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+ All rights reserved.This study conducted in Bangladesh reports the relationship of clinical complications with nutritional status and the prevalence of leukopenia among arsenic exposed patients living in the rural villages. A total of 115 exposed individuals diagnosed as arsenicosis patients were randomly selected from four known arsenic endemic villages, and age-matched 120 unexposed subjects were enrolled in the study program. The duration of arsenic exposure in about 37% of the patients was at least 10 yrs, while the population mean and range were 7.6 ± 5.2 yrs, and 1 – 25 yrs, respectively. The mean arsenic concentrations in the drinking water for the exposed and unexposed (control) population were 218.1 μg/L and 11.3 μg/L, respectively. The spot urine sample of the arsenicosis patients contained an average of 234.6 μg/L arsenic. Although very few patients showed elevated WBC count, 16% had leukopenia (below normal count), and the whole population had significantly low WBC count than the control subjects. Prevalences of neutropenia and lymphocytosis were observed in patients with chronic exposure to high levels of arsenic in water. The body mass index was found to be lower than 18.5, the cut-off point for malnutrition (underweight), in about 28% of the arsenicosis cases compared to 15% of the controls. The monthly income and total calorie consumption per day showed the patients were underprivileged than the controls. Arsenical symptoms and complications were more severe in the nutritionally vulnerable (underweight) patients than the overweight ones. Also, the incidences of leukopenia and anaemia were more common in the female patients who were underweight. The findings of this research demonstrate that the poor nutritional status of patients increases the complications of chronic arsenic toxicity; suggest the possibility of other sources of arsenic contamination different from drinking water in the study area; and establish a higher prevalence of leukopenia and lymphocytosis in arsenicosis patients.Endemic arsenic exposure emerged as a single catastrophe affecting millions of people mostly living in Bangladesh, India, Mexico, Taiwan and South America. In these regions, the concentrations of arsenic amount to several hundred micrograms per liter that considerably exceed the standard of 50 μg/L for drinking water, recommended by the World Health Organization. In Bangladesh, at least 25 million people are drinking arsenic contaminated water [1]. There is strong evidence from epidemiological studies of an association between chronic exposure to inorganic arsenic and hyperpigmentation, hyperkeratosis, and neoplasia in the skin as well as other diseases [2]. A higher prevalence rate of arsenical skin lesions with clear dose-response relationship has been found among Bangladeshi populations drinking arsenic contaminated well water [3]; and callus-like growths all over the extremities with changes in skin pigmentation have been reported [4].In humans, the liver rapidly detoxifies inorganic arsenic that is consumed in drinking water by transforming it to organic forms called monomethylarsonic acid (MMA) and dimethylarsenic acid (DMA) that are rapidly excreted in the urine to give an over-all arsenic half-life in the body of about 30 hours [5]. Thus, following exposure to arsenate (valency state v), the first step in the biotransformation is the reduction to arsenite (valence state iii); a process that may be considered a bioactivation [6]. There is evidence that the methylating capacity differs among individuals and populations and that different capacities would result in variation in tissue retention of arsenic. Environmental factors, particularly diet, might be important in explaining susceptibility to arsenic toxicity [7].Several studies have found that anaemia, leukopenia and thrombocytopenia are common effects of arsenic poisoning in humans following acute [8] and chronic oral exposures [9] at doses of 50 μg/Kg/day or more. These effects may be due to both a direct, cytotoxic or hemolytic effect on the blood cells [8,10] and a suppression of erythropoiesis [10]. The magnitude of exposure to patients examined by these researchers is not specified, although arsenic concentration in some of the wells in the area exceeded 1000 μg/L. However, there are reports that hematological effects are not observed in all cases of acute poisoning with arsenic [11,12], and the hematological abnormalities are reversible within weeks of termination of exposure [13].Patients with arsenic induced Bowen’s disease showed a defected cell mediated immunity and decreased percentages of T cell and T helper cell populations in the peripheral blood mononuclear cells, and arsenic exposure leads to diminution of phytohemagglutinin (PHA) stimulated T cell proliferation [14,15]. Electron microscopy study revealed that exposure to sodium arsenite caused alteration in the cytoskeleton, Golgi apparatus, mitochondria and perinuclear membrane of the T lymphocytes which ultimately changed intracellular secretion of proteins including IL���2 that might lead to an impaired proliferation of the cells when stimulated with PHA. It has been reported that both arsenite and arsenate are strongly toxic to macrophages and able to decrease the number of surviving cells to 50% at a concentration of 5 or 500 μM, respectively [16].This study was undertaken to determine the relationship of clinical complications of arsenic toxicity with nutritional status based on both measured and self-reported parameters in a randomly selected group of arsenicosis patients and on the profile of their peripheral blood leukocyte. The study area of this investigation was the northwestern district of Chapainowabganj where arsenic contamination in drinking water was first detected in Bangladesh in 1993. Samples were collected during November 2001 to August 2003 from patients living in the arsenic endemic rural villages of Rajarampur, Achinpara, Chandnai, and Bottola.Arsenic exposed subjects whose drinking water contained more than the maximum permissible limit of arsenic (50 μg/L, as recommended by WHO), were enrolled in this study. Their source of drinking water was from tube wells or artisan wells. Prevalence of arsenicosis was based on appearance of skin lesions among the enrolled patients. After informed consent, the researchers with the help of a public health nurse interviewed all the patients personally. The information on symptoms and complications of arsenic toxicity, anthropometric parameters (including age, height, and body weight), monthly income of the family, number of family members, housing, education, and food intake were recorded on a prepared questionnaire. The quality and quantity of food consumed by the patient on three consecutive days starting from the day before sample collection through the following day (from perspective list) were recorded. Since most of the patients were subsistence farmers, land-less farmers, day laborers, and their family members who could not include the monetary value of the produce consumed within the household, the monthly income of the family of all subjects was recorded solely on the basis of self-reported estimate. From the data on questionnaire, the socioeconomic and nutritional status of each subject was determined.Blood samples (2 – 3 ml) were collected from the subjects with their full consent to participate in this study. A total of 235 blood samples were collected of which 115 were from patients with arsenic toxicity (arsenicosis), 80 from age-matched subjects living in the same area as the patients but drinking safe water and with no sign of arsenic toxicity, and the remaining from unexposed subjects living in the city. Spot urine and drinking water samples were also collected from each individual. The fresh blood sample was used for blood type determination, total and differential counts of white blood cells.Arsenic content in the water and urine samples was measured at the Laboratory of Analytical Chemistry, School of Environmental Studies, University of Jadavpur, Kolkata, India. Some of the samples were also analyzed at the Analytical Research Division, Dhaka Laboratories, Bangladesh Council of Scientific and Industrial Research (BCSIR) for the content of arsenic. A Flow Injection - Hydride Generation - Atomic Absorption Spectrophotometer (FI-HG-AAS) was used for analysis of the water and urine samples.The total count of white blood cells (WBC) in a healthy subject usually varies from 4.0 – 11.0 X 106 cells/mL. To find out the effect of high concentrations of arsenic in drinking water on cells of the immune system, the WBC and differential counts were performed for each blood sample. The standard procedures were followed in preparing the samples, and total WBC and differential counts were determined. The percentages of neutrophils, eosinophils, basophils, lymphocytes and monocytes were calculated.Data analyses were carried out using the Statistical Package for Social Sciences (version 10.0 for Windows, SPSS Inc., Chicago, USA). The methods used were independent t– test for comparison of two groups (control and patients), correlations and statistical reporting. The results were considered significant when p was ≤ 0.05. Mean ± S.D. values were calculated for each studied parameter.Of the total 115 patients enrolled in this study, there were 36 males and 79 females with age ranging from 14 to 85 years. The age range of males was 14 to 85 years; while that of females was 14 to 75 years. The whole population of patients had an average monthly income of the family of Tk. 4,620 (about $ 80) with a median value of Tk. 4000 while the monthly income of the family varied from Tk. 1,000 - 16,000. A total of 41 patients had no formal education; while 36 had primary; and 27 had high school education. Only 11 patients had passed secondary school certificate examination, of which 2 completed graduation and another one had a master’s degree. The number of family members of the patients varied from 2 to 9 with an average of 5.02 ± 1.53 and median value of 5. A total of 39 patients lived in semi-buildings, 54 in tin-shades, 16 in mud-made houses, and 6 in huts. Of the total 120 unexposed control subjects, there were 81 males and 39 females. Six of them were illiterate, while the education levels of the rest were in the secondary to higher secondary and above. The monthly income of the families of the controls varied from Taka 1,500 – 25,000 with an average of Taka 6,730. Their family members varied from 2 to 6.The clinical symptoms based on skin manifestations, complications and duration of arsenic toxicity in the patients are shown in Table 1. There were diffused and spotted melanoses with black and white appearances, rough and mottled skin, keratosis or hardening of the skin with often formation of nodules. Spotted melanosis was more often seen on the throat, chest, back, or limbs. Many of the patients suffered from severe skin irritation. Prolonged exposure caused the skin to become rough and thickened due to diffused keratosis that usually developed on the soles of feet and palms of both hands with occasional formation of cracks on these areas. Patients with these clinical features were the worst sufferers of arsenic toxicity as they were unable to do their household works.Arsenicosis patients suffered from other complications like breathing problems including asthma, bronchitis and cough. Gastric and abdominal pain was also very common. Some patients had pain all over the body, while others had backache, headache, palpitation, anaemia and weakness. Anaemia was more prevalent in the females as 26 out of 30 patients identified with anaemia were females. The patients had pain in joints including the knee, burning sensation, tingling, frequent fever and cold. There were some less common complains like sleep disturbances, night blindness, depression, loss of appetite, edema on feet, goiter, tonsillitis, loss of hearing, lesion in mouth, gout and hepatomegaly. These symptoms were more common in the patients than in the control (unexposed) subjects living in the same villages.Depending upon the clinical symptoms in the patients, arsenicosis was divided into four categories: (i) melanosis (black and white pigmentation on skin), (ii) melanosis, keratosis, rough and mottled skin, (iii) melanosis, keratosis on palms and soles with and without cracks, nodule formation, and (iv) melanosis, keratosis, severe skin irritation, lump formation on feet. It was found that development of these symptoms was related to the duration of exposure as the mean duration was: 5.6 ± 4.1 yrs for (i); 7.2 ± 4.7 yrs for (ii); 8.6 ± 5.4 yrs for (iii) and 9.7 ± 6.6 yrs for (iv) (Table 1). Further, the duration of exposure in 42 of the patients was at least for 10 yrs, while 15 had the signs of poisoning developed for 2 yrs or so. The whole population of patients had an average duration of exposure for 7.6 ± 5.2 yrs with a median of 6.5 yrs, and varied from 1 – 25 yrs. The average levels of arsenic in the drinking water of the different categories of arsenicosis patients were: (i) 206 μg/L, (ii) 190 μg/L, (iii) 252 μg/L, and (iv) 271 μg/L.It was found that a total of 13 (11%) patients drank water that contained more than 600 μg/L of arsenic. There were 21 (18%) patients included in this study who switched to safe drinking water following the onset of arsenical skin lesions. The average level of arsenic in the drinking water of the patients was 218.1 ± 218.4 μg/L, with a median value of 156 μg/L. The levels varied from 3.0 – 875.0 μg/L. On the other hand, the mean arsenic concentration in the drinking water of the unexposed (control) population was 11.3 μg/L. The urine samples of only 15 (13%) patients contained arsenic that was within the permissible limit (5 – 40 μg/L). A total of 48 (42%) patients excreted more arsenic in urine than they had consumed through drinking water. This clearly indicated that there could be more than one source of arsenic contamination than drinking water. The average level of arsenic in the urine of the patients was 234.6 ± 311.5 μg/L while the median value was 129.7 μg/L and the levels varied from 20.0 – 1764.0 μg/L. Statistical analysis showed the levels of arsenic in the drinking water and urine samples of the patients was positively correlated at 1% level.Clinical symptoms and complications in arsenicosis patients in relation to the duration of arsenic exposureThe quality and quantity of food intake, calorie consumption, and anthropometric data of the patients showed that most of them were of poor health. The patients usually took rice with vegetables, potatoes and/or pulse everyday. They could manage to eat meat / fish once or twice a week. However, about 44% of the patients could eat meat / fish every day. The mean calorie consumption values were 1980 Kcal/day and 1850 Kcal/day for male and female patients, respectively. On the other hand, the quality and quantity of food intake by the control subjects were slightly improved and the mean calorie consumption values were 2100 Kcal/day and 1960 Kcal/day for males and females, respectively. The male patients had an average height of 1.67 meters (range: 1.47 – 1.85) and body weight of 54.4 Kg (range: 36.0–70.0); while the females had an average height of 1.49 meters (range: 1.27 – 1.82) and body weight of 48.3 Kg (range: 36.0 – 68.0). The control males had an average height of 1.70 meters (range: 1.55 – 1.82), body weight of 63.2 ± 8.0 Kg (range: 47 – 86) and body mass index (BMI) of 22.0 ± 2.3. The control females had an average height of 1.60 meters (range: 1.50– 1.68), body weight of 54.1 ± 7.8 Kg (range: 44 – 67) and BMI of 21.2 ± 2.5.The nutritional profile of the arsenicosis patients was evaluated based on their BMI. The male patients had an average BMI of 20.4 (range: 15.7 – 26.4), and that of the female patients was 21.4 (range: 16.2 – 31.5). The male and female patients were further divided into underweight (malnourished, BMI: < 18.5), normal weight (BMI: 18.5 – 24.9), and overweight (BMI: 25.0 – above) and their nutritional statuses were determined. The results are shown in Table 2. It was found that about 28% of the patients were underweight (malnourished) compared to 15% of the controls. The results showed kilocalorie consumed per kilogram body weight (Kcal/Kg) was highly significantly different between the female patients of different nutritional groups (p < 0.0001); but was significantly different only between the underweight and overweight male patients (p < 0.05). On the other hand, nutritional status of the male patients was directly related to the monthly income of the family; whereas data of the female patients did not show this correlation.To find out whether nutritional status played any role in the development of the toxic effects of arsenic, incidences of the different categories of arsenicosis were compared between the underweight, normal weight, and overweight patients (Table 3). The clinical symptoms that were evaluated were as described in the text and Table 1. The results showed that the severe symptoms of types (iii) and (iv) were absent in the overweight (well nourished) patients; whereas those were present in 9 out of 32 (about 28%) of the underweight (malnourished) patients. Further, some of the overweight patients suffered from toxicity for more than 10 years; still their clinical symptoms were not severe.The blood type data of the patients showed that 35% had O group, 31% had A group, 24% had B group, and 10% had AB group. On the other hand, the control subjects showed 27% had O group, 25% had A group, 37% had B group, and 11% had AB group. The blood group data showed O and A type blood were more common among the arsenicosis patients (pooled: 66%) than the unexposed controls (pooled: 52%). The total counts of the WBC in the patients and control subjects are shown in Fig. 1. It was found that the total WBC count in about 26% of the patients was ≤ 5 X 106 cells/mL, while 16% had leukopenia. Although leukocytosis was observed in about 4% of the cases, the whole population of patients showed an average WBC count of 6.71 ± 2.51 X 106 cells/mL compared to 7.34 ± 1.70 X 106 cells/mL in the control subjects. Statistical analysis showed the total WBC count in the arsenic patients was significantly lower than the control, unexposed subjects.Assessment of WBC counts in relation to nutritional statuses of the patients is shown in Table 4. The results showed that nutritional status was directly related to the WBC counts in the female patients who had low WBC counts than the males within the same nutritional group. Of the total 18 patients identified with leukopenia, about 78% of them were females (not shown in Table). Further, about 44% of the total patients with leukopenia belonged to the underweight nutritional group.The differential counts of WBC in the patients and control subjects are shown in Table 5. It was observed that about 26% of the patients showed neutropenia, having less than 50% neutrophils in the peripheral blood (normal range: 50 - 70%). This low neutrophil count was not compensated by elevated eosinophil or basophil counts. Further, the lymphocyte count showed high values in many patients (normal range: 20 - 40%); and 45 of them had more than 40% lymphocytes, of which 13 had ≥ 50% of the leukocytes belonged to this type. However, the whole population data showed none of the leukocytes in the patients varied significantly compared to that in the controls. The average neutrophil count in the neutropenic patients was 44.8 ± 3.7%, lymphocyte count 48.7 ± 4.4%, levels of arsenic in water and urine samples of 236.9 ± 248.3 μg/L and 195.9 ± 197.7μg/L respectively, and duration of arsenic exposure for 8.4 ± 5.9 yrs. These findings suggest that chronic exposure to high levels of arsenic in drinking water may cause neutropenia and lymphocytosis.This study has been conducted in an attempt to quantify whether nutritional status plays any role in the prevalence of chronic arsenic toxicity. Almost every physiological system in the human including the respiratory, immune, gastrointestinal, genitourinary, reproductive, and nervous systems have been reported to be adversely affected by prolonged exposure to high doses of arsenic [7,17,18]. There has been a significant dose-response relationship between skin cancer prevalence and chronic arsenic exposure as indexed by duration of consumption of high-arsenic artesian well water, average exposure, and cumulative exposure among residents of Taiwanese villages [19]. In these people, under nourishment due to a high consumption of dried sweet potato as a staple food has been found to be significantly associated with an increased prevalence of arsenic induced skin cancer. An epidemiological study conducted in West Bengal, India showed male gender and malnutrition correlated with increased prevalence of skin manifestations in arsenic exposed population [20]. In a recent study, an association between nutritional status and chronic exposure to arsenic has been reported among Bangladeshi patients [21].Nutritional status of arsenicosis patientsRelationship between clinical symptoms of arsenic toxicity and nutritional status of patientsComparison of the total white blood cell (WBC) counts in control subjects (n = 74) and arsenic patients (n = 115). Leukopenia (< 4 million cells/mL) had been found in 18 (16%) of the patients suffering from arsenic toxicity. The average WBC count in the patients was 6.71 ± 2.51 X 106 cells/mL compared to 7.34 ± 1.70 X 106 cells/mL in the controls, suggesting the patients had significantly lower WBC counts than the control subjects (p = < 0.05).Assessment of WBC counts in relation to the nutritional status of patientsDifferential counts of WBC in the arsenic patients and control subjectsIn this study, skin manifestations have been found as the prime and common features of arsenic toxicity that has been considered as definite exposure [22]. We found a significant direct relationship between the concentrations of arsenic in water to that in urine. The duration of exposure has been found directly related to the severity of clinical disease. Similar dose-response relationships have been reported in long-term arsenic exposure with ischemic heart disease mortality in Taiwan [23], and with skin lesions in Bangladesh [3].The calorie consumption per day values indicated that the patients have been nutritionally compromised and represented the common lower-middle class Bangladeshi populations; the control subjects from the same area showed similar data but have been in relatively better shape. BMI of the patients has been found inversely related to the duration of toxicity. Assessment of nutritional status in terms of BMI, calorie consumption per day and body weight has shown to have an inverse relationship to the clinical symptoms of arsenic toxicity. Our observations suggested that the underweight (malnourished) patients have been more prone than the overweight ones in developing the severe clinical symptoms of arsenicosis. Although BMI of the male patients has been lower than the females, we have not found an association of male gender with increased prevalence of arsenic-related clinical complications as has been reported [3].About 60% of our patients suffered from respiratory problems including asthma, bronchitis and cough. Prevalence of respiratory effects has been reported in a study conducted in West Bengal, India [24], and in Bangladesh [22]. In our study, some patients have had more than one arsenic-related complication as about 59% of them reported of having gastric and abdominal pain. About one third of the patients complained of weakness, headache and palpitation. These observations lead to the possibility that the respiratory and gastrointestinal systems are preferentially selected than other organs in human arsenic toxicity. Another observation of this study has been that since the study area located in the arsenic endemic region where the farmers use ground water for irrigation, and high concentration of arsenic present in these water might increase arsenic content of the paddy fields and ultimately contaminate nearby ponds, canals and other water bodies from where the residents collect water for cooking. Thus cooking rice and other foods in contaminated water might explain the reason for about 42% of the patients excreting more arsenic in urine than they have consumed through drinking water.The effect of inorganic arsenic on human erythrocyte morphology has been studied in an in vitro model and found changes in membrane integrity and deformability that could contribute to micro vascular occlusion and related peripheral vascular effects in chronic arsenic exposure [25]. Rodent models of arsenic exposure have also demonstrated disturbances in heme biosynthesis, characterized particularly by an increased urinary uroporphyrin excretion [26,27]. These observations explain the possible reason for 26% of our patients suffering from anaemia, also reported by previous workers [8,9]. However, we have seen a far higher prevalence of anaemia in the female patients than males, 26 out of 30; 14 of these female patients represented the underweight nutritional group. Accounts of leukopenia due to neutropenia or lymphopenia, and relative eosinophilia have been reported in some of the early studies [28,29]. However, we have seen neutropenia among 26% of the patients, incidences of lymphopenia and eosinophilia have been insignificant; instead, we report lymphocytosis in 40% of the patients. Neutropenia and lymphocytosis have been found associated with chronic exposure to high levels of arsenic in drinking water. Conditions of neutropenia may indicate persistent infection in the patients, possibly due to arsenical skin lesions. Further, arsenic-toxicity induced lymphocytosis may modulate the immune response of the patients.The majority of our patients have come from very poor socioeconomic class with about 56% unable to eat animal protein (meat / fish) everyday. Poor nutritional status of the patients may increase arsenic retention in the body and thus lead to severity of clinical complications. Patients with protein-energy malnutrition have inadequate supply of methionine from dietary sources, and studies in rabbits have shown low amount of methionine or protein in the diet decreased methylation of inorganic arsenic [30]; and deficiency of other dietary trace elements including zinc and selenium [31] may lead to arsenic accumulation and contribute to the toxic effects in the body. A recent study has shown that low intake of calcium, animal protein, folate and dietary fiber may increase susceptibility to arsenic-caused skin lesions [32]. Moreover, apart from the poor nutritional status, peoples in Bangladesh have greater daily intake of water (and thus consume more arsenic) than those living in other parts of the world that are not tropical, and thereby suffer from the adverse effects of arsenic toxicity.This work was supported by research grants from the Ministry of Science and Information & Communication Technology, Government of the People’s Republic of Bangladesh. We thank Professor Dipankar Chakraborti and Dr. Uttam K. Chowdhury, School of Environmental Studies, Jadavpur University, Kolkata, India, and Professor M. Amjad Hossain, Chairman, BCSIR, Dhaka, for helping us in analyzing the levels of arsenic in the water and urine samples; Dr. M. M. Hoque Bakul, Medical Officer, and Ms. Nasima Begum of Chapainawabganj; Mr. A. K. M. Mahbub Hasan and Mr. M. Sohel Shamsuzzaman of this department for their cooperation; and all the participants of this study.
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+ All rights reserved.Arsenical keratosis and skin cancer are among the most common health effects associated with acute and chronic exposures to arsenic. This study examines the acute and chronic dose-responses of arsenic in established human cell lines using keratinocytes (HaCaT), melanocytes (CRL1675) and dendritic cells (THP-1 + A23187). Chronic conditions were established by treating the three cell lines with at least 8 passages in 0.2 µg/mL arsenic trioxide. Cytotoxicity was assessed using the fluorescein diacetate assay after 72 hrs of exposure. Single cell gel electrophoresis (Comet assay) was used to measure DNA damage. Acute exposure to arsenic had LD10 and LD25 values of 0.38 µg/mL and 3.0 µg/mL for keratinocytes; 0.19 µg/mL and 0.38 µg/mL for melanocytes; and 0.38 µg/mL and 0.75 µg/mL for dendritic cells. Cytotoxicity assays for chronically exposed cells resulted in LD10, and LD25 values of 0.4 µg/mL and 0.8 µg/mL for keratinocytes; 0.10 µg/mL and 0.20 µg/mL for melanocytes; and 0.10 µg/mL and 1.0 µg/mL for dendritic cells. The Comet assay showed that arsenic was highly genotoxic to the three cell lines. No significant differences (p > 0.05) in DNA cleavage were observed between acute and chronic exposures. In acute exposure arsenic genotoxicity was more severe with dendritic cells while melanocytes were more sensitive to arsenic cytotoxicity. Similarly, chronically exposed dendritic cells showed the maximum genotoxic damage while melanocytes were more sensitive to arsenic cytotoxicity. In conclusion, this research shows that arsenic is dermatotoxic, showing a high degree of genotoxicity and cytotoxicity to skin cells.Arsenic exposure has been associated with skin keratosis and the development of many cancers, especially of the skin, lung and bladder, prostate, kidney and liver [1]. Low dose ingestion of arsenic does not immediate fatal consequences, however, prolonged arsenic exposure have been shown to significantly increase the risk of contracting these various forms of cancer. Because of increased risk of cancer associated with inorganic arsenic, the United States Environmental Protection Agency (U.S. EPA) has classified inorganic arsenic as a class A (known) human carcinogen [2].Following long-term exposure to arsenic, the first changes usually observed in the skin include pigmentation changes and then hyperkeratosis. Changes in pigmentation of skin are related to alterations brought about in the components of the epidermal-melanin unit. Chronic exposure to arsenic frequently results in skin cancers [3,4]. The skin is made up of two main layers, epidermis and dermis. The epidermis, outermost layer, provides the first barrier of protection from the invasion of foreign substances into the body. The principal epithelial cells of the epidermis are the keratinocytes. The dermis contains the melanocytes that migrate to the basal layer of the epidermis and reside there. Epidermis and surface epithelium dendritic cells are made up of immature cells known as Langerhan cells [5]. The dendritic cells arise in the bone marrow and migrate to and seed tissues throughout the body including the epidermis. The ratio for melanocytes: basal keratinocytes is 1: 10 [6] and the ratio of dendritic cells to keratinocytes is 1:10 [7].Arsenic has been found to be genotoxic [8]. Hei suggested that arsenic acts through a series of chemical reactions in the cell, interacting strongly with nearby molecules, and changing the structure of cellular components such as DNA [9]. Other studies have found that exposure to inorganic arsenic increases the frequency of micronuclei, chromosome aberrations and sister chromatid exchanges both in humans and experimental animals [10,11]. Arsenic has been previously shown not to affect DNA directly, but to intensify the toxic effects of other physical and chemical agents by inhibiting DNA repair, changing cell redox potential, and altering DNA methylation of cell-cycle control proteins [8]. In this study we evaluated the genotoxic and cytotoxic effects of arsenic in established human cell lines represented by keratinocytes (HaCaT), melanocytes (CRL1675), and dendritic cells (THP-1+A23187) [12], following acute and chronic exposures.Arsenic trioxide with 99.9% purity was purchased from Fisher Scientific (Suwanee, GA) and used throughout the experiments without further purifications. Reagents were purchased from Gibco (Grand Island, NY). Fetal bovine serum (FBS) was obtained from Hyclone Laboratories (Logan, UT). HaCaT cell line was obtained from Dr. N. Fusenig (Division of in vitro Differentiation and Carcinogenesis, German Cancer Research Center, Germany). THP-1 and melanocytes (CRL1675) were obtained from American Type Culture Collection (Rockville, MD). The cells were cultured in a humidified atmosphere with 5% CO2 at 37 o C.The standard growth medium was prepared according to recommendations for specific cell lines including Dulbecco’s Minimum Essential Medium for HaCaT, RPMI 1640 for THP-1 (dendritic cells), and Vitacell medium for melanocytes (CRL 1675). Media were considered complete with the addition of 10% FBS and 1% antibiotic (penicillin and streptomyocin).Cytotoxicity assay was carried out as previously described in our laboratory [13]. Briefly, cells were counted (20,000cells/well) and resuspended in complete medium. Aliquots of 100µl of cell suspension were placed in wells of microtiter plates, and 100µL of different concentrations of arsenic trioxide (0 to 200µg/mL) were used to treat the cells. The plates were incubated for 24, 48, and 72 hours, respectively. After incubation, cells were centrifuged, washed twice with PBS, PBS discarded and aliquots of 100µL of fluorescein diacetate (10ng/mL) added. The plates were incubated 35 min before being read using a Fluroskan II microplate reader (Helsinki, Finland) with an excitation wavelength of 485 nm, and an emission wavelength of 538 nm.In preliminary studies, several different doses of arsenic (< LD10) were used to grow cells by trial and error. After 3-4 passages, we encountered very sluggish growth with eventual cell death due to arsenic toxicity. In our experience with HaCaT, we found 0.2ppm to be well tolerated. Further literature research showed that 0.5 and 1.0µM (0.10-0.2 ppm) could be used in HaCaT for 5 months [4]. Thus, we arbitrarily chose 0.2ppm as our working dose to represent chronic exposure. The use of THP-1 +A23187 to mimic dendritic cells [12] was due to the fact that dendritic cells are found in the surface epithelium along with keratinocytes and melanocytes.Cells were counted (10,000 cells/well) and re-suspended in media with 10% FBS. Aliquots of 100µL of the cell suspension were placed in 96 well plates, treated with arsenic trioxide concentrations at doses of LD10 and LD25 determined from the cytotoxicity assay data, and incubated in a 5% CO2 at 37oC for 72 hrs. After incubation, the cells were centrifuged, washed with PBS, and re-suspended in 100 µL PBS. In a 2 mL tube, 20 µL of the cell suspension and 200 µL of melted agarose were mixed and 75µL pipetted onto a pre-warmed slide. The slides were placed in a refrigerator at 4° C for 10-20 min and placed in chilled lysis buffer for 45 min. Slides were washed twice for 5 min with TBE and electrophoresed in a horizontal gel apparatus at 25V for 10 min. Slides were placed in 70% ethanol for 10 min, removed, tapped to remove excess ethanol, and placed in an alkaline solution containing 99mL H2O, 100µL of 0.1mM Na2EDTA and 1g NaOH for 45 min. Slides were air dried for 2.5 hrs, stained with SYBR Green and allowed to set 4 hrs. The slides were viewed with an Olympus fluorescence microscope and analyzed using LAI’s Comet Assay Analysis System software (Loates Associates, Inc. Westminster, MD).Cell mortality data recorded from the cytotoxicity assay were plotted against arsenic trioxide concentrations and a linear regression analysis was performed to determine and characterize the dose-response relationship equation. This equation was then used to determine the LD10 and LD25 values used in subsequent genotoxicity experiments with the comet assay. For comet assay, photographs were taken to illustrate the changes in DNA morphology associated with arsenic exposure. The comet data for DNA fragmentation and tail length were expressed as means ± SDs with n = 70, and F-statistic ANOVA was applied to determine if there were significant differences in genotoxicity with regard to arsenic treatment and cell type. Differences were considered at p value ≤ 0.05.Cytotoxicity data in terms of LD10 and LD25 for acutely and chronically exposed cells are shown in Table 1. In acute experiments for cytoxicity, the LD10 for melanocytes was lower than that of keratinocytes and dendritic cells. LD25 dose for keratinocytes was the highest while that of dendritic cells was intermediate, andthat of melanocytes was the lowest. In chronic exposure for cytoxicity assay melanocytes and dendritic cells are more sensitive at LD10 than keratinocytes while melanocytes are more sensitive than keratinocytes and dendritic cells at LD25.Cytotoxicity (LD10 and LD25) of arsenic trioxide in acute and chronically-exposed skin cells.The comet assay or single cell gel electrophoresis revealed that treatment of the cells with arsenic causes severe damage to the cells’ nuclear DNA. Figure 1 is the representative picture for all the three cell types. As shown in this figure, the nuclear DNA of untreated cells is perfectly round, but the nuclear DNA of arsenic treated cells is severely fragmented. There are several ways to measure the severity of DNA fragmentation. Here listed are the percent of DNA fragmentation (percent of DNA in the Comet tail versus total DNA), and the length of the comet tail. The higher the percent of DNA fragments, the more severe is the damage. Similarly, the longer the comet tail, the smaller is the DNA fragment, the more severe is the damage.Figure 2 depicts the percentages of DNA fragmentation, and the lengths of comet tail in keratinocytes, melanocytes, and dendritic cells acutely (upper graphs), and chronically (lower graphs) exposed to arsenic trioxide at LD10 and LD25 doses. Damage to the nuclear DNA of melanocytes and dendritic cells as measured by both percent DNA damage and length comet tail of DNA fragments is similar for both LD10 and LD25, but significantly (p < 0.01) more severe than that of the keratinocytes (Figure 2).Results showed when compared to controls that there was no significant difference in tail length of acutely exposed keratinocytes at LD10, however, there was statistically significant differences (p < 0.001) for the LD25 exposure. In chronically exposed keratinocytes, there were significant differences at all exposure levels (p < 0.001) for tail length. The percentage of DNA damage and the length of the comet tail were up in LD10 dose but did not reach statistical significance, while DNA damage and tail length were up and reached statistical significance in LD25 (p < 0.001) when compared to controls. In the acutely exposed melanocytes, LD10 tail length increased slightly more than LD25 when compared to controls but neither reached statistical significance. In chronically exposed melanocytes, tail length for LD10 decreased slightly while LD25 increased significantly (p < 0.001) when compared to controls. The percentage of DNA damage in melanocytes with acute exposure was slightly increased at LD10 and more at LD25 when compared to controls but showed no statistical significant differences. Chronically exposed melanocytes showed %DNA damage to be significant at LD25 only.At acute exposures dendritic cells showed tail length to be slightly greater at LD10 than LD25 when compared to controls, but the difference in lengths were statistically significant at both LD10 and LD25 (p < 0.001). In chronically exposed dendritic cells, tail length increased at LD10 and more at LD25 than the controls, and reached statistical significance (p < 0.001). DNA damages in dendritic cells acutely exposed at LD10 and LD25 were greater than controls but did not reach statistical significance. Damages in chronically exposed dendritic cells at both LD10 and LD25 were greater than the controls and were significantly different (p < 0.001) (Figure 2).Human exposure to arsenic, a ubiquitous and toxic environmental pollutant, is associated with an increased incidence of skin cancer. It is a carcinogen that poses a significant health risk in humans. The mechanisms associated with arsenic-mediated toxicity, DNA damage and proliferation at low chronic levels of exposure remain to be examined in depth. Cell cytotoxicity measures the capacity of the intact cell to recover from the damage induced and thus forms the basis for measuring the sensitivity of the cell in question. Several studies have addressed cytotoxicity to keratinocytes using different forms of arsenic [14,15,16] but not arsenic trioxide.Comet assay depicting the genotoxic effect of arsenic (LD10) to melanocytes, dendritic cells and keratinocytes acutely and chronically exposed. The bottom pictures are two control keratinocytes wile the upper figures represent keratinocytes, dendritic cells and melanocytes exposed to arsenic trioxide at the dose levels of LD10. Y-axis represents total height while X-axis gives total length of the cometPercentages of DNA fragmentation and lengths of comet tail in acutely and chronically exposed keratinocytes (K), melanocytes (M), and dendritic cells (D) at LD10 and LD25.Arsenic toxicity studies on melanocytes and dendritic cells are also lacking in the literature. Our studies found the acute LD25 to be 3ppm for HaCaT and the chronic LD25 to be 0.8ppm. LD25 values of 0.38ppm and 0.25ppm in acute and chronic exposures were found for melanocytes, while values of 0.75ppm and 1.0ppm were recorded in dendritic cells, respectively (Table 1). These data indicate that arsenic is cytotoxic to the three tested skin cells, and that melanocytes appear to be more sensitive to arsenic toxicity while keratinocytes are more tolerant at both acute and chronic conditions. For the most part these data indicate that chronically-exposed cells are more sensitive to arsenic toxicity than in acute exposure with the exception of keratinocytes at LD10 and dendritic cells at LD25.Human activities have increased the possibility of exposure to naturally occurring metals causing a greater risk of exposure to toxic levels [17]. The exposure of metals such as arsenic constitutes a major health concern. Arsenite induces DNA damage referred to as genotoxicity in human cells within a pathologically meaningful dose range. Arsenic toxicity is cell specific; therefore, it is important that target cells be used for investigations [17]. Hamadeh et al. exposed normal human epidermal keratinocytes (NHEK) to nontoxic doses (0.005-5 µM) of arsenic (III) and that exposure simultaneously modulates DNA repair, and redox-related gene expression in NHEK [18]. Studies show that arsenic may not directly impact DNA but may inhibit some DNA repair [19]. Arsenic has been shown to induce DNA damage in human cells. Inorganic arsenic increases the frequency of micronuclei, chromosome aberrations and sister chromatid exchanges as well as inhibits DNA repair [10,11]. Specifically, a significant increase in comet tail-length at doses 0 to 6.45 mg/kg body weight demonstrated that arsenic trioxide cause DNA damage effectively [20,21]. Using the comet assay, we applied the alkaline treatment which aids in the unwinding and denaturation of DNA molecules, thus allowing for the sensitive detection of single-strand damage [22].In our studies using the three cell lines, we found for genotoxicity assay that in acute exposures there was a decrease in DNA fragmentation and tail length for keratiinocytes at LD25. This implies that probably a DNA repair mechanism may be activated thus affording protection to keratinocytes. Further, the damage to the nuclear DNA in melanocytes and dendritic cells was more severe than that of keratinocytes. In general, the DNA damage to chronically exposed keratinocytes was lower than dendritic cells and melanocytes. It can thus be conjectured that genotoxicity or cancer in the long term effect could be attributed to dendritic cells and melanocytes. When comparing dendritic cells and melanocytes one can further speculate that dendritic cells are playing a more important role as the damage to the dendritic cells is more severe than melanocytes. Thus, we have demonstrated that dendritic cells are more potent in genotoxicity implying that they may be the first trigger followed by melanocytes and then eventually affecting the key element keratinocytes in skin carcinoma.With respect to arsenic we propose that arsenic has a multifactorial effect on the cellular elements of the epidermis (Figure 3). According to our model it is postulated that the initial change caused by arsenic is on the dendritic cell at the DNA level. The antigen presenting cell then presents arsenic very effectively to the melanocyte which causes its cell death. If the death is partial then it triggers keratinocyte to alter its division and thus lead to carcinoma of the skin. If there is extensive death then the characteristic changes in the pigmentation of the skin occur.Our cytotoxicity data revealed that keratinoctes were more tolerant while dendritic cells were intermediate and melanocytes were most sensitive to arsenic toxicity. Pigmentation alteration due to long-term exposure of arsenic could be associated due to the direct cytoxicity of arsenic on melanocytes. This research is the first report investigating the in vitro effects of arsenic-induced cytotoxicity and DNA damage in melanocytes, dendritic cells and keratinocytes, concomitantly in the same study. It is anticipated that data from this report will serve as a base for furthering our knowledge on arsenic modulation of cytotoxicity and DNA damage in skin cells.A schematic representation of the damage caused by arsenic on the cellular elements of the skin involving keratinocytes, melanocytes and dendritic cells.This research was financially supported in part by a grant from the National Institutes of Health (Grant # 1G12RR13459) through the Center for Environmental Health, and in part by the U.S. Department of Education (Grant # PO31B990006) through the Title-III Graduate Education Program. The authors thank Jian Yan for the technical assistance.
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+ All rights reserved.Pentachlorophenol (PCP) is an organochlorine compound that has been widely used as a biocide in several industrial, agricultural, and domestic applications. Although it has been shown to induce systemic toxicity and carcinogenesis in several experimental studies, the literature is scarce regarding its toxic mechanisms of action at the cellular and molecular levels. Recent investigations in our laboratory have shown that PCP induces cytotoxicity and transcriptionally activates stress genes in human liver carcinoma (HepG2) cells [1]. In this research, we hypothesize that environmental exposure to PCP may trigger cytotoxic, mitogenic, and endocrine-disrupting activities in aquatic organisms including fish. To test this hypothesis, we carried out in vitro cultures of male channel catfish hepatocytes, and performed the fluorescein diacetate assay (FDA) to assess for cell viability, and the Western Blot analysis to assess for vitellogenin expression following exposure to PCP. Data obtained from FDA experiments indicated a strong dose-response relationship with respect to PCP cytotoxicity. Upon 48 hrs of exposure, the chemical dose required to cause 50% reduction in cell viability (LD50) was computed to be 1,987.0 ± 9.6 μg PCP/mL. The NOAEL and LOAEL were 62.5 ± 10.3 μg PCP/mL and 125.0±15.2 μg PCP/mL, respectively. At lower levels of exposure, PCP was found to be mitogenic, showing a strong dose- and time-dependent response with regard to cell proliferation. Western Blot analysis demonstrated the potential of PCP to cause endocrine-disrupting activity, as evidenced by the up regulation of the 125-kDa vitellogenin protein the hepatocytes of male channel catfish.Global production and use of organochlorine pesticides have resulted in ubiquitous residues in soil, water, and air samples. Exposure to this family of pesticides has been strongly connected to endocrine disruption activity, resulting in adverse human and wildlife developmental defects, disease, and potentially cancer [2,3]. Major groups of organochlorine pesticides include chlorinated benzenes, cyclodienes, cyclohexanes, and dichlorodiphenylethanes. Collectively, they have become key sources of environmental contamination because of landfill leachate, effluents from pesticide manufacturing plants, and exhaust gases from the combustion of woods [4]. Although some organochlorine pesticides such as hexachlorobenzene (HCH) and pentachlorophenol (PCP) are used as fungicides, most of them are primarily known for their effective biocidal action against a wide variety of insects and microorganisms. During the early 1940s, organochlorine pesticides were commercially manufactured in the United States and heavily used in agriculture for the control of insects. Widespread distribution of these pesticides peaked in the early 1960s because of extensive use in agricultural, industrial, and domestic applications. It has been found that parent organochlorine pesticides often contain impurities that are precursors of toxic metabolites. These by-products are considered to be more dangerous as well as persistent in the environment [5,6]. For example, PCP metabolizes to tetrachlorophenol, HCH, and various dioxins and furans. The isometric form, beta-HCH is more persistent than PCP. Subsequently, polychlorinated dibenzo-p-dioxin (PCDD) and polychlorinated dibenzofuran (PCDF) compounds are found in areas where PCP has been used or where PCP wastes have been disposed [2,7]. The slow biodegradation rate and lipophilic nature of organochlorine pesticides have contributed to pervasive contamination and persistence in the environment. For these reasons, the United States Environmental Protection Agency (U.S. EPA) banned many organochlorine pesticides in the 1970’s and 1980s. There is now accumulating evidence indicating that organochlorine pesticides can compromise the integrity of the male reproductive system of fish, reptiles and mammals [8,9,10]. Additionally, studies have shown that exposure to low levels of organochlorine residues can result in physiological disturbances and reproductive disruption in humans and other animals [3,11,12,13,14,15,16,17]. Organochlorine chemicals, including DDT, polychlorinated biphenyls (PCBs), dioxins, lindane, hexachlorobenzene, and PCP are generally cited as potential endocrine-disrupting substances [2,18,19]. Endocrine-disrupting mechanisms of these chemicals are ardently linked to mimicking hormonal estrogens [3,18,20]. Furthermore, chlorinated compounds, that mimic hormones, have the ability to induce proliferate responses [3]. Exposure to organochlorine pesticides in the aquatic environment has resulted in towering occurrences of vitellogenin (Vtg) synthesis in species of male fish [21,22,23]. Furthermore, the induction of Vtg in male fish exposed to environmental contaminants has become an attractive indicator of potential estrogenic potency [21,24,25,26]. PCP is a chlorophenol herbicide and prevalent wood preservative in the United States. It has been used worldwide in fungicide, insecticide, and herbicide applications. The biocidal action of PCP protects timber from fungal rot and wood-boring insects, thus extending the life of wood products. PCP is one of the most heavily used organochlorine pesticides in the United States, preceded by the herbicides atrazine, and alachlor [27]. It was first manufactured in 1841, and commercially produced in 1936 [28,29]. By 1967, sodium pentachlorophenate (Na-PCP) was used expansively in agricultural and industrial applications because pure-grade PCP was virtually insoluble in water [30,31]. Other applications of PCP include leather tanning, cooling-tower algae and fungi control, slime and fungus control in photographic solutions, and in other industrial activities for protection of plants and products from biological degradation [13,32,33]. PCP was banned as an herbicide, and over-the-counter sales were prohibited in 1987 [34]. In 1988, the U.S. EPA announced the restricted use of PCP in the pulp and paper industry where it is used in paper coatings, sizing, and adhesives and in inks. A decline in the use of PCP has resulted over years because of U.S. EPA regulations. Nonetheless, from 1987 to 1993, PCP releases to land and water totaled nearly 100,000 lbs of which about 80% was to land [34]. Various trade names for PCP include: Penchlorol, Dowicide, EP 30, Permagard, Permasan, Permatox DP-2, Fungifen, Dura Treet II, Glazd Penta, Woodtreat, PentaReady, PentaWR, Chem-tol, Cryptogil oil, Weedone, and Term-I-Trol [32,34]. PCP is an ominous pollutant whose environmental pathways lead to the contamination of soil, water, and food. Contamination occurs when the solution migrates from the interior of the wood to the exterior surface, thus leaching into the surrounding soil and groundwater. In addition, spills at industrial facilities and hazardous waste sites are key tributaries to soil contamination. Because of these pathways, drinking water, surface water, groundwater, rain, snow, air and aquatic biota have become common sources for PCP-contamination in the United States [35]. In aquatic ecosystems, levels of PCP are increased up to 10,000 times higher than the concentration in the surrounding waters [36]. PCP is acutely and chronically highly toxic to cold and warm water fish, and moderately toxic to other freshwater and marine organisms [37]. Bioaccumulation is a significant factor that affects several species of fish and invertebrates.PCP is a systemic toxicant to the nervous, immune, and reproductive systems and causes injury to hepatic and urinary organs [36,37]. PCP has been established as a carcinogen in animal laboratory studies, and a probable human carcinogen-Group B2 [38,39]. High doses of oral exposures and near lethal levels of PCP can disturb the developmental phase in animals. Series of experiments report that exposures to high levels of PCP are embryolethal in rats [40,41]. Also low doses of oral administration during the critical period of pregnancy in hamsters result in fetal deaths [40]. In addition, another study demonstrated that PCP-treated rams, from conception to necropsy at 28 weeks of age, increased in scrotal circumference, induced acute semi-niferous tubule atrophy, and reduced epididymal sperm density [11]. Another study characterized patterns of abnormal reproductive morphology when PCP exposure in female ewes caused an increase in severity of oviductal intraepithelial cysts [17]. Aggravation of the developmental phase in mink was reported when PCP caused a decrease in reproductive competence, whelping rate, and fertility [42]. More recently, another study reported that topical exposures to PCP in the Japanese medaka embryo increased embryos mortality and that the neurula stage was the most sensitive embryonic stage that was intolerant by a single topical exposure of PCP [43] Prolonged exposure conditions to PCP can result in adverse reproductive effects in humans that are associated with changes in the endocrine gland function, and immunological dysfunction [36]. Chronic exposure to PCP is often a widespread cause of reproductive toxicity in women. Epidemiological studies have confirmed that women with a history of spontaneous abortion, unexplained infertility, menstrual disorders, and mild ovarian deficiency have elevated levels of PCP in their blood [44].Recent investigations in our laboratory have shown that PCP triggers the induction of various genes in human liver carcinoma (HepG2) cells, and is involved in Phase I biotransformation, protein structure alteration, the inflammatory response, and DNA damage [1,45]. There are few reports in which PCP has been linked to endocrine-disrupting activity. In the present investigation, we hypothesize PCP-treatment of catfish hepatocyes may trigger cytotoxic, mitogenic, and endocrine-disrupting responses.Pentachlorophenol (C6Cl5OH, CAS No. 87-86-5, Lot No. 01530TS), with purity 98.0% was purchased from Aldrich Chem Co., (Milwaukee, WI). Dulbecco’s Modified Eagle’s Minimal Essential Medium (DMEM, Lot No. 109721) and tissue culture supplements were purchased from Life Technologies (Grand Island, New York). Twelve percent SDS-PAGE gels were obtained from ISC BioExpress (Kaysville, UT). c-fos primary monoclonal antibody, was purchased from Oncogene Research Products (San Diego, CA), while vitellogenin monoclonal anti-carp primary antibody was purchased from Biosense Laboratories (Bergen, Norway). Alkaline phosphatase conjugated donkey anti-goat IgG and goat-anti-mouse IgG secondary antibodies, and BCIP/NBT color development substrate were purchased from Promega (Madison, WI). Protein assay reagent was obtained from Xenometrix, Inc. (Boulder, CO). Accumax was purchased from Innovative Cell Technologies (San Diego, CA). Sexually mature male channel catfish, measuring 20 to 30 cm in length, were obtained from Grambling State University Aquaculture Facility, Grambling, LA. On day one, the catfish were transported to the laboratory. Subsequently, catfish were placed on ice for 10 minutes to numb incision area. An incision was made on the ventral side of the catfish where the liver was surgically removed. Hepatocyes were isolated by breaking up the liver into small chunks, several square millimeters. Accumax solution of 1.0 mL per milligram of tissue (w/v) was employed to dissociate hepatocytes from primary tissue. Tissue samples were incubated in Accumax on a platform rocker at 37oC up to 25 minutes. Following removal of cell debris, the hepatocytes were seeded into two 75 cm2 flasks containing 25 mL of fresh Dulbecco’s Minimum Essential Medium (DMEM) + 10% fetal bovine serum (FBS) + 1% Penicillin/Streptomycin, and incubated at 37oC in a 5% CO2 incubator. On day two, the medium was replaced with fresh +DMEM and hepatocytes were examined each day for cell density and morphology prior to treatment.The fluorescein diacetate (FDA) method was used to assess cell viability once primary cell cultures were established. FDA, an uncharged fluorescent dye, is rapidly esterified once it enters the cell. The viability of the cells is assessed from the hydrolysis product, fluorescein, which cannot escape from live cells. Fish hepatocytes grown to 80-95% confluency were washed with phosphate buffer saline (PBS), trypsinized with 5 mL of 0.25% (w/v) trypsin-0.03% (w/v) EDTA, diluted, and counted (2.5 – 5 x 105 cells/well). Next, cell suspensions of 160 μl DMEM (minus supplements) were seeded into a 96-well microtiter culture plate for 24 h at 37oC in a 5% CO2 incubator. The next day, 40 μl aliquots of PCP solutions at varying concentrations were added to each well. Cells were placed in the incubator for 48 h at 37oC in 5% CO2. Following 48 h period, medium mixture was removed and culture plate was washed once with 200 μl PBS/well. Aliquots of 100 μl FDA solution (10 μg/mL in PBS), were added column-wise to each well and incubated at 37oC for 30–60 minutes before reading the fluorescence. A Fluoroskan Ascent FL spectrofluorometer by Labsystems (Beverly, MA) at excitation and emission wavelengths of 485 and 538 nm, was used to measure absorbance.Primary catfish hepatocytes were grown to confluency in polystyrene multidish 6-well plates. About 1 x 106 cells per well were treated with duplicate 0, 750, 1500, and 2000 μg/mL concentrations of PCP for 48 hr. Untreated cells served as controls. After treatment with PCP, an equal volume (200 μL) of sample buffer (0.2 mol/L Tris, pH 6.8, 1% SDS, 30% glycerol, 7.5% β-mercaptoethanol, 0.1% bromophenol blue) was added to each well. Cells were mechanically dislodged, transferred to microcentrifuge tubes, and heated at 95oC for 10 min. Samples were then frozen until future use. The Bradford protein assay in a microtiter plate (Falcon 96-well microtiter plates) format was used for the determination of protein concentrations in samples. Bovine serum albumin (BSA) (1mg/mL) was dissolved in saline and used as the protein standard. Protein reagent (180 µl) was added to 21 wells of the microtiter plate. Triplicate volumes of 0, 2, 4, 6, 10, 15, and 20 µl of BSA (1mg/mL) were transferred into assigned wells. Distilled water was added to bring total volume to 200 µl per well. Without prior incubation, the total proteins for catfish cell lysates were quantitatively measured at 600 nm absorbance using the Multiskan Ascent microplate reader (Labsystems, Beverly, MA). Catfish hepatocytes were grown to confluency in polystyrene multidish 6-well plates. About 2 x 106 cells per well were treated with duplicate 0, 750, 1500, and 2000 μg/mL concentrations of PCP for 48 hr. Untreated cells served as controls. Whole cell extracts from catfish hepatocytes (20 μg/mL) containing an equal volume of sample buffer was heated at 100oC for 10 min and electrophoresed on a 12% SDS-polyacramide gel. Separated proteins were transferred onto a nitrocellulose membrane on ice in 20 mM Tris base, 150 mM glycine, 20% methanol, pH 8.0. Subsequently, membranes were blocked (10 mL of Tris-buffered saline 0.1 Tween-20 [TBST] with 5% nonfat dry milk) for 1 h at room temperature. Detection of the c-fos gene protein induced by PCP was probed with a c-fos (15:1000) primary monoclonal antibody that was further detected with a 1:750 dilution of alkaline-conjugated goat anti-mouse IgG, secondary antibody. A dilution of 1:5000 Vtg carp anti-mouse primary monoclonal antibody was used to elicit the Vtg protein expression. Vtg was further detected with a 1:100 dilution of alkaline phosphatase-conjugated donkey anti-goat IgG, secondary antibody. BCIP/NBT color substrate was incorporated to develop protein bands. Immunoblot 1-D protein bands were assessed for relative abundance by TotalLab computer software (Nonlinear USA Inc. Durham, NC).The effects of PCP on the viability of catfish hepatocytes are shown in Figure 1. PCP-cytotoxicity was determined in vitro using primary cultured hepatocytes from male channel catfish. Catfish hepatocytes were exposed for 48 h to seven incremental concentrations of PCP (62.5, 125, 250, 500, 1000, 2000 and 4000 μg/mL), excluding control. Exposure to PCP caused a dose-dependent decrease in cell viability; indicating acute toxicity. Cell viability percentage values of PCP-treated hepatocytes were compared to the untreated control (100% cell survival) to determine if there were any significant differences. Upon 48 hrs of exposure, the PCP concentration required to reduce cell viability by 50% (LC50), was computed to be 1987.0 ± 9.6 μg PCP/mL. NOAEL and LOAEL toxic endpoints were recorded as 62.5 ± 10.3 and 125.0 ± 15.2 μg PCP /mL, respectively. The mitogenic effect of PCP on catfish hepatocytes is shown in Figure 2. To determine the mitogenic activity of PCP, we expose catfish hepatocytes to sublethal concentrations of PCP (0, .484, .968, 1.94, 3.87, 7.75, 15.5, and 31 μg/mL). Untreated hepatocytes were compared to PCP-treated hepatocytes to determine significant differences in stimulatory patterns. Results from this experiment exhibited a concentration and time-dependent stimulatory effect on cell proliferation. A consistent fourfold to eightfold increase in cell proliferation was demonstrated with all 24 and 48 h test concentrations. For example, at 48 hrs of exposure, the percentages cell proliferation of catfish hepatocytes were about 100%, 600%, 800%, 500%, 600%, 700%, 800%, 900% in 0, .484, 1.94, 3.87, 7.75, 15.5 and 31 μg PCP/mL, respectively. These results suggest that low doses of PCP are mitogenic in catfish hepatocytes. Western Blot and densitometric analyses of c-fos expression in PCP-treated male catfish hepatocytes is shown in Figure 3. The 62-kDa c-fos gene protein was identified by Western Blot analysis. To determine the ability of PCP to transcriptionally induce c-fos expression, catfish hepatocytes were acutely exposed to various PCP-treatments (0, 750, 1500, 2000 μg/mL). PCP-treated hepatocytes were compared to the untreated control, to determine the magnitude of c-fos expression. Upon 48 hrs of exposure, a dose-dependent upregulation of the c-fos gene protein was observed at 2000 μg PCP/mL; indicating the mitogenic, inflammatory, and proliferative activity of PCP. Although no detectable protein expressions were observed at other concentrations, these data corroborate previous findings from our laboratory, where PCP simulated a dose-dependent overexpression of c-fos in HepG2 cells [1,45]. Densitometric analysis revealed that 2000 μg PCP/mL caused a substantial increase of c-fos abundance. Western Blot and densitometric analyses of Vtg expression in PCP-treated male catfish hepatocytes is shown in Figure 4. A qualitative identification of the 125 kDa-Vtg protein was made by Western Blot analysis. We tested the ability of PCP to stimulate the induction of the Vtg protein by exposing male catfish hepatocytes to various concentrations of PCP (0, 750, 1500, 2000 μg/mL) for 48 hrs. Untreated hepatocytes were compared to PCP-treated hepatocytes to determine the extent of Vtg expression. A strong induction of the Vtg protein was observed at 750 PCP μg/mL. The ability of PCP to stimulate Vtg production in male catfish hepatocytes, strongly suggests that PCP is an endocrine disrupting compound; and that PCP mimicks the estrogen hormone [3,18,20].Toxicity of pentachlorophenol to channel catfish hepatocytes. Cell viability of catfish hepatocytes exposed to PCP. The cells were treated with serial dilutions (0-4000 µg/mL) of PCP. Cell viability was measured by FDA assay as indicated in the methodology section. Absorbance readings were converted to percentages cell viability. Bars are means ± SDs, n = 3 with 8 replications per concentration. All values are significantly different (p ≤ 0.05) from control (0 µg/mL PCP).Mitogenic effect of pentachlorophenol on catfish hepatoctes, after 24 hr and 48 hr of exposure.Bars are means ± SDs, n=3 with 8 replications per concentration. All values are significantly different from control (0 µg/mL PCP), p≤ 0.001.Western Blot for c-fos expression and densitometric analysis in PCP-treated channel catfish hepatocytes. Inset shows representative Western analysis. Densitometric analysis shows significant c-fos abundance at 2000 ug/mL. Bar represents c-fos abundance; values are means ± SDs, n = 3. *Significantly different from control (0 µg/mL PCP), p ≤ 0.05.Western Blot for Vtg expression and densitometric analysis in PCP-treated channel catfish hepatocytes. Inset shows representative Western analysis. Densitometric analysis shows a significant increase in Vtg abundance at 750 ug/mL. Bar represents Vtg abundance; values are means ± SDs, n = 3. *Significantly different from control (0 µg/mL PCP), p ≤ 0.05.In the present study, we demonstrated that PCP is acutely toxic to catfish hepatocytes (Figure 1). The FDA assay was employed to assess cell viability. The sensitivity of the assay was based on the ability of the fluorescent dye to enter intact live cells, where is esterified by the esterase enzyme. The product of this hydrolysis, fluorescein, was demonstrated when exposure to PCP significantly reduced the viability of the catfish hepatocytes (LC50 = 1,987.0 ± 9.6 μg PCP/mL). PCP’s toxic potency to catfish hepatocytes was dose-dependent, indicating that increased levels of PCP caused increased toxicity to catfish hepatocytes. Interestingly, these results are consistent with previous findings from our laboratory that reported a strong dose-response relationship with respect to the cytotoxic effects of PCP in HepG2 cells [1,45].Cytotoxicity combined with the slow metabolic breakdown of organochlorine pesticides, can cause deleterious effects in species than consume fish or aquatic invertebrates [10]. Much of the current research shows that PCP is highly persistent in soil, having a half-life of up to 5 years. The persistence of organochlorine pesticides in the environment has been highly linked to pesticide bioaccumulation in aquatic species as well as exceeding guidelines for the protection of aquatic. For example, our PCP-cytotoxicity results are quite consistent with the literature; indicating that surviving fish such as bluegills, large mouth bass, and channel catfish can accumulate up to 19 mg/kg, 48 mg/kg, and 221 mg/kg of PCP in their muscle, gills, and liver, respectively for up to 10 months [46]. Previous studies have pointed out that the mitogenic activity of organochlorine compounds is strongly coupled to signal transduction through the G1 phase of the cell [3,47]. In the present study, an in vitro approach was used to investigate whether low levels of PCP have direct mitogenic effects on primary cultured catfish hepatocytes. As a consequence of exposure, accelerated proliferation rates were seen in PCP-treated catfish hepatocytes, consequently from 24 and 48hr exposure to sublethal doses. The 48 h exposure resulted in a dose-dependent, four-fold to eightfold increase in cell proliferation; indicating that PCP has a substantial mitogenic potential for channel catfish. Moreover, such mitogenic activity was also observed from our previous investigation where sublethal concentrations of PCP were mitogenic to HepG2 cell following 24 hrs of exposure [45]. The mitogenic activity of PCP is further supported by the results of the Western Blot analysis showing a potential activation of the c-fos protein catfish hepatocytes.Among the earliest responses to mitogenic signaling is the activation of transcription factors such as c-fos. c-fos has been implicated as a positive regulator of cell proliferation [48,49]. Recently we demonstrated that PCP has the ability to transcriptionally activate c-fos [1,45]. In this study, a dose-response relationship was observed with regard to PCP induction of the 62-kDa c-fos protein. We report that an appreciable expression of the c-fos protein was seen at the 2000 μg/mL level in PCP-stimulated catfish hepatocytes (Figure 3). The proto-oncogene c-fos has been exclusively implicated in cell growth, differentiation, and the inflammatory process. c-fos and other immediate early transcription factors are thought to be essential for mitogen-induced progress through the cell cycle. However, the induction of the c-fos gene involves both transcriptional and post-transcriptional machinery. Once stimulated, c-fos conjoins with c-jun, a transcription factor of the Jun family, and forms the heterodimeric complex, activator protein-1 (AP-1). The up-regulation of the c-fos protein is more consistent with the PCP-induced oxidative stress in catfish hepatocytes, since it was not expressed at a lower level of exposure. Estrogen plays a significant role in hepatic biosynthetic processes that are essential for oogenetic events in oviparous vertebrates. Vitellogenesis is an oogenetic activity by which estradiol, a natural estrogen, instructs the liver to synthesize and secrete Vtg. The activities of vitellogenesis are modulated by endogenous hormonal interactions between the hypothalamus, pituitary gland, and ovary follicle cell. Upon stimulation, the hypothalamus secretes gonadotrophin-releasing hormone that causes the pituitary gland to secrete gonadotrophic hormones in the blood. These hormones activate the follicle cell to produce estrogen, which instructs the liver to manufacture and secrete Vtg. The synthesis of Vtg is an exclusive female-specific mechanism in oviparous animals. Although male fish possess the Vtg gene, it is normally inactive [22]. Several studies reported the induction of Vtg synthesis in male fish and other aquatic species is caused by the exposure to estrogenic compounds in the environment [50,51,52]. For example, organochlorine pesticides, including alachor, β-hexachlorocyclohexane (a lindane residue), DDT, and endosulfan II, bind to the estrogen receptor [53,54] and cause, detrimental population consequences, depressed gonadal sex steroids, altered sex characteristics, and delayed sexual maturity [55,56,57,58]. An example of these deleterious effects was reported when a massive spill of dicofol, the DDT-like pesticide, was released into Florida lakes in the early 1980s. This accident resulted in a high incidence of altered sexual differentiation of the male reproductive tract and abnormal feminized steroid hormone profiles in juvenile alligators [55]. In the current study, the estrogenic activity of PCP was demonstrated by the up-regulation of the 125-kDa Vtg protein in the hepatocytes of male channel catfish (Figure 4). Upon 48 hrs of exposure, a time-dependent expression of vitellogenin to sublethal concentrations of PCP was observed. The up-regulation of this steroid-inducible protein constitutes a clear indication of the endocrine-disrupting activity of PCP. This observation is consistent with a similar study that reported increased levels of Vtg in the male Japanese Medaka (Oryzias latipes) after exposure to 4-tert-octylphenol (OP), a chlorinated compound and known environmental estrogen [50]. Acute exposure to pentachlorophenol significantly reduces the viability of catfish hepatoctyes; the LC50 was computed to be about 2000 μg PCP/mL; indicating that PCP is acutely toxic to catfish hepatocytes. Exposure to lower levels of PCP caused a four-fold to eight-fold in cell proliferation, indicating the mitogenic effect of PCP in catfish hepatocytes. The over expression of the c-fos protein in this record is more indicative of the inflammatory response, in response to PCP-induced oxidative stress in fish hepatocytes. PCP has a potential to cause endocrine-disrupting activity in channel catfish as well as mimic hormonal estrogen. This was evidenced by the expression of the 125-kDa vitellogenin protein.This research was financially supported in part by a grant from the U. S. Department of Education (Grant No. P031B990006-01) through Title III Graduate Education Program, and in part by a grant from the National Institutes of Health (Grant No. 1GR12RR13459) through the RCMI-Center for Environmental Health at Jackson State University.
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+ All rights reserved.Pentachlorophenol (PCP), an organochlorine fungicide, is extensively used in the United States for the protection of wood products. Moreover, widespread agricultural, domestic, and industrial applications have caused PCP-contaminants to enter the food chain from the environment. There is accumulating evidence indicating that PCP is highly toxic to humans, and causes injury to major organs including the lung, liver, kidneys, heart, and brain. While PCP has been shown to induce systemic toxicity and carcinogenesis in several experimental studies, the literature is scarce regarding its toxic mechanisms of action. Recent investigations in our laboratory have shown that PCP exerts both cytotoxic and mitogenic effects in human liver carcinoma (HepG2) cells [1], and in primary culture of catfish hepatocytes [2]. In the present study, we hypothesized that PCP exposure will trigger similar cytotoxic and mitogenic responses in AML 12 Mouse hepatocytes. To test this hypothesis, we performed the MTT assay for cell viability in PCP-treated and control cells. Data obtained from this experiment indicated a biphasic response with respect to PCP toxicity; showing a hormosis effect characterized by mitogenicity at lower levels of exposure, and cytotoxicity at higher doses. Upon 48 hrs of exposure, PCP chemical doses required to cause 50% reduction in the viability (LC50) of AML 12 mouse hepatocytes was computed to be 16.0 + 2.0 µg/mL. These results indicate that, although the sensitivity to PCP toxicity varies from one cell line to another, its toxic mechanisms are similar across cell lines.Chlorinated aromatic pesticides are pervasive environmental contaminants. They are well-known for their biocidal action against a broad spectrum of insects and microorganisms. The manufacturing of these compounds is usually achieved by direct chlorination of phenol or alkaline hydrolysis of synthetic chemicals [3]. A vast majority of chlorinated pesticides enter the environment via agricultural applications and wastewater effluents from chemical industries [4]. Exposure to this family of pesticides is a common and highly recognized cause of adverse health effects in humans and wildlife species. When released into the environment, these contaminants can interfere with the physiological performance of the endocrine, nervous, and reproductive systems [5] and influence sex differentiation in wildlife animals [6,7,8]. Pentachlorophenol (PCP) is a chlorinated aromatic fungicide that is commercially manufactured by direct chlorination of phenol with chlorine gas. It is a structurally halogenated hydrocarbon, composed of a benzene ring to which is attached a hydroxide radical making a chlorinated phenol. For more than 100 years, PCP has been the prevalent industrial wood preservative in the United States. Although creosote and chromated copper arsenate are popular wood preservatives, PCP has been used expansively in agricultural and domestic applications [9]. Agricultural applications include its use as an herbicide, algicide, defoliant, and fungicide [10]. Domestic applications of PCP can be found in veterinary supplies, disinfectants, fabrics, military uniforms, home-care and pharmaceutical products. In undeveloped parts of the world, PCP is used as a rot proofing agent to protect raw cotton or loom state fabric during transportation. As of 1977, an average of 50 million pounds of PCP was produced annually in the United States [11]. The widespread use of PCP has resulted in soil, water, and food contamination. At least 313 of 1,585 Superfund sites from the National Priorities List have been identified by United States Environmental Protection Agency (U.S. EPA) where PCP was found [12]. Exposure to PCP can result in the modulation of neurological responses; impairment to DNA that may lead to cancer; and hematological and immunological dysfunctions [13,14]. Pulmonary absorption of vapors, aerosols, dusts, and absorption via the skin and gastrointestinal tract, represents relevant routes of low-level exposure to PCP. Respiratory manifestations such as congestion of lungs, irritation of respiratory tract, sneezing, coughing, shortness of breath are the symptoms of PCP inhalation [15]. Symptoms of abdominal pain, nausea, fever, and aggravation of the eye, skin, and throat are heavily linked to inhalation [16]. Moreover, high levels of PCP vapors have been associated with obstruction of the circulatory system, permanent visual impairment, and central nervous system damage [16]. Alternatively, dermal contact is the most dangerous route of exposure for humans, although, inhalation of PCP is considered to be a common route of exposure in the workplace. Hodgkin’s disease, acute leukemia, and soft-tissue sarcoma have been associated with occupational exposure to technical-grade PCP [17]. Animal studies show toxicant insult to the cardiovascular, hepatic, immune, and central nervous systems from acute oral exposure to PCP [17]. Median lethal concentration and dose (LC50 and LD50) tests in rats and mice have shown PCP to have high toxicity from inhalation exposure and extreme toxicity from oral exposure [17]. At high-level exposure, PCP has the potential to induce tumorigenic [18] and carcinogenic activity [19] in mice. Similarly, oncogenic activity in mice has been documented as infrequent liver tumors, adrenal medulla pheochromocytomas, and hemangiomas [20]. Intermittent delirium, rigors, flushing and excitement are common symptoms to both children and adults from acute exposure. Subsequently, PCP exposure can also cause neurological disorders of tachypnea, cerebral edema, and swelling of the myelin sheath [21]. Chronic exposure to PCP causes injury to the liver, kidneys, and central nervous system. Blood levels of children on the average are 1.8 times higher than adults; indicating that children are more susceptible to PCP exposure [21]. Babies that nurse liberally, experience tachycardia, hepatomegaly, progressive metabolic acidosis, proteinuria, azotemia, irritability followed by lethargy, pneumonia or bronchitis, and aplastic anemia [21]. Evidence of human mutagenicity and carcinogenesis due to PCP exposure is inadequate. However, the U.S. EPA has established PCP as a probable human carcinogen-Group B2, based on suggestive evidence of carcinogenicity from laboratory animal studies [17]. It is believed that the mechanism by which PCP exerts its toxic action involves uncoupling mitochondrial oxidative phosphorylation, thereby causing accelerated aerobic metabolism and increasing heat production [22,23]. Previous investigations in our laboratory have shown that PCP induces acute toxicity and transcriptionally activates a constellation of stress genes in HepG2 cells [24]. From our laboratory findings, we demonstrated that PCP has the ability to undergo Phase I biotransformation in the liver (CYP1A1 and XRE), to cause cell proliferation (c-fos), to cause growth arrest and DNA damage (GADD153 and p53), to influence the toxicokinetics of metal ions (HMTIIA), and to induce proteotoxic effects (HSP70 and GRP78) in HepG2 cells [1,24]. Our laboratory has recently reported that PCP exerts both cytotoxic and mitogenic effects in human liver carcinoma (HepG2) cells [1], and primary culture of catfish hepatocytes [2]. In the present study, we hypothesize that PCP exposure will trigger similar cytotoxic and mitogenic responses in AML 12 Mouse hepatocytes.Pentachlorophenol (C6Cl5OH, CAS No. 87-86-5, Lot No. 01530TS), with purity 98.0% was purchased from Sigma-Aldrich Chem Co., (St. Louis, Missouri). Dulbecco’s Modified Eagle’s Medium (DMEM) was purchased from Hyclone (Lot No. ANK19799; Logan, Utah) and tissue culture supplements were purchased from American Type Culture Collection (ATCC) Manassas, VA. Dulbecco’s phosphate buffered saline (Lot No. 1163547) was obtained from Invitrogen Corporation (Grand Island, New York). Thiazolyl blue trazolium bromide CAS 298-93-1, purity 97.5%, and dimethyl sulfoxide were purchased from Sigma-Aldrich (St. Louis, Missouri). Alpha mouse liver 12 (AML 12) hepatocyte cultures have been established from a mouse transgenic for human transforming growth factor α (ATCC CRL-2254, Manassas, VA). The cells were stored in liquid nitrogen in the laboratory until use. Next, cells were thawed for 2 min in a dry bath at 37oC. After thawing, the content of each vial was transferred to a 75 cm2 tissue culture flask diluted with DMEM, supplemented with 10% fetal bovine serum (FBS) and 1% streptomycin and penicillin, and incubated at 37oC under an atmosphere of 5% CO2 incubator in humidified air to allow the cells to grow and form a monolayer in the flask. Subsequently, cells grown to 80-95% confluence were washed with phosphate buffer saline (PBS), trypsinized with 5mL of 0.25% (w/v) EDTA, diluted, counted (5 x105 cells/well), and seeded in 96-well microtiter tissue culture plates prior to treatment.The MTT [3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide] assay was performed for assessing cytotoxicity and cell proliferation activity in AML 12 mouse hepatocytes exposed to PCP. Viable cells are able to convert MTT to a water-insoluble formazan dye. AML 12 mouse hepatocytes were maintained in Dulbecco’s Minimum Essential Medium (DMEM) supplemented with 10% fetal bovine serum (FBS) and 1% penicillin/streptomycin, and incubated at 37oC in a 5% CO2 incubator. On day one, 180 μl/well aliquots of cell solution (5.0 x105 cells/mL) were seeded in 96 well polystyrene tissue culture plates. On the day of exposure, old medium was removed and replaced with 180 µl of fresh medium. Subsequently, 20 μl aliquots containing varying concentrations of PCP were added to each well. Cells were placed in the incubator for 48 h at 37oC in 5% CO2. One hour before the end of the incubation period, 20 μl aliquots of MTT solution (5mg/mL) were added to each well containing cells. Culture plates were incubated at 37oC for 1 h. Medium containing MTT was removed and 200 μl aliquots of DMSO were added to each well. The culture plates were placed in the 37oC incubator for 5 min to dissolve air bubbles, and transferred to a Bio-Tek Model – EL 800 microplate reader where absorbance was measured at 550 nm. Absorbance readings of 550 nm from cell viability experiments were transferred to percentages to assess significant differences in treated cells compared to control cells. Graphs were made to illustrate the dose-response relationship with respect to cytotoxicity or mitogenic activity. Standard deviations were determined, and the Student’s t-test values were computed to determine if there were significant differences in cell viability in PCP-treated cells compared to control cells. We used the MTT-assay to assess the viability of PCP-treated AML 12 mouse hepatocytes. Mouse hepatocytes were exposed to increasing levels of PCP for 48 h. Metabolically-active cells were able to convert MTT to water-insoluble dark-blue formazan crystals. The cytotoxic effect of PCP on AML 12 mouse hepatocytes is shown in Figure 1. A strong biphasic-response pattern was demonstrated with respect to the cytotoxicity of PCP. Cell viability percentage values of PCP-treated hepatocytes were compared to the untreated control (100% cell survival) to determine if there were any significant differences. The percentages for cell viability were recorded as 163 + 2.3%, 197 + 2.8%, 136 + 3.5 %, 48 + 2.1%, and 40 + 1.9% for concentrations of 1.95, 3.95, 7.8 15.6, and 31.2 μg PCP/mL, respectively. A statistically significant (p < 0.05) increase in cell viability within the concentration range of 0 - 7.8 μg PCP/mL was observed along with a concomitant decrease within the concentration range of 15.6 - 31.2 μg PCP/mL. The 48h-LC50 was computed to be 16.0 + 2.0 μg PCP/mL, indicating that PCP is acutely toxic to AML 12 mouse hepatocytes. Toxicity of pentachlorophenol on AML 12 mouse hepatocytes. The cells were treated with serial dilutions (0-31.2µg/mL) of PCP. Cell viability was measured by MTT assay as indicated in the methodology section. Absorbance readings taken from survival cells were converted to percentage cell viability. Bars are means + SDs, n = 3 with 8 replications per concentration. *Significantly different (p < 0.05) from control (0 µg/mL PCP).To determine the mitogenic activity of PCP, we exposed mouse hepatocytes to sublethal concentrations of PCP (0, .975, 1.95, 3.95, and 7.8 μg/mL). The mitogenic effect of PCP on AML 12 mouse hepatocytes is shown in Figure 2. Untreated hepatocytes were compared to PCP-treated hepatocytes to determine significant differences in stimulatory patterns. Data from this experiment exhibited a concentration and time-dependent stimulatory effect on cell proliferation. All 24 h and 48 h test concentrations within the range of .975 – 7.80 μg PCP/mL induced a one-fold to three-fold increase in proliferative activity. For example, at 24 h of exposure, the percentages of cell proliferation of AML 12 mouse hepatocytes were about 100%, 125%, 150%, 250%, and 174% in 0, .975, 1.95, 3.95, and 7.8 μg PCP/mL, respectively. These results suggest that low doses of PCP are mitogenic in AML 12 mouse hepatocytes. Mitogenic effect of pentachlorophenol on AML 12 mouse hepatocytes cells. AML 12 mouse hepatocytes were treated with sublethal concentrations of PCP (.975 - 7.80 μg/mL) for a 24 h and 48 h period. Cell proliferation was determined based on the MTT assay. Untreated hepatocytes were compared to PCP-treated hepatocytes to determine significant differences in stimulatory patterns. Bars are means + SDs, n = 3 with 8 replications per concentration. All values are significantly different from control (0 µg/mL PCP), p < 0.001.We examined cell morphology with regard to duration and intensity of PCP-treated AML 12 mouse hepatocytes to determine cell injury. Morphologic changes in AML 12 mouse hepatocytes treated with PCP is shown in Figure 3. Morphologic injury was assessed by light-field microscopy and by determination of cell viability (MTT) post-exposure. PCP-treated hepatocytes (B), were distinguishable from that of untreated control (A). Upon 48 h of exposure, a 15.6 µg PCP/mL concentration caused distortion of monolayer morphology, changes in cell shape, and decreased viability (48 + 2.1%). We conclude that changes of cell morphology are closely related to the extent of stress and functional status of AML 12 mouse hepatocytes caused by PCP exposure. Morphologic changes in AML 12 mouse hepatocytes treated with pentachlorophenol. Toxicity to mammalian cells is frequently an elaboration of dose-dependent occurrences as well as complex processes that respond to genotoxic stress. In the present study, we demonstrated that PCP (100 μg/mL in 1% DMSO) is acutely toxic and causes injury to AML 12 mouse hepatocytes (Figure 1). This was evidenced by the ability of PCP to cause a 50% reduction in the viability of AML 12 mouse hepatocytes (LC50 = 16.0 + 2.0 μg PCP/mL). Interestingly, these results are consistent with previous findings from our laboratory that reported strong dose-response relationships with respect to the cytotoxic effects of PCP in HepG2 cells (23.0 + 5.6 μg PCP/mL) and catfish hepatocytes (1987 + 9.6 μg PCP/mL) [1,2,24]. Although the mechanisms by which PCP exerts its toxic effect is somewhat unclear, we have demonstrated that the toxic potency of PCP is a dose-dependent event with regard to AML 12 mouse hepatocytes, HepG2, and catfish hepatocytes.The mechanisms by which stress triggers signal transduction pathways that lead to mitogenic activity are of great interest. Steroid-like compounds, such as PCP, can influence rapid activation of cell proliferation and facilitate signal transduction through the G1 phase of the cell [25,26]. We previously reported that sublethal concentrations of PCP were mitogenic to HepG2 cells and catfish hepatocytes following 24- and 48 h of exposure [1,2]. In the present study, we demonstrated that low level concentrations of PCP were found to stimulate proliferative activity in AML 12 mouse hepatocytes (Figure 2). The greatest mitogenic response was observed at 3.95 μg PCP/mL. Stress gene expression that facilitates cell proliferation, migration, differentiation, and survival, is the eventual outcome of signaling pathways that coordinate long-term cell adaptation. We have previously shown that PCP has the ability to transcriptionally activate the c-fos proto-oncogene, in HepG2 cells and catfish hepatocytes, [1,2,24]. c-fos has been implicated in proliferative machinery and is thought to be essential for mitogen-induced progression through the cell cycle [27]. Mitogen-activated protein kinase (MAPK) is one of the major signal transduction pathways activated when cells are exposed to inflammatory cytokines or stress [28]. Our findings strongly suggest that stress-activated protein kinases (SAPK)/Jun N-terminal kinases (JNK), members of the MAPK family, could be involved as a result of signal transduction within PCP-treated hepatocytes. Therefore, we believe that the SAPK/JNK cascade is likely to be involved in facilitating the mitogenic response in PCP-treated AML 12 mouse hepatocytes.When cells are exposed to a variety of environmental stresses, cell death is the consequence of severely disturbed extracellular environmental conditions. However, if the injury is acute, the cell can survive in a damaged state and adapt to the injury (reversible) or it can die (irreversible or cell death) [29]. In the present study, morphologic changes in AML 12 mouse hepatocytes were the consequences of 48 h exposure to PCP (Figure 3). Severe changes in cell morphology were observed at 15.6 µg PCP/mL. We have shown that PCP causes injury to AML 12 mouse hepatocytes.It can therefore be concluded from the findings of this research that: 1) PCP is acutely toxic to AML 12 mouse hepatocytes, with an LC50 of 16 + 2.0 μg/mL after 48 hours of exposure; 2) upon 24- and 48 h of exposure to sublethal concentrations, PCP demonstrated a strong mitogenic activity; 3) PCP causes injury to AML 12 mouse hepatocytes. This was confirmed by the morphologic changes in AML 12 mouse hepatocytes as a consequence of 48 h exposure; and 4) the toxic mechanisms of PCP are similar across cell lines.This research was financially supported in part by a grant from the U. S. Department of Education through Title III Research Excellence Fund, Grambling State University, and in part by Title III Graduate Education Grant No. P031B990006-01 to Jackson State University. We thank Dr. Abdul Mohamed, Dean of the College of Science, Engineering and Technology, for his technical support of this research project.
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+ All rights reserved.The effect of bonny-light crude oil was assessed in adult albino rats. The rats were administered with 200, 400, and 800mg/kg body weight of the crude oil orally for 7 days. Fluid intake was measured daily, initial and final animal body was recorded. The toxic effects on the kidneys were assessed and histological studies carried out. The results revealed that the kidney cells were damaged; crude oil caused a destruction of the renal reserve capacity. There was a significant increase (p ≤ 0.05) in creatinine in the high dose group (800mg/kg), and a significant decrease (p ≤ 0.05) in urea concentration. Histological examination indicates that crude oil induced severe pathologic changes in the forms of necrosis and oedema.Crude oil has been described as a complex mixture of over 6000 potentially different hydrocarbons and metals [1]. Crude and refined petroleum, and oil field chemicals and emissions are highly complex chemical mixtures. Crude petroleum contains hundreds of compounds and the chemical composition varies between geologic formations [2].Knowledge of human responses to acute exposures to petroleum components comes from studies with several solvents containing benzene and petroleum [3]. Recognized human biochemical and physiological responses associated with acute exposures to natural crudes are mainly transient and short lived unless the concentrations of the components are unusually high [4]. Pollution implies that a particular level of contamination has deleterious effects, which may take the form of hazards to human health, interference with human activities, reduction of human amenities or harmful effect on living resources. Substantial amounts of those potentially toxic substances have been introduced into the environment. There is a concern that workers and other individuals exposed to crude oil might have an increased incidence of organ damage. After absorption via pulmonary or gastrointestinal routes, crude oil is transported in plasma initially bound to albumin and other larger proteins to the liver. In Nigeria, crude oil is predominantly found in the riverine areas. Over the years the local population has used crude oil for various ailments such as gastrointestinal disorders, burns, foot rot and leg ulcers, poisoning and witchcraft. The present study, investigates the renal effect of bonny-light crude oil.Matured albino rats weighing between 124-130g were obtained from the Animal Facility Center of the Department of Pharmacology and Toxicology, National Institute for Pharmaceutical Research and Development, Abuja, Nigeria. The animals were acclimated to housing conditions for at least one week prior to commencement of the experiment. Animals were housed singly at ambient temperature of 23 ± 3° C and a 12 h light, 12 h dark cycle. Water and food were provided ad libitum for the animals.Animals were divided into 4 groups of 5 male rats each. The three test groups received by gavage for seven days freshly prepared Bonny light crude oil at doses of 200 mg/kg, 400 mg/kg and 800 mg/kg while the 4th group received a solution of Tween 80 in water only as control. These doses were based on that used by the local population in folkloric medicine. Food and fluid intake were measured daily. The final body weights of the rats were recorded.Nigerian bonny-light crude oil was obtained from Nigerian National Petroleum Corporation (N.N.P.C.) Research Laboratory Port Harcourt, Nigeria. The bonny light crude oil was dissolved in Tween 80. At the end of the seven days exposure period, the animals were weighed and sacrificed under chloroform anesthesia, and serum obtained. Sodium and potassium in serum were determined with Corning 410C clinical flame photometer (Corning Instruments, England) at 589 and 768nm respectively. Titrimetric method of Schales and Schales (1941) [5] was used for the determination of chloride. The potassium and sodium were estimated using the flame emission spectrophotometer (corning brand). Urea was analysed spectrophotometrically at 525nm [6]. Creatinine was determined using the method of Taussky 1956 [7] at 520 nm. Serum potassium, sodium, bicarbonate and chloride were measured by the use of commercial kits (Biosystem, Spain).The animals were then sacrificed under chloroform anaesthesia and the kidneys were harvested, weighed and fixed in 10% buffered formalin for 48 hrs. The kidneys were processed using an automatic tissue processor, embedded in paraffin wax, and sections (5µm thick) cut using a rotary microtome. The sections were stained by haematoxylin and eosin (H&E) method, and examined and photographed by using a light microscope. Two Histopathologists examined the sections, independently.The values were reported as means ° SEMs. Data were analyzed using the Student’s t-test, while Duncan’s multiple-range test was used to test for differences between treatment groups using Sigma-stat 2.0 software. P < 0.05 was considered statistically significant.Table 1 shows the effect of Bonny light crude oil on body weight, fluid intake, absolute and relative weights of kidney after 7 days of sub-chronic exposure to male rats. There was no significant difference in the mean fluid intake between the three dose levels. The animals demonstrated a progressive increase in body weight in the control and lower doses during the exposure period and there was a significant reduction in body weight in the high dose group (800mg/kg) compared to the control. There were no significant differences (p ≥ 0.05) in the absolute and relative kidney weights of the animals in the three experimental groups.The results show a significant increase (p < 0.05) in serum creatinine in the 800mg/kg crude oil group; whereas there was significant increase in potassium in the 200 and 400mg/kg crude oil treated groups. There was also a significant decrease (p < 0.05) in serum urea, sodium, bicarbonate and chloride in the bonny-light treated group when compared with the control (Table 2).The histological changes of kidney in the different treatment groups are shown in Figure 1. The control group showed normal histological structure. Low dose crude oil produced slightly congested glomeruli and hyperemia, a congestion caused intraglomerular space between the tuft and capsule. The glomeruli appeared infiltrated and congested creating wider spaces between the tuft and the capsule in the medium dose crude oil group. Also in this group some of the glomeruli appeared degenerated. The high dose group showed empty and widened capsules with severely degenerated glomeruli and noticeable necrosis.Exposure of humans and animal to crude oil, which is increasing in terms of the environmental levels, and application to body, may be toxic. Crude oil is used in folkloric medicine in the Niger-delta area of Nigeria for the treatment of various ailments including stomach up-set, wound, and burns [8]. The route of administration is mostly oral and external application for burns and wounds. In several organs, mainly heart and liver, cell damage is followed by increased levels of a number of cytoplasmic enzymes in the blood, a phenomenon that provides the basis for clinical diagnosis of heart and liver diseases e.g. liver enzymes are usually raised in acute hepatotoxicity but tend to decrease with prolonged intoxication due to damage to the liver cells. The Nigerian Bonny crude oils are classified as light crude oils, with aromatic hydrocarbons accounting for up to 45% of the total hydrocarbons. As aromatic hydrocarbons are relatively soluble in water [9], it is expected that the potential of this light crude oil to have adverse toxic effects is higher than for heavier, less water–soluble crude oils. It is known that lipophilic xenobiotics may have the characteristics of both electron uncouplers and energy inhibitors [10]. Since a large proportion of the crude oil components is lipophilic in nature biological membrane may be the target sites where adverse effects occur.The kidney can suffer considerable damage before losing sufficient function to modify the normal clinical indication of renal disease such as the serum creatinine concentration [11]. Approximately 50% or more of renal capacity can be lost before serum creatinine become abnormal and disease is detectable clinically. A battery made up of a combination of different types of test can aid in the detection of damage by a nephrotoxin and also allows for the determination of various threshold damage. The detection of renal damage at a reversible stage is necessary before effective preventive measures can be taken to halt the progress of damage to the irreversible stage. Effect of Bonny light crude oil on body weight, fluid intake, absolute and relative weights of kidney after 7 days of sub-chronic exposure to male rats.Numbers in parenthesis indicate initial body weight of rats.N = 5*Significantly different from control at p < 0.05†Effect of Bonny light crude oil on serum electrolytes, urea and creatinine after 7 days of sub-chronic exposure to male rats. N = 5. *Significantly reduced (p < 0.05) compared to control.#Significantly increased (p < 0.05) compared to control.Photomicrographs of H&E stained sections of rat kidney (Magnification X250).The absolute and relative kidney weights, as well as body weight gain were significantly (p < 0.05) decreased by exposure to the bonny–light crude oil. Researchers have reported that absolute kidney weight is a relatively sensitive indicator of nephrotoxicity for known nephrotoxicants, with nephrotoxicity defined as increased kidney weight (absolute or relative) [12].There were no significant (p < 0.05) differences in the absolute and relative kidney weights of the animals in the three test groups. This observation seems not to conform with the above definition of toxicity probably due to the short duration of the exposure to the bonny-light crude oil. However the significant (p < 0.05) increase in serum creatinine and potassium in the treated groups and also the significant (p < 0.05) decrease in serum urea, sodium, bicarbonate and chloride in the bonny-light treated group when compared with the control is suggestive of renal pathology. This in part is in agreement with Kluwe definition of renal toxicity [12].
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+ All rights reserved.We studied the acute toxicity of a raw effluent from a battery manufacturing plant (Pilcam) in Douala, Cameroon, to a freshwater fish (Oreochromis niloticus), and subsequently evaluated its sub-acute effects on water quality and the biota in freshwater microscosms. The acute toxicity test was based on 96 hrs static renewal bioassays that resulted in 96-h LC50 and LC90 values of 16 and 20.7% (v/v), respectively. The sub-acute experiments were conducted by exposing several species of aquatic organisms (plankton, macro-invertebrates and mollusks) to lower effluent concentrations [1.6%, 8.0%, 16% (v/v)] for six weeks, and monitoring their survival rates, as well as the physical and chemical characteristics of water. These concentrations were based on 10%, 50%, and 100% of the 96 h - median lethal concentrations (LC50) of the effluent to the freshwater fish, Oreochromis niloticus. Significant effects on functional parameters, such as, chlorophyll-a and total protein could not be demonstrated. However, the activity of alkaline phosphatase was significantly inhibited at all concentrations tested. Phytoplankton, zooplankton, macro-invertebrate communities and snails were negatively affected by the effluent application at concentrations ≥ 8% (v/v), with chlorophyta, ciliates, ostracoda, annelida, planaria and snails being the most sensitive groups. The snails were eliminated after 24 h exposure from microcosms treated with effluent at concentration ≥ 8% (v/v). Effluent exposure also caused significant effects on water quality parameters (DO, pH, hardness, conductivity, color, turbidity, ammonia) in general at concentrations ≥ 8% (v/v). Temperature and alkalinity were not significantly affected. Overall, data from this research indicate that a dilution of the Pilcam effluent down to 1.6% does not provide protection against chronic toxicity to aquatic organisms. Further studies are needed to determine the no observable adverse effect level (NOAEL), as well as a chronic reference concentration for this effluent.Cameroon is located along the end of the West side of West African Coast of the Gulf of Guinea in the Atlantic Ocean.The portion of the Gulf of Guinea which is under Cameroon’s territorial waters has the most diverse and productive ecosystems. The country's coastline is 402 km long and extends from the Rio del Rey Estuary at the boundary with the Federal Republic of Nigeria, and to the Kribi coast along the boundaries with Equatorial Guinea and Gabon. It is the major sink for discharge of pesticides and persistent organic pollutants (POPs) by inland rivers that drain into the Atlantic Ocean. The region has the most industrialized areas of the country, high population density, as well as accounting for about 80% of the national annual GDP. A traditional meal in this area includes fish foods. About 39% of the population in this zone has an annual average consumption of 7.02 kg of fish harvested from the Atlantic Ocean.About 95% of Cameroon's industries including the battery production factory (Pilcam) lie along the coastal town of Douala, the economic capital and most populous town in Cameroon. Most of these industries discharge their liquid wastes (often without treatment) into the urban drainage network which in turn empties its content into the Atlantic Ocean. Of these industries, less than 10% have on-site waste water treatment facilities, and those that own such facilities operate under different standards. The main contributors to environmental pollution from these industrial wastes include heavy metals, pesticides and polychlorinated biphenyls (PCBs) which are persistent and have high potential to bio-accumulate in the food chain.The production of batteries in Pilcam generates very toxic liquid waste containing heavy metals such lead, cadmium, mercury, zinc, copper and chromium [1,2]. Cadmium causes necrosis and chlorosis in aquatic organisms. It is lethal to plankton at 2 µg/l. Aquatic invertebrates seem to resist more to the toxic effects of this element than fish [3].The bio-accumulation of lead in crayfish Palaemon elegans induces a significant depletion of red blood cells [4]. Bacteria transform mercury into toxic and less biodegradable methyl mercury that bio-accumulates in the food chain [1,
2
+ 5,
3
+ 6,
4
+ 7]. Mercury interferes with the respiratory function in mollusks and crustaceans [8]. High levels of zinc in water suffocate fish, following the destruction of the gill epithelium [3,5]. In mollusks and crustaceans, lethal and sub-lethal levels of zinc have been reported to cause an internal development of hypoxia that blocks respiration [8]. Several studies have shown that there has been a steady increase in the destruction of marine life (decreasing fish and mangroves population, in particular), due to pollution from the sources cited above [9]. Microbial communities and aquatic organisms tested in laboratory microcosms have previously shown sensitivities to several toxic compounds at environmentally realistic levels [10,11,12,13,14,15,16]. In this research, we evaluated the acute toxicity of a raw effluent from a battery manufacturing plant (Pilcam) in Douala, Cameroon, to a freshwater fish (Oreochromis niloticus), and subsequently studied its sub-acute effects on water quality and the biota in freshwater microscosms. Both taxonomic (phytoplankton and protozoan species richness and composition and macro-invertebrates communities and snails) and non taxonomic (biomass, stored material and enzyme activity) responses were measured. Juvenile Oreochromis niloticus measuring 5.33 ± 0.64 cm and weighing 2.50 ± 1.10 g were collected from the Yaounde Municipal Lake. They were acclimatized to laboratory conditions in all-glass aquaria (150 cm x 70 cm x 60 cm) for four weeks prior to experimentation. The aquaria were supplied with aerated spring water (temperature 24.1 ± 0.2 °C; total hardness 62.0± 1 mg/l as CaCO3; pH 6.20± 0.02; alkalinity 75.0± 0.6 mg/l as CaCO3; NH3-N and total residual chlorine were below the detection limits of 20 μg/l and 5 μg/l, respectively).A composite sample of the raw chemical industrial effluent was collected from its discharge pipe in the urban drainage network and transported to the laboratory. This sample (60l in polyethylene cans) was obtained by mixing proportionally to their flow rates. Grab samples (mean flow rate 26 m³/d) were collected every 15 minutes for 6 hrs representing a working shift period at the factory. The physico-chemical properties of this raw effluent included temperature = 24.1 ± 0.2 °C; salinity = 0.70 ± 0.05% total hardness = 164.0 ± 0.5 mg/l as CaCO3; pH = 5.15 ± 0.02; conductivity =1926 ± 3 µS/cm; and total suspended solids= 965± 2 mg/l. Its heavy metals content was not evaluated.The acute toxicity tests were conducted according to Standard Methods [17] and EPA methods for static-renewal acute toxicity tests [18]. Five test concentrations (dilutions) of the effluent and a control were used. Each effluent concentration was tested in duplicate and the experiments were repeated once to evaluate variability and the average values of toxic end-points determined. The LC50 and LC90 were estimated using the U.S. EPA probit analysis computer program version 1.3 used for calculating effective concentrations written by C. Stephen of the Duluth U.S. Environmental Protection Agency's Environmental Research Laboratory. Sub-acute concentrations of the raw effluent [1.6%, 8.0%, 16% (v/v)] were used in subsequent freshwater microcosms studies. These test concentrations were respectively 10%, 50% and 100% of the 96-h median lethal concentration (LC50) obtained from the test fish (Oreochromis niloticus).Laboratory microcosm’s experiments were conducted in glass microcosms (60 x 30 x 35 cm). Exposure medium was natural water collected from the pelagic zone of the Yaounde Municipal Lake, using a small ZODIAC boat model MR II GR. The average temperature, pH, dissolved oxygen, harness, alkalinity, conductivity and total dissolved solids (TDS) of the water were 24.07°C, 7.12, 5.13 mg/l, 70 mg CaCO3/l, 245 µs/cm and 164.7 mg/l, respectively. A total of 40 l of medium were added to each of the 4 microcosms. The test organisms for these microcosms’ experiments included the natural microbial and benthic macro-invertebrate communities obtained by filtering (< 64 µm plankton net) lake water from the pelagic and herbarium littoral zone, respectively. Five liters of filtrate were added to each microcosm. Snails of Lymnea, Bulinus, Physa, and Biomphalaria genera were also added to each microcosm. The microcosms thus constituted were allowed to acclimate to laboratory conditions for 3 weeks. The water loss by evaporation in microcosms was compensated by adding distilled water. The composition of these microcosms in terms of organisms prior to the experiment is shown in Table 1.The experiment consisted of 6-weeks of exposure and monitoring. The microcosms were dosed with a single application of raw effluent at the nominal concentrations of 0, 1.6, 8.0 and 16% (v/v). Sampling was done every 24 h for the first 4 days, the seventh day, and subsequently every 7 days until the end of the experiment. Microcosms were lighted with daylight-equivalent bulbs (Color Rendering Index >90; Durotest, Corp.) at an intensity of approximately 5000 lux on a 16h / 8h of light/dark photoperiod. Temperature was uncontrolled and fluctuated with ambient laboratory temperature (25-27°C).The physical and chemical conditions monitored included microcosms temperature, pH, alkalinity, color, total suspended solids, dissolved oxygen, turbidity, ammonia-N, total hardness, and conductivity. Microbial biomass accumulated in microcosms after each exposure period was estimated by measuring total protein. An aliquot of each sample was concentrated by gentle centrifugation (1000 rpm for 10 minutes). Aliquots of resulting residues were extracted using 0.5 N NaOH [19] and protein analyzed by the method of Bradford [20]. Aliquots of each sample residues were also analyzed for chlorophyll a [17] and alkaline phosphatase activity [21].Protozoan and phytoplankton densities and species richness within the microbial communities were used as an indicator of community complexity, since it was not feasible to identify all microbial species present. The size and the number of protozoan species in each sample were determined within 6 h of collection by microscope examination [22]. Samples for phytoplankton density and species distribution were collected and preserved with acid Lugol's. Organisms were concentrated in settling chambers and counted using a Wild inverted microscope. Distribution changes of species were recorded from counts of one sample per microcosm at detection limits of 23 organisms per ml [23].Plankton species in each sample were assigned to a taxa group to identify any major compositional shifts (e.g the disappearance of taxa). Most protozoan species fell within four major groups (rotifer, ciliates, copepoda, and ostracoda), whereas phytoplankton species fell within three (cyanophyta, chromophyta and chlorophyta). Therefore, only these taxa groups were used in subsequent analyses.Benthic macro-invertebrates were also examined, and assigned to a taxonomic group to identify differences in sensitivity to the effluent. These benthic organisms were distributed in groups including annelids, planaria and nematodes. The number of living organisms in each taxa was recorded after each exposure period.Table 1 showed that the biota composition (density and specific richness) varied from one microcosm to another. Consequently, all results were expressed in terms of percent of the initial level [(level at t / level at to) x 100] and were presented in histograms using software Excel version 5.0. The mean values of water quality parameters of the untreated and treated microcosms were compared using the paired Student t-test from the Systat software version 5.0The acute toxicity of the effluent to fish (Oreochromis niloticus) is shown in Table 2. The 48-h LC50 and 48-hLC90 for Oreochromis niloticus were 20.8% (v/v) and 26.6% (v/v) of effluent respectively, whereas the 96-h LC50 and 96-h LC90 values were 16% (v/v) and 20.7% (v/v) respectively. These results show that Pilcam effluent exerts a high toxic effect on the freshwater fish Oreochromis niloticus. When exposed for 48h and 96h to a 14% (v/v) diluted effluent, 0% and 30% fish kill were recorded respectively. When the percent dilution of the effluent was increased to 20%(v/v), fish mortalities of 40% and 90% were obtained after 48h and 96h of exposure respectively.As this effluent contains mostly heavy metals (Hg, Cd, Pb, and Zn) [1,2], high levels of zinc in water were reported to suffocate fish by damaging the gill epithelium, thus blocking respiration [3,5]. Mercury is probably the most serious metallic pollutant of the seas. In oxygenated marine sediments bacteria may convert the less toxic, inorganic form of the metal to the more toxic, organic form of methyl mercury. This chemical form is relatively mobile in the environment and tends to accumulate in fish. Some fish, such as tuna have naturally high levels which may approach or even exceed acceptable limits for human consumption in large, older specimens even without apparent adverse health effects in the fish [24].Histopathological damages in Oreochomis niloticus exposed to petroleum refinery effluent containing Pb, Fe, Zn, Cd, Cu, Ni, and Hg have been reported [25]. This fish, collected from High Dam and Nasser lakes in Egypt has been shown to concentrate in its dorsal fins and liver high levels of lead, cadmium and zinc. Untreated industrial effluent from factories near Islamabad and containing Pb, Fe, Zn, Cd, Cu, Ni and Hg caused 100% mortality in carp fishes exposed for 24 hours [26].. The mean values of the studied water quality parameters are shown in Table 3. The variations of the temperature ranged from 98% to 106% of initial values in all microcosms during the study. The pH variations ranged between 80% and 130% of initial values with a significant increase in the 8.0 and 16.0 % (v/v) microcosms. Total alkalinity (as CaCO3) variations were in general small ranged between 96% and 108% of initial values in microcosms, whereas the total hardness (as CaCO3) variations were irregular and ranged between 50% and 250% with a significant increase in 8% and 16% (v/v) treated microcosms. The water remained highly mineralized in effluent treated microcosms with conductivity fluctuations ranging from 110% to 180% of its initial value. The total suspended solids generally decreased in all microcosms. It varied from 10% to 115% of its initial value with significant decrease in the 8 and 16% (v/v) microcosms. The values of the apparent color and turbidity decreased in all microcosms. The color and turbidity ranged from 18% to 125% and 18% and 105% respectively of their initial values with significant decreased in the 8 and 16% (v/v) microcosms. Initial biota composition in microcosms (t = 0 day; n=number)Experimental data from the acute toxicity test of Pilcam effluent on fish (Oreochromis niloticus) a Average number of 2 replicatesMean values of water quality parameters studied of control and treated microcosms with various concentrations of the effluent during the study (6 weeks of exposure)TSS: Total suspended solids; Alk phos: Alkaline phosphatase activity;T. hardness: Total hardness; T. alk: Total alkalinity. * Parameter with Effluent addition significantly different (p < 0.05) from no Effluent addition, using the paired Student t-test.* Parameter with Effluent addition significantly different (P<0.05) from no Effluent addition, using the paired Student t-test.The significant decrease of total suspended solids, apparent color and turbidity could be due to sedimentation and the reduction in organic loading due to fish kill. Dissolved oxygen ranged from 1.72 to 6.84 mg/l in all microcosms. It generally decreased in all microcosms during the first three weeks of exposure, probably due to the death of organisms and fouling of water. This decrease was significant in the 8% and 16% (v/v) microcosms. This is in accordance with the findings that showed that the increase in heavy metals contents of seven study sites in coastal areas in Kuwait caused a depletion in dissolved oxygen [27]. Ammonia-N also significantly decreased (p < 0.05) in microcosms treated with 8% and 16% (v/v) of effluent. Ammonia-N being the final major product of the protein catabolism excreted by aquatic animals [28], its significantly low excretion could be ascribed to the decrease of the density of organisms in these microcosms. Chlorophyll a measures the biomass of autotrophic organisms accumulated in the microcosms. The chlorophyll-a of phytoplankton communities in the microcosms increased steadily during the first 4 days of exposure in all microcosms with some stimulation in all treated microcosms (up 1150% of initial value), and declined rapidly and remained low in all microcosms after 21 days of exposure. Based on the intrinsic variations of the chlorophyll-a content relative to initial value in each microcosm, the inhibitive effect of effluent on the photosynthesis could not be demonstrated (Fig. 1).Three days following the application of Pilcam effluent, microbial biomass (protein) in the 1.6% (v/v) microcosm was increased while it slightly decreased in the 16% (v/v) microcosm as illustrated in Fig. 1. After 4 days, the biomass generally tended to decrease with increasing effluent concentration and exposure time, but the effect was not significant (p > 0.05) over the entire exposure period.The activity of alkaline phosphatase enzyme measures the rate of cleavage of phosphorus organic compounds by the microbial community. One day following the application of the effluent, the alkaline phosphatase activity significantly decreased in all treated microcosms relative to the untreated microcosm (Fig. 1). Of the parameters analyzed for the plankton community function, this parameter seemed to be the most sensitive to Pilcam effluent-induced stress. Cyanophyta were the most abundant taxa and showed a global decrease in density throughout the experimentation. This decrease seemed more pronounced in 8 and 16% (v/v) microcosms (Fig. 2). The density of chlorophyta was not regular with increasing exposure period. They disappeared completely from the 1.6% (v/v) microcosm after 28 days of exposure, and from the 8 and 16% (v/v) microcosms after 3 and 14 days of exposure, respectively (Fig. 2). The density of chromophyta moderately varied during the first week of exposure with a significant decrease in the 1.6% (v/v) microcosm, and subsequently was eliminated from all microcosms as the fourteen day of exposure. In the control, an abrupt re-colonization of chromophyta was observed on the 42nd day (Fig. 2). The species present in all microcosms and occasionally abundant in the 1.6% (v/v) microcosm throughout the study were Planktothrix mougeotii and Oscillatoria putrida with putrida being more sensitive to the effluent toxicity than mougeotii.The presence of cyanophyta in all microcosms throughout the study as compared to chlorophyta and chromophyta may indicate a probable resistance of this taxa to Pilcam effluent stress. This is consistent with reported findings showing that cyanobacteria are qualitatively and quantitatively predominant in the Godavari stream polluted by heavy metals [29]. This is also in accordance with the observation that cyanophyta are extremely opportunistic, and are able to bloom in disturbed ecosystems [30]. Moreover, the dominance of cyanophyta in various microcosms could also be attributed to the production of toxins that inhibit the growth of other algae, therefore promoting their own proliferation [31,32].After one day of exposure in 8% and 16% (v/v) treated microcosms, protozoa density and species richness and composition significantly reduced relative to initial values and remained so for the remaining sampling periods (Fig. 3 and Fig. 4). The protozoan community structure in the 1.6% (v/v) microcosm was affected gradually with some recovery for the ciliates density. The inhibitory effects of Pilcam effluent on zooplankton density and species richness in all treated microcosms at all sampling periods, were most obvious for ciliates and to some extent, rotifers (< 28 days) whose biomass shifts followed a concentration-response pattern expected for single species tests. Dominant ciliates species present in untreated and 1.6% (v/v) microcosms and for short period of exposure in 8% and 16% (v/v) microcosms were in order of decreasing sensitivity to the industrial effluent as follows: Strombidium gyrans → Disematostoma sp. → Coleps hirtus → Uronema sp.The effect of Pilcam effluent on the chlorophyll A, microbial biomass (protein) content and alkaline phosphatase activity of untreated and treated microcosms. Significant effluent levels can be found in Table 3.The effect of Pilcam effluent on the phytoplancton density from untreated and treated microcosmsThe effect of Pilcam effluent on the protozoan density from untreated and treated microcosmsThe effect of Pilcam effluent on the protozoan species (A: Ciliates; and B: Rotifers) richness and composition from microcosms The effect of Pilcam effluent on benthic macro-invertebrates density of untreated and treated microcosmsThe effect of Pilcam effluent on aquatic snails’ density of untreated and treated microcosmsThe population dynamic of benthic macro-invertebrates under Pilcam effluent stress was similar to that of the protozoans. Their density in treated microcosms was in general reduced relative to the initial values after one day of exposure. The inhibitory effects of the industrial effluent at all sampling periods were most obvious for annelids whose biomass shifts followed a concentration-response pattern expected for single species tests and were significant in 8% and 16% (v/v) microcosms (Fig. 5). These benthic organisms were eliminated from the 16% (v/v) microcosm after one day of exposure. Snails were very sensitive to Pilcam effluent as they were all killed in 16% and 8% (v/v) microcosms after one and two days of exposure respectively (Fig. 6). These invertebrates’ organisms have been reported as being the first ones to be eliminated when the environment is saturated with heavy metals [3], especially with Hg and Zn that interfere with their respiratory function [8].The addition of Pilcam effluent adversely affected the water quality of test microcosms by causing a significant increase in conductivity, salinity, and total hardness. Industrial effluent concentrations of 8% and 16% (v/v) induced significant decreases in color, turbidity, total suspended solids, dissolved oxygen, and ammonia-N. All tested organisms, including fish, plankton, benthic macroinvertebrates and snails, were sensitive at various degrees to the industrial effluent, even at the lowest tested concentration of 1.6%. Further studies are needed to determine the no observable adverse effect level (NOAEL), as well as a chronic reference concentration for this effluent.This research was financially supported in part by the International Foundation for Science, Stockholm, Sweden, and the Organization for the Prohibition of Chemical Weapons, in the Hague, (the Netherlands), through a grant (No. C/2488-2) to Dr. Adolphe Monkiedje, in part by IRD AIRE Development Program through grant No. 01-02-CMR-31-1 to “Equipe Hydrobiologie”, and in part by “Fonds Universitaire d’Appui a la Recherche (FUAR)” of the University of Yaounde I, through a grant No. 99/5A7.
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+ All rights reserved.2,4-dichlorophenoxyacetic acid and monosodium methanearsonate are often sold in commercial mixtures. Bioconcentration studies have been performed for each of these herbicides individually, but little information exists concerning long-term exposure to a mixture of these herbicides. The following study examined the uptake of arsenic in crawfish after long-term exposure to this mixture, and the health risks associated with consumption of these crawfish. Bioconcentration and depuration experiments using a 50:50 by concentration mixture of the two herbicides, with and without surfactant, were performed to quantify how much arsenic is concentrated in the edible tissue of the crawfish. Of the three tissues (muscle, gill, and hepatopancreas) sampled hepatopancreas bioconcentrated the highest amount of arsenic. Surfactant significantly reduced this uptake but did not affect bioconcentration of arsenic into other tissues. Surfactant had no effect on depuration of arsenic from any of the tissues. Cooking lowered hepatopancreatic arsenic content, possibly as a result of structural changes in the hepatopancreas. Assessment of the human health risk associated with consuming these crawfish showed an exposure dose at the high end of consumption that was approximately twice the reference dose for arsenic. Cancer risks were averaged at approximately 7 extra tumors in a population of 10,000 and 6 extra tumors in a population of 10,000 resulting from a lifetime consumption of crawfish exposed to the herbicide mixture without and with surfactant, respectively.Both 2,4-dichlorophenoxyacetic acid and monosodium methanearsonate are used to control weed growth on public rights-of-way and in sugar cane, a major Louisiana money crop. These herbicides are often applied as a mixture to increase effectiveness to target vegetation. However, mixed herbicides may combine to form potentially harmful compounds or may enhance the toxic effects of each individual ingredient. One possible nontarget organism that may be affected is the red swamp crawfish Procambarus clarkii. Crawfish are an important food source in Louisiana, both to fishing industries and to recreational and subsistence fishermen, as well as an important part of the food chain for many native organisms [1,2].Data are available regarding the effects of these herbicides individually on red swamp crawfish, but there is no information available regarding the combined effects of these two herbicides on this species. The following study examined the uptake and excretion of a mixture of these herbicides in the red swamp crawfish, and addressed the question of whether consumption of exposed crawfish would produce higher or lower health risks than those recorded for single-herbicide exposures.Red swamp crawfish were obtained from the KJEAN Seafood Company of New Orleans, Louisiana. Crawfish were fed oatmeal three times per week and acclimated in holding tanks for 2 weeks. During the bioconcentration and depuration experiments, crawfish were allowed approximately ninety minutes three times weekly to eat oatmeal flakes. Any oatmeal left after feeding concluded was removed to prevent fouling of the water by accelerated bacterial growth.For the bioconcentration tests, 80 L of aerated and dechlorinated tap water per tank was dosed with one of three concentrations of 2,4-D dimethylamine salt (active ingredient 38.8% 2,4-D) and MSMA (active ingredient 46.33% As), or three 2,4-D/MSMA plus surfactant concentrations. The dosages of the herbicide mixtures were 0.342 mg/L, 0.684 mg/L, and 3.42 mg/L. The highest mixture concentration used was one percent of the 96-hour LC50 identified in an earlier series of acute toxicity tests [3]. The lowest mixture concentration was tenfold less than the highest concentration. A control tank of crawfish was also included in the assay [4].Circulation of the pesticide doses throughout each tank began three days before the addition of crawfish. On a daily basis, 25 L of water was siphoned from the lower portion of each tank and replaced with an equal amount of fresh dechlorinated water and the appropriate dose of the herbicide mixture. Effluent from the tanks was filtered through activated carbon to prevent the herbicides from entering the municipal water supply. On the first day of the bioconcentration experiment, one hundred seventy-five randomly chosen mixed-sex crawfish were placed in each test tank. Three crawfish were removed from each tank according to the following time schedule: 0 hours, 24 hours, 48 hours, 96 hours, and every fifth day thereafter up until the 47th day, at which point it was decided that the unusually high level of predation within the tanks necessitated an abbreviation of the bioconcentration assay. Ten milliliter water samples were taken from each tank on each sample day, before and after the water were refreshed, to confirm that no more than 20% variation for each herbicides occurred due to the replacement process [4]. Each of the sampled crawfish was dissected into muscle, gill, and hepatopancreas tissues. All samples were frozen until quantification of total arsenic could be performed. Crawfish mortality was recorded daily. At the end of the bioconcentration period, 3 control crawfish and 3 crawfish from the tank containing the 3.42 ppm mixture concentration were boiled for 20 minutes in one tablespoon of Zatarain’s crab and shrimp boil and 3 teaspoons of salt (as recommended in the Zatarain’s cooking instructions). Boiled crawfish were dissected as described earlier. These tissues and water samples from the boiling liquid were frozen for later analysis by ICP-MS for total arsenic quantification, as recently described in our laboratory [3].Crawfish remaining from the bioconcentration phase were transferred to fresh dechlorinated water to determine the rate at which they excreted 2,4-D and MSMA from their systems. The total volume of water in each tank was refreshed on a daily basis. Crawfish mortality was recorded daily. The planned length of the depuration assay was abbreviated due to high predation within the test tanks. Three crawfish were taken from each tank on days 3, 8, 22, and 50 (the final day of the depuration assay). Depuration tissue samples were analyzed using the same methodologies described for the bioaccumulation experiment [3].All samples were digested in a CEM MDS-2000 microwave to reduce interference from organic substances and to convert the arsenic to a form that could be analyzed by ICP. For microwave-assisted digestion of water samples, 9 mL of sample and 1 mL of concentrated HNO3 were initially heated to 160 ± 4°C or 70 psi in 10 minutes. For the second stage, the temperature of the samples was raised to 165 -170°C (or 85 psi) for 10 minutes [3].Each tissue to be analyzed was defrosted overnight under refrigeration. Aliquots of 0.5 g wet weight of defrosted tissues were microwaved with 9 mL HNO3; if the weight of the sample was less than 0.5 g ± 0.01g, the actual wet weight was recorded for use in calculating the inorganic arsenic concentration after ICP analysis. Samples were initially heated to 180°C in 12.5 minutes. For the second stage, samples were held at 180°C an additional 9.5 minutes. Each digested sample was diluted by 50% with distilled deionized water before ICP analysis to prevent acid damage to the ICP [3].Analysis of samples for arsenic concentration was performed using an Agilent Technologies 7500 series inductively coupled plasma mass spectrometer (ICP-MS) with Plasma chromatographic software. Since the molecular weight of arsenic (74.9216) is 46.33 % that of MSMA (161.7), the concentrations of arsenic obtained through ICP analysis was considered to be 46.33% of the actual concentration of MSMA. To correct for the dilution during microwave digestion, ICP results for water samples were multiplied by 1.11. ICP results for tissue samples were multiplied by a correction factor of 0.0526.Statistical analyses of total arsenic content in tissue samples from the bioconcentration/depuration experiments were performed using three-factor analysis of variance (ANOVA) run by the SAS PROC GLM program [5]. Significant differences were explored using the Newman-Keuls post hoc procedure. Time-specific comparisons were also performed to explore the dose by exposure time effect.Differences in arsenic concentrations in tissues from the crawfish boil experiment were analyzed for statistical significance using the Student-t test to determine whether boiling removed significant amounts of arsenic from crawfish muscle tissue [6]. The average doses of arsenic for a person eating crawfish exposed to the herbicide mixtures were calculated to assess the associated health risks. This dose was then compared to the reference dose for arsenic, the dose considered not to affect human health. The equation used for risk assessment was as follows:
2
+
3
+
4
+
5
+ where, X = the amount of bioconcentrated arsenic in crawfish tissue on the last day of bioconcentration;  Y = the average amount of crawfish flesh ingested per person per day;  Z = 70 kg for adult body weight, and 10 kg for child body weight.The Louisiana Crawfish Farmers Association reports a rule of thumb for a crawfish boil of 5 lbs of whole crawfish (shell and all) per person and approximately 1lb of tails (214.75 grams) for every 3 people when cooking crawfish etoufée [7], or 72 grams per person. This quoted average daily intake of tail meat per person was used as the high end of the range for the amount of crawfish consumed. The low end of the range for this variable was drawn from the conventionally accepted value of 33 grams per person of seafood (fish or shellfish) consumed daily. The risk or margin of exposure (MOE), for each treatment was calculated by dividing each treatment’s exposure dose by the accepted reference dose for arsenic.The cancer risk for a lifetime of ingestion of crawfish exposed to a 3.42 ppm mixture of 2,4-D and MSMA, with and without surfactant, was determined using the following equation:
6
+
7
+ Cancer Risk = Dose × Oral Slope Factor
8
+
9
+ in which Dose equals the exposure dose calculated in the systemic (non-cancer related) health risk evaluation, and the oral slope factor used was the EPA’s estimate of 1.5 (mg/kg/day)-1.Statistical analyses of inorganic arsenic content in tissue samples from the bioconcentration/depuration experiment yielded the results listed in Table 1. Statistical analysis of bioconcentration / depuration dataThe average inorganic arsenic concentrations detected in each tissue over total time of the bioaccumulation/depuration experiment are listed in Table 2. Calculated across all of the tissues over the total time of the assay, the arsenic tissue concentrations observed were 0.59 ppm for the control, 1.24 ppm for the 3.42 treatment, and 1.12 ppm for the 3.42 ppm + surfactant treatment. Arsenic concentrations measured in control tissues were significantly lower than those measured in the test groups (p < 0.01). The presence of surfactant was not observed to cause a significant difference in arsenic concentrations measured across tissues over the total time of the bioconcentration experiment. Average inorganic arsenic concentrations detected in each tissue over total time Table 3 lists the dose by tissue interactions of arsenic in sampled tissues. Hepatopancreas bioconcentrated the highest amounts of arsenic of the three tissues sampled, at 1.88 ppm and 2.24 ppm total arsenic for tissues exposed to the herbicide mixture with and without surfactant respectively. These concentrations were lower than those found by Abdelghani et al. [1] of 3.7 ppm arsenic at the long-term exposure to an MSMA-only concentration of 1.1 mg/L. A significantly greater amount of arsenic was bioconcentrated in the absence of surfactant (p < 0.01).Dose by tissue interactions of arsenic concentrationsGill tissue was observed to bioconcentrate significant amounts of arsenic (p < 0.01) in both treatments (0.75 ppm without surfactant and 0.93 ppm with surfactant), but its uptake of arsenic was not significantly affected by the presence of surfactant. Muscle tissue displayed no significant difference in long-term arsenic concentration between the control and the treatment with surfactant, with average concentrations over time of 0.32 ppm and 0.55 ppm, respectively. Surfactant was not observed to significantly affect the accumulation of arsenic in muscle tissue, with concentrations averaging at 0.55 ppm and 0.73 ppm, in the treatments with and without surfactant respectively. Muscle tissue sampled from the treatment alone did, however, contain significantly higher concentrations of arsenic (p < 0.01) than those sampled from the control group, at overall concentrations of 0.32 ppm and 0.73 ppm, respectively. Abdelghani et al [1] reported that after long-term exposure to MSMA, gill tissue displayed the greatest total arsenic loss of the three tissues under study (71%-78%). In contrast, gill tissue after long-term exposure to the 50:50 2,4-D/MSMA mixture displayed a total arsenic loss of approximately 41% (from 0.68 ppm to 0.40 ppm). Gill tissue after long-term exposure to the mixture with surfactant added displayed a total arsenic loss of approximately 46% (from 0.72 ppm to 0.39 ppm). Hepatopancreas tissue samples lost 63% of their arsenic content (from 3.21 ppm to 1.18 pm) in the absence of surfactant and 67% of their arsenic content (from 3.26 ppm to 1.08 ppm) in the presence of surfactant. Muscle tissues lost 50% of arsenic content (from 1.15 ppm to 0.58 ppm) in the absence of surfactant and 29% (from 0.68 ppm to 0.48 ppm) in the presence of surfactant. The order of loss based on percent of total arsenic depurated from the tissues in the absence of surfactant is as follows:
10
+
11
+ hepatopancreas > muscle > gills
12
+
13
+ and the order of loss based on percent of total arsenic depurated from the tissues in the presence of surfactant is as follows:
14
+
15
+ hepatopancreas > gills > muscle
16
+ This is in contrast to Abdelghani’s findings from a bioconcentration system with MSMA exposure alone:
17
+
18
+ gills > hepatopancreas > muscle
19
+ Average arsenic concentrations from the final day of bioconcentration to the final day of depuration are listed in Table 4 and Table 5. Statistical comparisons of the highest herbicide mixture treatments with and without surfactant showed a significant loss of arsenic across tissues from both treatments during the first 24 hours of depuration. Total arsenic concentration in tissues decreased from 1.68 ppm to 1.10 ppm in the mixture treatment and from 1.56 ppm to 0.97 ppm in the treatment with surfactant (both at p < 0.05). Significant differences in arsenic concentrations were also found between the last day of bioconcentration and the last day of depuration, with tissue arsenic decreasing from 1.68 ppm to 0.72 ppm in the mixture treatment and from 1.56 ppm to 0.65 ppm in the treatment with surfactant (both at p < 0.01).Tissue arsenic concentrations from final bioconcentration day to final depuration day, 3.42 ppm treatmentTotal arsenic concentrations from final bioconcentration day to final depuration day, 3.42 ppm + surfactant treatmentTable 6 lists the average arsenic concentrations in tissues used to examine the effects of cooking on arsenic content. Boiled muscle tissue from exposed crawfish contained a significantly higher amount of total arsenic than boiled muscle tissue from control crawfish (p < 0.003), at 0.76 ppm and 0.29 ppm respectively. Boiled treated muscle tissue did not, however, contain a significantly different concentration of arsenic than uncooked treated tissue (p ≤ 0.310), at 0.76 ppm and 1.15 ppm, respectively. The only significant difference in arsenic content between the uncooked treatment tissues and the boiled treatment tissues was found in the hepatopancreas tissue concentrations (p < 0.006), with total arsenic concentrations of 3.21 ppm and 0.58 ppm, respectively. Total arsenic content in boiled crawfish (in ppm)The gill tissues sampled for this assay were not observed to have bioconcentrated a significant amount of arsenic 0.5L (p < 0.605), which is contrary to findings from the bioaccumulation assay. This is probably due to random differences in the uptake of arsenic from crawfish to crawfish and to the fact that the sampling pool for the boiling assay was a relatively small one.Human health risks were calculated using arsenic tissue concentrations from the second to last day of bioconcentration sampling. Use of the risk assessment equation described in section 5.7 yielded the following results:
20
+
21
+
22
+ Herbicide mixture dose = 2.9 x 10-4 to 6.4 x 10-4 mg/kg/day ( an MOE of 1 to 2.1)
23
+
24
+
25
+ Herbicide mixture with surfactant dose = 2.3 x 10-4 to 5 x 10-4 mg/kg/day (an MOE of 0.77 to 1.6)
26
+
27
+ Herbicide mixture dose = 2.9 x 10-4 to 6.4 x 10-4 mg/kg/day ( an MOE of 1 to 2.1)Herbicide mixture with surfactant dose = 2.3 x 10-4 to 5 x 10-4 mg/kg/day (an MOE of 0.77 to 1.6)The averages of these doses are higher than the accepted reference dose for inorganic arsenic of 3 x 10-4 mg/kg/day. Both herbicide treatments yielded margins of exposure that were approximately twice the acceptable level of one. Estimated arsenic-related cancer risks for ingestion of crawfish that underwent long-term exposure to the 2,4-D/MSMA mixture were quantified at 4 - 10 extra tumors in a population of 10,000 over a lifetime consumption of crawfish exposed to the herbicide mixture, or 4 - 8 extra tumors in a population of 10,000 over a lifetime of consumption of crawfish exposed to the mixture plus surfactant. These cancer risks averaged to approximately 7 extra tumors in a population of 10,000 resulting from a lifetime consumption of crawfish exposed to the herbicide mixture and 6 extra tumors in a population of 10,000 resulting from a lifetime consumption of crawfish exposed to the herbicide mixture with surfactant. This method computes the 95% upper bound for the risk rather than the average risk, which results in there being a very good chance that the risk is actually lower. These calculated cancer risks are considered to be good within one order of magnitude; in other words, 10-4 may in actuality be 10-5Crawfish have an “open” circulatory system with arteries that eventually terminate after leaving the heart, allowing circulatory fluid, or hemolymph, to bathe the internal organs [8]. All dissected tissues were therefore in constant contact with the herbicide mixture introduced into the crawfish hemolymph through absorption from the gills [9]. Arsenic binds to the sulfhydryl groups of hemocytes in the hemolymph as it passes through the gills, which are directly exposed to the contaminate medium.The arsenic-laden hemolymph then moves through the hepatopancreas, where the bound metals are concentrated and sequestered to minimize toxicity [1,10]. The crawfish hepatopancreas serves in a variety of physiological processes, including digestion, absorption and storage of digested foods, detoxification, and storage of heavy metals [10,11,12]. In vertebrate hepatopancreas and liver tissue, arsenic induces production of metallothioneins, a class of low molecular weight proteins which bind metals such as arsenic, thereby rendering them unavailable to cause cellular damage [13,14]. In invertebrates, arsenic-induced metallothioneins do not actually bind arsenic [15]; instead, the primary method of arsenic sequestration in invertebrate hepatopancreas appears to be the formation of intracellular vacuoles [11,16]. Approximately 27% of the arsenic also binds to the lipid fraction of the hepatopancreas [1]. This multifaceted ability to concentrate and sequester arsenic explains why the hepatopancreas was not only observed to accumulate higher amounts of total arsenic than gill or muscle tissue but was also found to contain the highest concentrations of arsenic in control tissues.The significantly lower arsenic concentrations found in hepatopancreas samples exposed to the mixture with surfactant added may be due to the ability of surfactant to adsorb metal ions into precipitates of metal-surfactant. Surfactants added to a solution can render hydrophilic mineral surfaces hydrophobic through the formation of neutral metal-surfactant molecules [17]. This adsorptive activity led to arsenate removals of over 90% in the remediation studies of Lazaridis et al [18]. In comparison, the 1.88 ppm average total arsenic concentrated in the hepatopancreas exposed to surfactant in the bioconcentration experiment was 84% less than the 2.24 ppm average total arsenic concentrated in the hepatopancreas samples in the absence of surfactant.The gills are in direct contact with water and present a relatively large permeable surface for exchange of water-borne chemicals [8,19]. Oxygen consumption in gill tissue decreases in the presence of heavy metals. Respiratory stress may therefore affect the overall metabolic processes involved in the concentration and elimination of arsenic [8]. Gill tissue bioconcentrated significant amounts of arsenic, reaching an arsenic plateau around day 30 of the bioconcentration assay. The rest of the arsenic entering the gills would have been transported away as gill tissues were flushed by hemolymph [1].Crawfish abdominal muscle has consistently been found in literature to contain the lowest concentration of metals of all sampled crawfish tissues, a finding that was paralleled by the results from the bioconcentration assay. For example, Jorhem et al [20] reported a total arsenic bioconcentration of 0.18 g/g (ppm) in muscle tissue, 4.5 times less arsenic than the concentration they found in hepatopancreas (0.81 g/g (ppm)). In the 2,4-D/MSMA mixture bioconcentration assay, muscle bioconcentrated 3.1 times less arsenic than hepatopancreas (0.73 ppm versus 2.24 ppm). These findings are important in risk assessments to human health, since muscle tissue is the most often consumed portion of the crawfish. Composed of 81% water, crawfish abdominal muscle is likely to have fewer arsenic binding sites than the other tissues studied in this experiment [1]. The adsorption of arsenic by the surfactant may also have prevented a significant portion of arsenic from concentrating in muscle tissue.2,4-D can affect enzymatic activity during long-term exposure (Neskovic et al, 1996). It is therefore possible that the presence of 2,4-D in the herbicide mixture altered the ability of hepatopancreas to depurate stored products of MSMA. 2,4- D may also have decreased the ability of gill tissue to bioconcentrate MSMA, thereby reducing the amount of MSMA products available for depuration.The amount of arsenic bioconcentrated into muscle tissue was not significantly affected by boiling. This may be due to tight binding of arsenic at available sites within muscle tissue [1]. Muscle proteins are denatured during cooking, but this does not seem to significantly affect the proteins involved in arsenic sequestration.The significant difference in arsenic content between the uncooked exposed hepatopancreas and the boiled exposed hepatopancreas may have been due to the loss of hepatopancreatic lipids to the boiling medium. The hepatopancreas showed a dramatic alteration in size and consistency after boiling. The structural dissolution of this tissue would release arsenic bound to hemocytes and lipids and sequestered within intracellular vacuoles.Risk assessments for consumption of tissues exposed to both herbicide treatments yielded margins of exposure that were approximately twice the acceptable level. An unknown fraction of the total arsenic findings would actually be present in relatively nontoxic organic forms such as arsenobetaine [21]; therefore the risk assessments performed from this experiment may be misleading. Estimated arsenic-related cancer risks for ingestion of these crawfish yielded an average risk of 6 and 7 tumors in a population of 10,000; the accepted reference dose for arsenic yields a cancer risk of 4.5, or 5 extra tumors in a population of 10,000 over a lifetime’s consumption of exposed crawfish.One element not examined by this series of assays concerns the forms of arsenic that the bioaccumulated element might have been stored in. Different species of arsenic have different levels of toxicity [21,22]. For example, arsenobetaine, which is the major converted form of arsenic found in various marine animals, has been shown to be relatively non-toxic [21]. It would be valuable to determine how large a fraction of the arsenic bioconcentrated in this study was actually stored in this and other relatively nontoxic forms. Such information would give a clearer understanding of the actual risk involved in consuming these crawfish.
Med-MDPI/ijerph_1/ijerph-01-02-00132.txt ADDED
@@ -0,0 +1,7 @@
 
 
 
 
 
 
 
 
1
+ All rights reserved.The toxic effect of metal ions like chromium (Cr3+), cobalt (Co2+), nickel (Ni2+), copper (Cu2+), cadmium (Cd2+) and lead (Pb2+) on biochemical oxygen demand (BOD) of synthetic wastewater samples has been studied at different temperatures i.e., 15°C, 20°C, 25°C and 30°C. Experiments were conducted for BOD exertion in presence (10 ppm of each metal ion) and in the absence of metal ions at different temperatures. Transition metal ions like Cr3+, Co2+, Ni2+ and Cu2+ show an increase in relative percentage inhibition with increasing atomic number. BOD inhibition in presence of Cd2+ and Pb2+ is relatively large. The metal ions under study are found to be highly toxic to microbes.Since both the aerobic biological treatment process and BOD measuring technique (using dilution method) are based on the same principle, the presence of metals like copper, zinc, lead and other heavy metals will have an influence on both processes. Because of the limited solubility of oxygen in the aqueous medium in BOD bottle, the effect of heavy metals will be sufficiently large as compared to that in the treatment plant because the effluent in the treatment plant gets a regular supply of oxygen from continuous aeration.Trace quantities of heavy metals such as nickel, manganese, lead, chromium, cadmium, zinc, copper, ferrous and mercury are common constituents of most wastewaters. Some of these metals are necessary for growth of biological life and the absence of sufficient quantities of them could limit growth of algae. However, the presence of any of these metals in excessive quantities will interfere with many beneficial uses of water because of their toxicity. Nriagu [1] reported the effects of Hg, Cu, Zn, Cd and Pb on algae photosynthesis in the lakes. Zn and Cd inhibited the photosynthetic activity strongly in summer.Very little work is available in literature on the effect of temperature on heavy metal toxicity to BOD. Apparently, BOD is expected to increase with the increase in temperature of incubation because of the loss of dissolved oxygen from the medium of heavy metal ions. It will be quite interesting to see the effect on BOD particularly with an increase in temperatureIt has been reported that BOD5 is suppressed significantly by even small concentrations (1-2 mg/L) of copper or chromium. Stone [2] measured the percentage of suppression caused by 1 mg/L of selected heavy metals on the BOD of domestic sewage. Gray [3] found that 6 mg/L mercury chloride, 40 mg/L of copper sulphate or 30 mg/L of potassium dichromate were required to completely inhibit bacterial activity on a sample of glucose. Mayo [4] studied the effect of temperature and pH on the growth of micro organisms and has defined the optimum temperature and pH as those at which the growth rate of micro organisms is the highest.According to another study by Mittal and Ratra [5], the presence of metal ions like Al, Co, Ni, Cu, Zn, Pb and Hg in the effluent samples significantly affects BOD values. Metal ion addition results in the inhibition as well as increase in BOD, depending on its concentration.The present study was undertaken to determine the effect of heavy metal ions on BOD when the temperature of the medium is varied in the ambient temperature range. The reported work may have more relevance for common effluent treatment plants, and municipal wastes that contain both inorganic and organic effluents.The ambient temperature range (15oC-30oC with an incremental change of 5oC) was selected for the study, as a variety of microbes responsible for BOD exertion are optimally active in this temperature range. Experiments were conducted at selected temperatures of incubation to measure BOD in the absence and in the presence of some heavy metal ions. Five–day BOD (BOD5) measurements were used as assessment end points.Samples were prepared synthetically in the laboratory and stored in the refrigerator at 4°C. At the time of conducting the experiment, samples were first brought at the room temperature.Incubation bottles (300 mL capacity) were washed, rinsed thoroughly with distilled water and drained before use. As a precaution against drawing air in the dilution bottle during incubation, a water seal was used. To reduce evaporation of water from the seal, during incubation placed an aluminium foil cap over flared mouth of the bottle. Four BOD incubators thermostatically controlled at 15oC, 20oC, 25oC and 30oC (± 0.1oC) were used for the tests.All reagents were prepared as per procedures given in Standard Methods for the Examination of Water and Waste [6].A desired volume of distilled/deionised water was put in a suitable container, and 1 mL/litre of each of the phosphate buffer, MgSO4, CaCl2 and FeCl3 solutions were added to the seeded dilution water, as described in following paragraph. Dilution water was brought to room temperature before use, and saturated with dissolved oxygen by shaking for 15 minutes.A compatible seed from treatment plant of a local milk plant, Sirhind Road, Patiala was collected. The seed was taken from a corner of aeration tank and immediate before the settling tank. The effluent water (seed) was aerated and then allowed to settle. The supernatant liquid was then used as a “seed” for the BOD test. 1 mL of seed per litre of dilution water was added in nutritioned dilution water.BOD of seeding material was measured and dilutions were made so that its DO uptake comes between 0.6 to 1.25 mg/L.Synthetic waste sample was prepared from raw milk (Verka, Standard Shakti) in the laboratory. Synthetic sample was preferred to the actual waste samples from industrial effluents because the work is of relative nature and effect of toxicity was to be observed. Because industrial sample may contain matter other than organic material such as detergents, heavy metals, chemicals and other foreign materials, which may interfere along with the toxic metals under test. Raw milk has BOD about 70,000 to 100,000 mg/L and was diluted so that its BOD comes in the range of 700-1000 mg/L. For that 5 mL of raw milk was added in about 500 mL distilled water and diluted further to make it one litre.Metal ions selected for study were chromium, cobalt, nickel, copper, cadmium and lead. Appropriate amounts of their salts were dissolved in distilled water and dilutions were done to prepare a 1000-ppm stock solution of each metal ion.BOD5 measurement was done in presence of each one of all the six metal ions taken for the present study. Based on our previous study (Mittal and Ratra, 2000) a concentration of 10 mg/L was selected. For each metal concentration, three replicas were taken and the mean was used for subsequent calculations.1.5mL (1: 200) sample was taken in each bottle (300 mL capacity) so that the BOD lies between 700-1000 mg/L. Therefore, maximum expected DO depletion would have been between 3.5 to 5.0 mg/L. As the saturated DO in summer days is usually around 7.0 mg/L, a minimum of 1-2 mg/L of DO would have remained in the test solution even in case of maximum depletion, which is an essential norm for the BOD test.All the bottles were filled carefully with seeded dilution water to the top of the brim. Water was not allowed to overflow as it would disturb the concentration of metal ions in bottle. Initial DO of first bottle of each set was taken and remaining three were capped, water sealed and incubated for five days at 15oC, 20oC, 25oC and 30oC in BOD incubators. After five days incubation, final DO of samples was determined.Initial and final DO values were determined by titration method. Alsterberg modification of Wrinkler’s method was adopted. Percentage of inhibition in BOD values (due to toxicity of metal ions) were calculated for each metal ion at different temperatures, and are shown in Table 1.BOD Inhibition as a function of temperature in presence of different metal ionsA number of measurements were carried out for each experiment of BOD. The results are reported in Table 2 and the standard deviations for each measurement are given. Mean BOD value in presence of a given metal ion is subtracted from the mean BOD without the metal ion to calculate the inhibition. Histograms of the BOD without metal ion as well as those in presence of metal ion are plotted. Error bars for each histogram represent the standard deviation in the BOD values. Inhibitions in BOD due to the presence of metal ion are shown in the figure as separate columns. No error bars can be given for BOD inhibition as the results are the differences in mean BOD values.BOD exertion is affected by factors like temperature, availability of organic matter as food [7], seed [8,9], and pH of the medium [10]. Present study is undertaken to study the effect of some heavy metal ions like Cr3+, Co2+, Ni2+, Cu2+, Cd2+ and Pb2+ on BOD at different temperatures, i.e., 15°C, 20°C, 25°C and 30°C. Experiments were conducted for BOD exertion in presence of 10 ppm of each metal ion. This concentration was selected on the basis of our previous study [5] where an appreciable change in BOD was recorded due to the presence of 10ppm of the metal ion concentration. BOD5 was determined in the absence as well as in the presence of the metal ions.All metal ions are found to inhibit the BOD (Table 1). Absolute values of BOD in presence of metal ions are compared with those in the absence of metal ions at different temperatures and are shown in Figure 1, Figure 2, Figure 3 and Figure 4. These Figures also show inhibition in BOD in the presence of metal ions at a given temperature. Relative percentage inhibition increases with increase in temperature (Table 1) up to 25°C and further up to 30°C it tends to decrease, probably because of the relatively large increase in BOD exertion for the blank system from 25°C to 30°C (an increase of ~400mg/L as compared to ~100mg/L and ~140mg/L for temperatures 15°C to 20°C and 20°C to 25°C, respectively).In all the metal ions, relative percentage inhibition is more at 25°C than at all other temperatures. Lead is found to be the most toxic element at all temperatures. Cd2+ and Pb2+ show relatively larger inhibition than that of transition elements. It shows that transition metals as well as Cd2+ and Pb2+ are highly toxic to microbes.Table 2 shows that BOD in the absence of any metal ion increases from 857±15 mg/L at 15°C to 970±26 mg/L at 20°C. This increase of 100mg/L is relatively less as compared to the increases of ~150 and ~400mg/L for the increase in temperatures from 20°C to 25°C and 25°C to 30°C, respectively. This is due to the reason that upto 20°C, population of the nitrifying bacteria is not enough to contribute appreciably to the normal BOD5 exertion, but at 25°C and 30°C nitrogenous BOD adds up to the normal BOD. There are primarily two reasons for the increase in BOD from 15°C to 30°C;
2
+
3
+ Depletion of dissolved oxygen (mg/L) with increase in temperature. Solubility of oxygen in water at 760 mm of mercury pressure and 100% relative humidity are: 10.07 (15°C), 9.08 (20°C), 8.27 (25°C) and 7.59 (30°C).
4
+
5
+ Due to the toxic behaviour of metal ions towards microbes (metal ions complexing with microbial cells) resulting in a lower demand of dissolved oxygen.
6
+
7
+ Depletion of dissolved oxygen (mg/L) with increase in temperature. Solubility of oxygen in water at 760 mm of mercury pressure and 100% relative humidity are: 10.07 (15°C), 9.08 (20°C), 8.27 (25°C) and 7.59 (30°C).Due to the toxic behaviour of metal ions towards microbes (metal ions complexing with microbial cells) resulting in a lower demand of dissolved oxygen.A peer look into Table 1 shows that the relative percentage inhibition increases almost linearly with increase in temperature up to 25°C in the presence of each metal ion. It indicates that the metal microbe complex is stabilized with increase in temperature up to 25°C. For most of the metals, except copper, the relative percentage inhibition lies between 40% and 60%. Presence of lead in the BOD bottle leads to an exceptionally large inhibition with very little dependence on temperature.The recorded inhibition at 30° C is besides the effect of nitrifying bacteria. In their absence the inhibition would have been still larger. These bacteria are known to be most active in a temperature range 30°C - 40°C.BOD exertion as a function of temperature in presence of different metal ionsHistogram of BOD in presence of metal ions at 150CHistogram of BOD in presence of metal ions at 200CHistogram of BOD in presence of metal ions at 250CHistogram of BOD in presence of metal ions at 300CMeasurements of BOD can be influenced by the temperature of the solution by acting on properties of substances present like microbes, metal ions in the sample. Complexation phenomenons are strongly dependent on temperature. It is therefore recommended that BOD be measured for different temperature values.The authors are thankful to the Director, Thapar Institute of Engineering and Technology, Patiala for providing research facilities.APHA, AWWA and WPCF, Standard Methods for the Examination of Water and Wastewater, 20th Edition 2000, Jointly edited by Eaton, Andrew D.; Clesceri, Lenore S.; Greenberg, Arnold E.
Med-MDPI/ijerph_1/ijerph-02-01-00001.txt ADDED
@@ -0,0 +1 @@
 
 
1
+ Recent developments in biomedical sciences and environmental medicine have significantly improved our understanding of risk factors associated with human diseases and health disparities. Although poor access to health care, genetic, cultural, behavioral, and socio-economic factors have been linked to disease status, environmental determinants including physical, chemical, and biological factors have also been shown to play a significant role in human health. Environmental health problems are increasingly taking center stages. Environmentally-related illnesses including infectious diseases, and chronic conditions such as cancer, are continually inflicting new health burdens. As these challenges emerge in new and more complicated forms, it is essential that we keep pace with the scientific understanding that is critical in defining strategies for sound environmental health decision-making.In its summary recommendations for improving the scientific basis for environmental decision-making, the National Academy of Science pointed out that “90% of voters believe that the environment plays a significant role in health…yet environmental science and environmental health communities are too frequently independent of one another, funded by different agencies and consisting of different researchers. If these disciplines fail to push ahead collectively with further research and prevention, the many burdens of environmentally-influenced illness imposed upon our society may become even heavier”. In response to this call for action, Jackson State University (JSU) hosted the “First International Symposium on Recent Advances in Environmental Health Research” at the Marriott Hotel in Jackson, Mississippi, from September 19 through September 22, 2004. This important event was that first World Congress held in Jackson, MS, USA, on important issues related to environmental quality and human health. Its overarching objective was to promote interdisciplinary discussions and international scientific collaborations, as well as to advance the participants’ understanding of local, regional, and global environmental issues as they relate to the quality of life and human health.In an attempt to contribute global solutions to these environmental challenges, scientists around the world, have been more and more involved in bioenvironmental research, studying the toxic mechanisms of action of various environmental agents, developing new approaches for detecting or remedying environmental damage, identifying and characterizing genes involved in the manifestation of environmentally-related diseases, and providing the public and policy makers with scientific tools that are critical for environmental health decision-making. Therefore, the “First International Symposium on Recent Advances in Environmental Health Research” served as a strong forum for environmental and biomedical scientists-biologists, chemists, toxicologists, public health scientists, engineers, and policy makers interested in bringing about substantial contributions to addressing global environmental, and sustainable development issues, to communicate the latest advances in scientific research and new developments on critical environmental and human health topics including the following:New Frontiers in Environmental Health Research: The causes of most human diseases have been attributed to the complex interactions between genetic factors and environmental exposures. Hence, control and prevention measures highly rely on the understanding of the cause and effect relationships between these factors and disease development. In recent years, new areas of research such as toxicogenomics, proteomics, and functional genomics have emerged, with the aim of understanding molecular mechanisms of health and disease. Also, the recent advances in the molecular biology of the cell cycle regulation have given new life to our understanding of cancer in particular, and the idea that defects of regulation in cancer cells may partially explain successes that have been achieved in cancer chemotherapy. Specific areas of symposium research presentations included gene expression studies, proteomics, gene-environment interactions, functional genomics, biomarkers of effect, sensitivity and effect, signal transduction and gene activation; and molecular targets of disease chemotherapy.Environmental Toxicology and Health Risk Assessment: Growing public awareness of the potential risk to humans from toxic chemicals in the environment has generated demand for new and improved methods for toxicity assessment and rational means for estimating health risk. Many environmental agents such as metal ions, polycyclic aromatic hydrocarbons, pesticides/herbicides, UV-light, food additives, and viruses are known to induce various types of illnesses including cancer in humans. Several symposium presentations dealt with research elucidating the cellular and molecular mechanisms by which these environmental agents induce toxicity, mutagenesis, and carcinogenesis, as well as research on hazard assessment of exposure to physical, chemical and biological agents; dose-response evaluation and model development; exposure assessment analysis; and health risk characterization; and management.Emerging Topics in Computational Biology, and Environmental Modeling: Using of computational methods and procedures to investigate environmental and biological phenomena has made remarkable progresses. This field includes analysis of human genome data, prediction of DNA and protein structure and function, design of biomaterials and therapeutic agents, studies into small molecule-biomacromolecule interactions, and other related computational method development. Therefore, several symposium presentations dealt with the computational analysis of the physical and chemical properties of several environmental compounds, as well as on quantitative structure activity relationship (QSAR) studies for developing predictive toxicology models associated with exposure to these compounds.Health Disparities and Environmental Security: In recent years health disparities and biological and chemical terrorism have emerged as major issues in public safety and homeland security. With recent advances in laboratory technologies, it is often possible to measure specific genetic variations as risk factors for specific types of disease. Equally important is the evaluation of the role of modifier factors such as environmental exposures or other genes that may exacerbate the genetic risk leading to differences in disease susceptibility among individuals. Since the events of September 11, 2001 regarding the attacks on the World Trade Center and the Pentagon, and the subsequent anthrax attacks on several people, our collective thinking with regard to our vulnerability to terrorism has completely changed. The specific areas of research presentations included the following: health disparities and cancer; health disparities and heart disease; health disparities and infectious diseases; and bioterrorism/chemical terrorism.Medical Geology and Human Health: Recent concerns over health-related issues arising from exposure to environmental substances have raised substantial interest in a new field termed “medical geology”. In fact, naturally-occurring toxic metals such as arsenic, cadmium, lead, and mercury are now known to cause serious public health problems in several areas of the world. Likewise, the geographical distributions of several infectious diseases such as malaria, meningitis, and schistosomiasis, have been linked to intrinsic climatic and environmental factors. Research on this topic dealt with disease ecology, toxicology, pathology and/or epidemiology with regard to the emerging subject of medical geology.Natural Resources Damage Assessment and Management: Several environmental influences including natural and anthropogenic factors have been linked to ecosystem vulnerability. Monitoring and assessment data are therefore needed for science-based decision-making with regard to environmental management. Papers for presentation on this topic included those related to: a) conceptual modeling for ecological risk assessment, b) assessment of the physical, chemical, and biological characteristics of specific ecosystems, c) applications of GIS and remote sensing technology to environmental assessment and management, and d) bioindicators for environmental management.The symposium attracted more than 300 participants from 21 countries representing all five continents, and more than 150 scientific presentations across the disciplines of environmental health and biomedical sciences. As stated above, the scientific program was composed of six plenary sessions where oral/platform presentations were given by more than 40 invited speakers. In addition, there were two poster sessions – one for faculty and professional scientists, and one for students – with more than 100 abstracts. The submitted full length manuscripts were peer-reviewed, and selected for publication by experts in their respective fields. The accepted papers are being published in two volumes as special issues of the International Journal of Environmental Research and Public Health.We wish to extend special thanks to Dr. Sidney McNairy, Associate Director of the National Center for Research Resources at the National Institutes of Health for his vision and commitment to providing resources for the support of the RCMI Program activities including the RCMI-Center for Environmental Health at Jackson State University, and for serving as the First Distinguished Speaker for the Honorary Biomedical and Health Information Lecture Series” at the symposium. He made a distinguished lecture on the issue of human environment and health disparities, and pointed out the critical role that institutions like Jackson State University should play in addressing these issues. Thanks are also extended to all our conference presenters, session chairs, and keynote speakers, and especially congressman James E. Clyburn of South Carolina who spoke very elegantly of the critical issue of environmental justice. We are also grateful to all members of the planning and implementation committees for their significant contributions to the successful organization of the conference. Many thanks to Mrs. Zelma Leflore and her students for providing the technical assistance with the technology needed for platform presentations.Special thanks are extended to Dr. David Potter (Mississippi Commissioner of Higher Education), Dr. Ronald Mason, Jr. (President), Dr. Felix Okojie (Vice President for Research Development, and Federal Relations), Dr. Velvelyn Foster (Interim Provost and Vice-President for Academic Affairs), and Dr. Mary Myles (Director of Title III Program) for their administrative support. We would like to acknowledge the authors for their involvement and cooperation, and for their outstanding contributions to advancing science and sound decision-making in the critical area of environmental health sciences. Special thanks are also extended to all the peer-reviewers who took time off their busy schedules to carefully and critically review each of the manuscripts.On behalf of the entire organizing committee, the greatest acknowledgments go to our major symposium sponsors including the U.S. Department of Education Title III-Strengthening the Environmental Science Ph.D. Program at JSU, National Institutes of Health RCMI-Center for Environmental Health, JSU Office of Research Development and Federal Relations, and U.S. Environmental Protection Agency. The contribution of the National Library of Medicine (NLM) to the successful organization of the pre-symposium workshop on the “NLM’s Toxicology Network and Environmental Health Information Databases” is gratefully acknowledged.
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+ The environment plays a pivotal role as a human health determinant and presence of hazardous pollutants in the environment is often implicated in human disease. That pollutants cause human diseases however is often controversial because data connecting exposure to environmental hazards and human diseases are not well defined, except for some cancers and syndromes such as asthma. Understanding the complex nature of human-environment interactions and the role they play in determining the state of human health is one of the more compelling problems in public health. We are becoming more aware that the reductionist approach promulgated by current methods has not, and will not yield answers to the broad questions of population health risk analysis. If substantive applications of environment-gene interactions are to be made, it is important to move to a systems level approach, to take advantage of epidemiology and molecular genomic advances. Systems biology is the integration of genomics, transcriptomics, proteomics, and metabolomics together with computer technology approaches to elucidate environmentally caused disease in humans. We discuss the applications of environmental systems biology as a route to solution of environmental health problems.The relationship between the external environment and human health was recognized by ancient societies. The Greek physicians Alcmaeon of Croton and Hippocrates are credited with hypotheses linking environment and health [1]. In Roman times it was known that a source of potable water was necessary for human health, thus in addition to building aqueducts to supply necessary drinking water. Roman laws concerning public health were severe and strictly enforced [2]. Remnants of association between environment and disease survive to this day in some of the names associated with diseases. Malaria, for example, literally means “bad air”, which was associated with the onset of the disease. With the discovery that bacteria could cause disease, the Germ Theory of Disease was promulgated, largely from the work of Lister, Koch and Pasteur [3–6]. The germ theory recognized infectious agents of biological origin such as bacteria and viruses as the cause of much of human disease, subsequently leading to discovery of antibiotics that control bacteria and development of new regimens of immunization to control viral diseases [6–8]. Together with greater understanding of vector control and use of antibiotics and vaccines, the ability to manage diseases increased and the environment was largely overlooked as a causative agent of human disease.With the elucidation of the structure of DNA in the early 1950’s and the growth of molecular biology, the genetic basis of non-infectious diseases blossomed, and great emphasis on genetics as a cause of diseases was emphasized in medicine [9–12]. In fact, chronic diseases for which no specific cause was known were largely attributed to genetics or even “bad genes” [13].Awareness of the environment as an agent that affects human health gained momentum with publication of some popular press books, notably Silent Spring[14]. Incidents such as that which occurred at Love Canal, inspired the environmental movement, and government action and research into the environment and disease. The creation of the Environmental Protection Agency (EPA) and the National Institute of Environmental Health Sciences (NIEHS), an institute of the National Institutes of Health (NIH) [15, 16] focused on government sponsored environmental health research. The presence of hazardous pollutants in the environment is now often implicated in human disease [17]. That pollutants cause human diseases however is often controversial because data connecting exposure to environmental hazards and human diseases are not well defined, except for some cancers and syndromes [18].The complex nature of human-environment interactions and the role those interactions play in determining the state of human health are becoming more appreciated [19]. Observational epidemiology studies undertaken to assess potential causal relationships between exposure and human health are limited because excess disease occurrence is often small and difficult to identify [20, 21].The Human Genome Project was undertaken as an international collaboration to sequence the entire human genome [22]. It was found that the human genome consists of between 20,000 to 25,000 genes, 3 billion base pairs, and that about 99.9 % of which are identical in human populations [23]. It has been estimated that approximately 1,200 genes are responsible for about 1,600 diseases [24]. The “genome” was originally defined by a German botanist; Hans Winkler in the 1920’s to refer to all genes within a set of chromosomes [25]. The term was expanded to mean all DNA in chromosomes, because it was found that genes comprise only 2 to 3 percent of the human genome [25]. Sequencing the human genome is the most ambitious and important effort in the history of biology. It was thought that through sequencing the entire human genome a complete genetic blueprint for human life would be provided, which would yield important insights into human health and development [26]. The genome sequence has provided many tools for researchers to ask questions that were not addressable before the human genome project. While there is hope for improved medical care and public health resulting from the advances made by the human genome project, the genome sequence is not yet used as widely in public health or medical practice as it is in research.It was quickly realized that the sequence of the genome alone was not going to yield all the answers, thus we quickly entered the post-genomic age, which focuses not only on the study of the genome, but also on products of the genome, which essentially follows the central dogma of molecular biology proposed by Watson and Crick more than 50 years ago [27], with the addition of enzymes and metabolism: (Figure 1). Thus, the genome (all DNA) gives rise to the transcriptome (all messenger RNA; mRNA), the proteome (all proteins in a cell, including enzymes) and the metabolome (all metabolites and enzymes that generate metabolites) in the cell.The human genome project yielded huge data sets containing large numbers of DNA sequences stored and being analyzed on computers all over the world. These data are being sorted, annotated and developed in various ways using computer software to organize integrated maps of DNA involving genetic and physical information [28]. Recognition of the need to be able to handle large data sets came early when GenBank was established in the mid 1960’s [29, 30]. This marriage of biology and computer technology led to the emergence of the new science of bioinformatics.As the picture of environmentally-caused diseases continues to emerge, we are gaining a greater appreciation that it is the interaction of the environment with our genes that leads to most disease states in humans. Sequencing the human genome served to underscore this. Understanding risks to human health in light of the human genome-environment interaction is one of the more compelling challenges in environmental public health [31, 32]. With approximately 99.9 % of human genomes being identical, the remaining 0.1% (or about 3 million base pairs) appears to dictate differences in susceptibility to environmental challenges among human populations. As a result, much research has focused on single nucleotide polymorphisms (SNPs), which are stable heritable changes abundant in the genome, as the source of human variation [33]. We are learning that it is not as simple as a single SNP alone, but rather it is differences in patterns of SNP polymorphisms, called haplotypes, that may be at least partly responsible for differences in susceptibility to environmental conditions of human populations [32, 34, 35]. Active research to elucidate haplotype maps and patterns among different population groups is currently underway [36]. Haplotype mapping and pattern recognition is a potentially powerful tool to identify populations at risk for environmentally caused diseases. Thus certain SNPs or groups of SNPs (haplotype) confer susceptibility of individuals in a population to disease [37].Because of our increased knowledge of genetics and genomics it is now apparent that most diseases are not carried in our genes as deterministic factors of disease, but rather our genomes carry variations in populations that result in differences in susceptibility to disease. So, with the sequencing of the human genome, renewed interest in understanding the role of the environment as a cause of human disease has occurred [38]. Genes are expressed in response to the environment. Thus, when individuals in a population carry variations in the genome that results in altered expression of certain genes, disease results in susceptible populations [39, 40].Even with availability of large sets of sequence data and genomic information, it is not yet possible to determine the role that exposure to the environment plays in affecting health outcomes such as birth defects, developmental deficiencies, chronic respiratory disease, multiple sclerosis, Parkinson’s or Alzheimer’s disease [41]. The term toxic genomics has been applied to the study of gene-environment interactions [42]. However that term is self-limiting to consideration of pollutant chemicals and does not embrace the concept that the environment encompasses more than pollutant toxicants.We use the term “Enviromics” to mean interactions of the complete environment, or envirome, with human genomes (Figure 2) [43]. The envirome encompasses every interaction between humans and the external environment. It includes where we live, what we eat, drink, or breathe, our social economic status, behavior, social interactions, occupation, and exposure to pollutants. The concept of the enviromics is all encompassing in its scope and understanding how the envirome affects human health, both positively and negatively. To gain a full understanding of these interactions, new tools and approaches must be developed. The science of genetics has been a powerful tool in environmental public health practice to identify rare conditions and syndromes, chromosomal aberrations birth defects, inborn errors in metabolism and reproductive errors, and as a tool for genetic counseling [44]. Genetics however is a linear science, which examines single genes, one at a time. A multidimensional approach is required to derive a more accurate assessment of the dynamic processes associated with living systems.Genomics looks at all the genes as a dynamic system, over time, to determine how they interact and influence biological pathways, networks and physiology, in a much more global sense than genetics. Thus, genomics shows great promise for identifying groups of genes involved in complex disorders to understand and intervene in environmentally caused diseases [45].When considering environment-genome interactions as a factor in complex disease, we understand that the genome cannot be changed, at least for now. However, once identified, it is possible to reduce exposure or modify the lifestyle element that is the environmental factor in the disease [46, 47]. Gene-envirome interactions can occur by direct interactions with active metabolites at specific sites of the genome to yield mutations, which could result in a human disease [48]. Indirect interactions with the human genome can occur via intracellular receptors that act as ligand-actived transcription factors, which regulate gene expression maintaining cellular homeostasis, or with an environmental agent to cause harmful effects (Figure 3) [49]. This type of envirome-gene interaction may be more easily examined than direct interaction because markers of this type of interaction are numerous and easily measured before onset of disease. Some examples of this include expression of cytochrome P450 genes after exposure to environmental agents, such as the polyaromatic compound benzo[a]pyrene, that bind to the Ah receptor [50–52]. Epigenomic change brought about by exposure to environmental agents is another important example of indirect environment-gene interaction [53, 54]. These changes, which are not considered mutations, result in silencing or enhancing specific gene expression by hyper-or hypo-alkylation processes.Our ability to measure envirome-gene interactions has exceeded our understanding of the mechanisms of envirome-disease linkages. Current approaches to understanding risk to human health after environmental exposure are based on studies of single chemical exposure and limited health effect, or single gene-environment interactions [55, 56]. We are becoming more aware that the reductionist approach promulgated by traditional research methodology has not, and will not yield answers to the broad and most important questions of population health risk analysis [57].The question most people have is “will the environment adversely affect my family’s health?” This is obviously not an easy question to answer. There are many common chronic diseases for which we do not have a clear understanding of causes, etiology, gene involvement, or susceptibility and we certainly do not have causal links [58]. These diseases are ones which are common in our society, including asthma, prostate and breast cancer, autism, Parkinson’s disease, Crohn’s disease, or diabetes. In addition, we lack knowledge of the molecular mechanisms of pathology of diseases caused by exposure to lead, mercury, or pesticides of various kinds. This is true in spite of a large body of research to try to pick apart those diseases and exposures. We have not really progressed to the point that we have detailed knowledge of how genes are involved or what processes and pathways influence individual susceptibility to disease after interaction with the envirome. This is a result of using a reductionist approach to piece together the larger picture one component at a time. We need an integrated approach that draws on data from the environment, biomarkers of exposure, gene expression patterns and parameters, and physiology, for public health practice to benefit from modern genomics technology [59]. Systems biology is an emerging science that integrates genomics, transcriptomics, proteomics, and metabolomics together with computer analysis and modeling to understand interacting gene networks that maintain cellular homeostasis. Because of the unique problems we face in environmental health, environmental system biology teams must include environmental anthropologists and sociologists, exposure assessors, epidemiologists, ecologists as well as toxicologists, molecular scientists, computer modelers and statisticians. Systems biology can thus can be applied to the understanding how the envirome can modulate the tightly regulated circuitry of the human organism to cause disease in the broadest sense [60].That cells and organisms have interconnected pathways that regulate metabolism is well known and reflected in the metabolic pathways found in every textbook of biochemistry. Similarly, signal transduction networks are becoming better understood. However, understanding the complex gene regulation networks expressed in the transcriptome, proteome and metabolome downstream of the signal transduction pathways is much more complex [61]. Gene array technology together with computers for statistical analysis and modeling techniques has been used to establish gene networks (see Figure 3) [62]. Proteomics and metabolomics are more complex than genome analysis and have lagged in application to environmental health; however the development of protein chips and other analytical advances will result in exponential growth in those fields [63].The recognition that using gene array technology can elucidate genomic and envirome factors in understanding human health and disease are a focal point in modern environmental public health [64]. We will soon be in a position to organize data components into modules amenable to systems biology approaches to modeling of environmental disease. Thus data on environment, exposure, and gene networks that describe the transcriptome, proteome and metabolome will provide insights into the identity and character of genome-envirome interactions, giving us opportunities to effectively target intervention strategies. Complex databases of genome sequences from genomic and toxicant information combined with modern methods of data mining, information retrieval and statistics will provide comparative information on the molecular basis of toxicity and disease.The science underlying genomic and system biology approaches to environmental diseases is readily available. However, application of these powerful methods is lagging, in part because at first glance, genomics and public health practice are at polar opposites. Public health is practical and utilitarian, where the rights of the majority out weigh the rights of the minority, resulting in interventions that can be perceived as coercive. For example general immunization and isolation or quarantine has been justified over individual civil rights to protect general health of the population [65]. On the other hand, using systems biology to identify susceptibility to environmental diseases other is highly personalized [66]. There is no guarantee that individual findings will be generalizeable to the population at large, consequently, there is potential for clashes between public health and new genomics approaches [67]. Another major concern includes, ethical, legal and social issues regarding the accumulation and proper application of the data derived from such studies [68–70]. These points will have to be addressed before modern genomic approaches can be widely accepted in the practice of environmental public health.Human population studies using clinical or epidemiological data that associate environmental exposures with health endpoints and disease can now be studied using systems biology approaches incorporating enviromics, and metabolomics. Together with the use of population genetic histories, understanding human genetic variation and genomic reactions to specific environmental exposures will allow us to uncover the causes of variations in human response to environmental exposures providing important new tools in assessing risk of human disease [31].Central Dogma of Biology: Modern-omics technologies follow the pattern established by the central dogma of biology proposed more than 50 years ago by Watson and Crick [27], with the addition of active enzymes and metabolities, which taken together reflect human phenotypes. Here we include enzymes as part of the metabolome because metabolities are regulated by enzyme patterns.Indirect Environment-Gene Interaction: Hormones and vitamins interact with the genome via ligand-activated transcription factors yielding a “normal” cellular response to maintain homeostasis. Environmental agents can mimic natural ligands or bind to other intracellular receptors that yield different information from homeostatic regulation. The result is an altered cellular response yielding an adverse health effect.Simple Interaction Gene Regulatory Network: In the simple model, three interacting genes form a network in a cell. Here Gene A activates Gene B. Gene B activates Gene A and Gene C, and Gene C inactivates Gene A. Thus several levels of regulation are possible with the three interacting genes.Supported in part by grants from National Science Foundation (0234143), National Institute of Environmental Health Science (ES-10956 and ES-013379), Department of Energy (DE-FC26-00NT40843), and Centers for Disease Control and Prevention.
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+ Although nickel and cobalt compounds have been known to cause induction of the transcription factor hypoxia-inducible factor 1 (HIF-1) and activation of a battery of hypoxia-inducible genes in the cell, the molecular mechanisms of this induction remain unclear. The post-translational modification of HIF-1a, the oxygen-sensitive subunit of HIF-1, regulates stabilization, nuclear translocation, DNA binding activity, and transcriptional activity of the protein. Among the enzymes regulating the post-translational modification of HIF-1a, the factor inhibiting HIF-1 (FIH-1) hydroxylates the protein at asparagine 803, suppressing the interaction of HIF-1a with transcription coactivators p300/CBP and reducing the transcriptional activity of the protein. ARD-1, the acetyltransferase, acetylates HIF-1a at lysine 532, which enhances the interaction of HIF-1a with pVHL. Therefore, FIH-1 and ARD-1 negatively regulate the transcriptional activity and the stability of HIF-1a. We examined the mRNA levels of FIH-1 and ARD-1 genes after exposure nickel (II) or cobalt (II) to the cell and found that both genes were down-regulated by the chemical treatment, which may lead to reduced levels of both proteins and result in increased level of HIF-1a and its transcriptional activity.According to the International Agency for Research on Cancer (IARC 1990), both soluble and insoluble nickel compounds have long been established as human and animal carcinogens [1]. And cobalt compounds are carcinogenic in animals [2]. Epidemiological studies have shown that environmental or occupational exposure to nickel or cobalt compounds could cause lung and nasal cancers, asthma, fibrosis, pneumonitis and some other lung injuries [1–5]. Despite the various differences among the compounds of these two metals regarding the biochemical and molecular mechanisms of their toxicity and carcinogenicity, they mimic hypoxia to induce the HIF-1a transcription factor and hypoxia-inducible genes [5, 6], which are believed to play important roles in carcinogenesis. However, the complete mechanisms by which nickel and cobalt compounds induce HIF-1a are still unknown, although several sites of their impact on HIF-1a have been described [7–9].The transcription factor hypoxia-inducible factor 1 (HIF-1) plays an essential role in cellular oxygen homeostasis [10, 11]. HIF-1 is a heterodimeric complex composed of alpha and beta two subunits [12]. The beta subunit is also the heterodimerization partner for the aryl hydrocarbon receptor (AhR) and thereby called aryl hydrocarbon receptor nuclear translocator (ARNT) that is constitutively expressed; whereas the alpha subunit (HIF-1a) is highly oxygen-sensitive and is rarely detectable under normal oxygen tension but is dramatically induced with hypoxia [12]. Under reduced oxygen tension, HIF-1a is stabilized and translocates to the nucleus, where it dimerizes with ARNT. Then the active HIF-1 stimulates the transcription of genes involved in angiogenesis, cell survival, glucose transport and metabolism [13, 14].HIF-1a is regulated by a reduced oxygen level largely at its post-translational modifications, resulting in stabilization, nuclear translocation, DNA binding activity, and transcriptional activity of the protein. The post-translational modifications of HIF-1a include prolyl hydroxylation at proline 402 and 564 within the oxygen-dependent degradation (ODD) domain by HIF-prolyl hydroxylases (HPHs) [9, 15–17], asparaginyl hydroxylation at asparagine 803 in the C-terminal activation domain (C-TAD) by factor inhibiting HIF-1 (FIH-1) [8, 18, 19], acetylation of lysine 532 in the ODD domain by an acetyltransferase ARD-1[20], phosphorylation induced by p42/p44 mitogen-activated protein kinase (MAPK) activity [21], as well as ubiquitination by the von Hippel-lindau (pVHL) complex [22–25].In normoxia, proline 402 and 564 of HIF-1a are hydroxylated, which is required for the binding of pVHL complex and leads to the ubiquitination of the protein, resulting in targeting of HIF-1a for proteasomal degradation [15–17, 24]. The interaction of HIF-1a with pVHL is enhanced by ARD-1-mediated acetylation at lysine 532 [20]. The acetylation of this lysine residue by ARD-1 is critical to the proteasomal degradation of HIF-1a since a mutant with arginine substituting lysine 532 shown no acetylation by ARD-1 was stabilized and had a decreased interaction with pVHL [20]. At the same time, hydroxylation of asparagine 803 during normoxia suppresses interaction of HIF-1a CAD with transcription coactivators p300/CBP and reduces the transcriptional activity of the protein [18, 26]. In addition to interacting with HIF-1a, the asparaginyl hydroxylase FIH-1 also interacts with pVHL, allowing the formation of complexes containing HIF-1a, FIH-1, and pVHL [27].In hypoxia, decreased level of prolyl hydroxylation due to the limiting oxygen prevents pVHL binding to HIF-1a, resulting in rapid accumulation of HIF-1a protein [15, 16, 24]. The acetylation level of HIF-1 gradually decreases as the length of hypoxic exposure time increases, which is due to the reduced expression of ARD-1[20]. Meanwhile, stabilized HIF-1a protein is able to bind to p300/CBP to execute its transcriptional activity due to decreased level of hydroxylation at asparagine 803[18, 19]. In addition, during hypoxia, p42/p44 MAPK activity induces phosphorylation of HIF-1a and promotes its transcriptional activity [21]. As a consequence, HIF-1a accumulates and promotes hypoxic tolerance by activating gene transcription.Among the enzymes that regulate HIF-1a, prolyl hydroxylases (HPHs) and asparaginyl hydroxylase (FIH-1) belong to the family of 2-oxoglutarate-dependent dioxygenases and require Fe2+, 2-oxoglutarate, O2, and ascorbate for their reactions [8, 15, 16]. ARD-1 acetyltransferase acetylates HIF-1a by transferring an acetyl group from acetyl-CoA [20].Nickel (II) and cobalt (II) exposure in the presence of oxygen causes accumulation of HIF-1a protein and induction of its transcriptional activity [28–30], in part resulting from the inability of HIF-1a binding to pVHL due to the inhibition of HIF-1a hydroxylation by the metals [24, 31]. Furthermore, Cobalt (II) has been shown to inhibit activities of recombinant asparaginyl and prolyl hydroxylase in vitro [8, 9].In this study, we have examined the effects of nickel (II) and cobalt (II) on the gene expression of FIH-1 and ARD-1 using reverse transcriptase PCR (RT-PCR). Both genes were down-regulated in the metal-exposed cells, which might lead to reduced level of both proteins and result in increased level of HIF-1a and its transcriptional activity.Human lung adenocarcinoma A549 cells (CCL185) were purchased from American Type Culture Collection (Manassas, VA). Cells were maintained in F-12K medium (Life Technologies, Inc., Gaithersburg, MD) supplemented with 10% fetal bovine serum and 1% penicillin/streptomycin (equivalent to 100 units/ml and 100 μg/ml, respectively) at 37°C as monolayers in a humidified atmosphere containing 5% CO2.Nickel chloride hexahydrate and cobalt chloride hexahydrate were purchased from Sigma (St. Louis, MO). Cells were seeded into 100-mm dishes and allowed to attach overnight. When cells reached 70–80% confluence, nickel chloride hexahydrate (0.5 mM, 1.0mM), or cobalt chloride hexahydrate (0.2mM, 0.4 mM) was added to the medium for 24h.At the end of the treatment, total RNA was isolated from the cell using the Trizol reagent (Invitrogen, Carlsbad, CA). The mRNA was isolated using Oligotex kit (Qiagen, Germany). Reverse transcription was carried out with SuperScript first-Strand Synthesis System for RT–PCR (Invitrogen) and 1 μg of mRNA was used for first-strand cDNA synthesis according to the manufacturer’s protocol. The PCR was carried out in a total volume of 50μl and 1μl of first-strand cDNA was used for amplifying genes. The primers used for the PCRs were: FIH-1 forward primer, 5′-GCCAGCACCCACAAGTTCTT-3′; FIH-1 reverse primer, 5′-CCTGTTGGACCTCGGCTTAA-3′ ARD-1 forward primer, 5′-TGGGGTGAGGAGGGGATGG-3′; ARD-1 reverse primer, 5′-GGGAAGATTGTGGGGTATG-3′. Primers for amplifying GAPDH gene were purchased from BD Biosciences Clontech. Amplification conditions for FIH-1 and ARD-1 genes were 95°C for 2 min, 22 cycles at 95°C for 45 s, 57°C for 45 s, 72°C for 1 min, and 72°C for 5 min; the same for GAPDH gene except that 16 cycles were performed. PCR products were then resolved on 1.5% agarose gels containing Ethidium bromide.The investigation on the mechanisms of the induction of HIF-1a by nickel (II) and cobalt (II) has been focused on the effects of the metals on the enzyme activities of HPHs and FIH-1. Besides the possible effects of the metals on the enzyme activities, nickel (II) and cobalt (II) might affect the gene expression of the enzymes, so that the levels of the enzymes in the cell could be affected. Therefore, we examined mRNA levels of FIH-1 and ARD-1 by exposure A549 cells to nickel (II) and cobalt (II). As shown in figure 1, nickel (II) and cobalt (II) decreased FIH-1 and ARD-1 mRNA levels in a dose-dependent manner. GAPDH gene served as an internal control. The house keep gene 60S acidic ribosomal protein has also been used for measuring the loading control, which gave the same pattern as that of GAPDH gene (data not shown).Nickel (II) and cobalt (II) have long been known to induce hypoxia-like stress by activating HIF-1a [28–30]. However, the mechanisms of this induction are still unclear, although several studies have shown effects at sites of HIF-1a regulation, particularly focusing on the effects of the metals on the hydroxylases activities[7–9]. Since the HIF-1a hydroxylases require iron (II) as a cofactor for their activities and iron, cobalt, and nickel are adjacent in the transition metal group, it has been suggested that nickel (II) and cobalt (II) may substitute for the iron (II) in the hydroxylases, thereby, causing the loss of the enzymatic activity[7]. Unfortunately, there is no direct evidence to support this hypothesis. Recently, the cellular ascorbate, another co-factor required for the hydroxylases activities, has been shown to be depleted by both metals. Since the role of ascorbate is maintaining iron in its reduced state (iron II), this depletion may favor enzyme-bound iron oxidation, which may lead to the inactivation of the hydroxylases[31].Besides their effects, direct or indirect, on the hydroxylases activities, nickel (II) and cobalt (II) are very likely to induce HIF-1a at additional sites. We demonstrated here that the mRNA levels of FIH-1 and ARD-1 genes were down-regulated by both metals, which could result in reduced levels of the protein products of these genes. Since both FIH-1 and ARD-1 proteins negatively regulate HIF-1a, decreasing of them would lead to accumulation of HIF-1a and increase of its transcriptional activity.Furthermore, we suspect that nickel (II) or cobalt (II) may affect the acetylation of HIF-1a not only by down-regulating the acetyltransferase ARD-1, but also by decreasing the level of the cellular acetyl-CoA, the supply of the acetyl group, by the way of shutting down the Kreb cycle as well as increasing the ratio of NAD/NADH thereby causing the deacetylation of histones. Since both nickel and cobalt have previously been shown to inhibit histone acetylation and increase the methylation of histone H3 at lysine 9, this effect may play a role in the down-regulation of these two important genes which impact negatively upon HIF-1a activation (Costa et al. Mutation Res. in press).Our study revealed a possible new mechanism of the nickel- and cobalt- induction of hypoxia-like stress in the cell, contributing insight to better understanding of their carcinogenesis.Nickel (II) and Cobalt (II) down-regulate mRNA levels of FIH-1 and ARD-1genes in A549 cells A549 cells were treated with NiCl2·6H2O (0.5mM and 1 mM) or CoCl2·6H2O (0.2 mM and 0.4mM) for 24 h. The cells were harvested at the end of the treatment and the mRNA was isolated. The FIH-1, ARD-1, and GAPDH genes were amplified by RT-PCR.
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+ L1 and Alu elements are among the most active retroposons (mobile elements) in the human genome. Several human diseases, including certain forms of breast cancer and leukemia, are associated with L1 and Alu insertions in functionally important areas of the genome. We present data demonstrating that environmental pollutants, such as heavy metals, can stimulate L1 retrotransposition in a tissue culture system using two different types of assays. The response to these agents was equivalent when using a cell line with a stably integrated L1 vector (genomic) or a by introducing the L1 vector by transient transfection (episomal) of the cell. Reproducible results showed that mercury (HgS), cadmium (CdS), and nickel (NiO) increase the activity of L1 by an average of three (3) fold p<0.001. This observation is the first to link several carcinogenic agents with the increased retrotransposition activity of L1 as an alternate mechanism of generating genomic instability contributing to the process of carcinogenesis. Our results demonstrate that mobile element activation must be considered as one of the mechanisms when evaluating genomic damage/instability in response to environmental agents.LINEs (LINE-1s or L1s) are long interspersed repeated elements with the capability of generating new copies that insert throughout the genome. LINE amplification has been highly successful through evolution contributing approximately 17% of the human genome [1]. In addition, it is believed that LINEs have also been responsible for the 11% of the genome made up of Alu elements [2]. L1 retrotransposition also can serve as a vehicle to mobilize non-L1 sequences such as exons or promoters into existing genes [3]. Overall, LINE activity has greatly contributed to the evolution of the human genome.From a clinical view point, there are several reported examples of diseases caused by L1 insertions, including muscular dystrophy [4] and hemophilia A [5]. Mobility of L1 and Alu elements has also been shown to cause cancer, probably through somatic mutations [6, 7]. A variety of reports present data suggesting that increased activity of mobile elements may be involved in neoplastic progression [8]. For example, rodent LINE expression is increased in tumors [9, 10]. In humans, L1 expression is increased in breast cancers [11] and testicular cancers [12]. Yet, to date, little is known about the environmental factors that can influence human retrotransposition. Until relatively recently, it was practically impossible to measure retrotransposition activity in mammals, particularly at the somatic level. There are a number of examples of stress-inducing factors that influence the expression and rate of transposition for an assortment of elements in various organisms [13–16]. Two reports evaluate different and environmental factors on L1 expression using an L1 promoter-luciferase assay, providing an initial indication that L1 may respond to external factors [17, 18]. However, this method only evaluates L1 promoter response in an artificial setting that can not accurately evaluate the retrotranspositional activity of an element. There is one report on the stimulation of Alu retrotransposition by genotoxic compounds, suggesting that these types of mobile elements can be influenced by the exposure of exogenous agents [15].In 1996, the Kazazian laboratory designed a genetically marked L1 vector such that the selectable marker would be activated only following an RNA-mediated retrotransposition event [19]. This development immediately presented the application of this L1-vector in quantifying mammalian retrotransposition using tissue culture. We utilized this L1 insertion assay to create a model system and evaluate selected environmental pollutants, specifically heavy metals. Two approaches were utilized: one using a cell line with a stably integrated copy of the L1-vector that more accurately reflects the “natural” state of an L1 element, and the other using the transient introduction of the vector into the cells. We show that both approaches are useful in the evaluation of L1 activity after the exposure to environmental agents.Metal pollutants pose a threat of toxicity to both humans and wildlife because of their wide distribution in the environment and workplace and their high persistence [20]. Humans get exposed to these heavy metals from numerous sources including contaminated air, water, soil and food. Exposure to heavy metals is not uncommon; in particular people who smoke, consistently expose their lungs to several of these metals. Reports indicate that epidemiological data predict that 1 to 18 lung cancer deaths/10,000 smokers may be attributed to inhaled cadmium in cigarette smoke [21]. Several heavy metals like cadmium and mercury are on the EPA’s list of extremely hazardous substances hazardous chemicals: [http://yosemite.epa.gov/oswer/ceppoehs.nsf/CAS]. Among their many hazardous effects, cadmium is recognized as a carcinogen and a teratogen, while mercury is classified as a neurotoxin (reviewed by [22]). Although both nickel and cadmium are classified as carcinogens, they are poor mutagens, suggesting an indirect mechanism of action [23, 24]. The precise mechanism in which these heavy metals induce carcinogenesis is undefined. In this manuscript, we report:Both stable and transient transfection assays are successful in measuring L1 retrotransposition in a reproducible and consistent manner.Mercury, nickel and cadmium significantly increase L1 retrotransposition in a dose-dependent manner.These stimulatory effects are not universal for all metals, as cobalt, zinc and magnesium had no stimulatory effect on the L1 retrotransposition.This is the first report demonstrating the stimulatory effects of carcinogenic environmental pollutants on L1 retrotransposition. Our results also demonstrate the need to take into account DNA damage through mobile element activation as one of the mechanisms of action of genotoxic agents.The following compounds were purchased from SIGMA-Aldrich: cadmium sulfide, CdS 99.999%, mercury (II) sulfide, HgS, 99.5+% HgS and nickel oxide (NiO), 99.9%. The filtered stock solutions of HgS and CdS were quantified professionally by AccuLab Inc. of Louisiana using method numbers 213.1 for CdS and 245.1 for HgS as described in USEPA Methods for Chemical Analysis of Water and Wastewater. The concentrations of the stock suspensions were determined to be 1.37 mg/L (ppm, parts per million) or 1370 ppb (parts per billion) for HgS and 1.15 ppm or 1150 ppb for CdS. The doses tested in the assays were various dilutions (example from 50 fold to 3000 fold) of the stock solutions using regular DMEM medium lacking any antibiotics as a diluent. During any procedure (preparation or experimental), the suspensions were continuously shaken in the hood at low speeds to ensure the uniform suspension of the particles of CdS and HgS.JM101/L1.3ΔCMV (a kind gift of Dr. John Moran) was used in the previously described retrotransposition assay [3]. The plasmid contains a full-length functional L1 element (schematic shown in Figure 1A). PIRES2-EGFP (Clontech) contains a neomycin resistance expression cassette and was used in parallel in the retrotransposition assays as a combined control for transfection and cytotoxicity. All plasmid DNA was purified by alkaline lysis and twice purified by cesium chloride buoyant density centrifugation. Final evaluation of the DNA quality was performed from the visual assessment of ethidium bromide stained agarose gel electrophoresed aliquots.HeLa cells (ATCC CCL2) were grown in a humidified, 5% CO2 incubator at 37°C in Earl’s minimal essential medium (EMEM). EMEM was supplemented with 10% fetal bovine serum. HeLa cells were seeded in T-75 flasks at a density of 1.5 × 105 cells/flask and grown for 20 hours prior to transfection. Cells were transfected with the Lipofectamine Plus (InVitrogen) for three hours using 1μg of either the L1 or 0.3μg the neomycin control plasmid with 18μl of the plus reagent and 12μl of Lipofectamine, following the manufacturer’s protocol. The transfection mix was aspirated and replaced with the complete media supplemented with the appropriate dose of the compound evaluated. After a 48 hour period, the treatment was removed and the cells were grown for two weeks using selection media (containing 400μg/ml G418) to obtain the neomycin resistant (neoR) colonies. Cell colonies were fixed and stained for 30 minutes with crystal violet (0.2% crystal violet in 5% acetic acid and 2.5% isopropanol). A schematic of the time line is shown on Figure 3B.L1-stable cell lines of NIH3T3 (ATCC CRL1658) with the integrated human L1 retrotransposition cassette were generated. To integrate the plasmid the method of electroporation was selected as it usually generates clones with the integration of only one copy of the plasmid. 106 cells were electroporated with 6μg of JM101/L1.3ΔCMV using 300V and 800μF. The electroporated cells were grown under hygromycin selection to obtain colonies containing the integrated plasmid. A mouse cell line (NIH3T3) was utilized to facilitate the evaluation of the hygromycin resistant clones to detect the integrated plasmid that contains a human L1. Once the individual colonies were grown, they were subsequently evaluated for their L1 retrotransposition capability as determined by the generation of neomycin resistant (neoR) colonies. Two clones (3E and 28E) contained an integrated copy of the L1 vector and were also L1 retrotranspositionally competent. Clone 28E was utilized for most of the experimental evaluations, as it generated a low inherent background.Stable L1-vector clonal cell lines (described above) were used to evaluate L1 retrotransposition. Each assay lasted for 10 weeks and consisted of two parts a) retrotransposition assay and b) parallel toxicity control:Testing the chemical for effects on retrotransposition: To keep the background retrotransposition rate to a minimum, a fluctuation analysis was carried out by first seeding 1–2 cells of clone 28E into 12-well plates. The cells were fed with DMEM medium containing 5% FBS (Invitrogen) and 0.01% non-essential amino acids (Invitrogen), but lacking any antibiotics or antifungal agents. When the cells reached the 100-cell stage, they were treated as described in the results with the heavy metal for two nonconsecutive 24 hour exposures (Figure 2C). Upon allowing the cells to recover for 24 hours, 250,000 cells were seeded per T75 flask and subjected to selection (G418 400μg/ml) for 14 days. A schematic of the assay time line is shown on Figure 2C. As per the design of the L1 vector, only retrotransposed cells would be resistant to G418. The retrotransposed clones were stained with crystal violet for 3 hours, washed, allowed to dry and counted as raw data. For the fluctuation analysis, a minimum of twelve T75 flasks (each originating from a separate 1–2 cell seeding) were used for the determination of L1 retrotranspositional rate.Toxicity Control: To set up the toxicity control, two populations of NIH3T3 cells were utilized:Native NIH3T3 cells lacking the L1 construct and sensitive to the antibiotic G418;The neoR control cells, carrying an L1 insert and resistant to the antibiotic G418.Both cell types were propagated from a 1–2 cell stage, and dosed at the 100-cell stage as in the experimental treatment. Then the cells were allowed to recover and divide until their numbers reached about one million (106). However, at the stage of G418 selection, this protocol differed from the basic assay described above. Here, a mixture of 100 of the neoR cells plus 250,000 native cells were seeded into a T75 flask and grown under G418 selection for 14 days. The resultant clones were stained with crystal violet and counted and used to calculate the survival curve. The untreated group should generate an average of 100 colonies of neoR cells. If the chemical has a toxic effect on the cell viability, treated samples would present a lower number of retrotransposed clones than the expected 100. Finally, if the chemical has any effect on cell proliferation, the number of retrotransposed clones should be higher than the expected 100 in the dosed flasks.Analysis for the experiments using stable transfectants was performed using fluctuation analysis for mutations following the method originally described by Luria-Delbrück [25] with the appropriate modifications for the application to tissue culture assays [26]. Results obtained from transient transfection assays were evaluated using paired t-test.About half a million L1 elements are found dispersed throughout the genome. However, most of them are 5′ truncated [27] and estimates suggest that about 1000–3000 of them are full-length copies [28, 29] and only approximately 10% of those are likely to be active [30]. For our evaluation of the effects of L1 retrotranspositional activity, we designed an approach to parallel as closely as possible the conditions of the active L1 elements within the genome. The previously described L1 retrotransposition reporter vector, JM101/L1.3ΔCMV plasmid [19] was stably integrated into the genome. Individual clones were selected and amplified to homogeneity. The clones were evaluated for integration of the plasmid and for L1 retrotranspositional capability. The L1 vector used contains the sequence of an active L1 element driven by its endogenous promoter, and designed to allow expression of the neomycin resistance (neoR) gene only when the expressed L1 element goes through retrotransposition (schematic in Figure 1A).The use of a L1-stable cell line allows for the evaluation of potential epigenetic factors, such as methylation, that can affect the activity of a genomic L1 element [31–33]. However, the integrated L1 vector is continually active, generating a low level of integrated copies that contain a functional neomycin resistance gene. Thus, the simple task of growing the cell line will generate a mixed population of neomycin sensitive (no new L1s) and neomycin resistant cells (Figure 1B). Not only is the population mixed, but the proportions change relative to when in the cell expansion period L1 integration occurs. For example, if we start with 10 cells and at this stage 1 insert occurs, this will translate in a 10% neoR background. However, if the cells had divided and the L1 insert occurs at the 100-cell stage a 1% neoR background is observed. In contrast, the transient transfection assay introduces the plasmid into the cells at the same time as the treatment, eliminating the background levels (Figure 1C). Thus, to properly evaluate effects of the treatment on the stable L1-cells, we performed the assay in a way to be able to apply fluctuation analysis (details in materials and methods). Basically, the assay is started with 1–2 cells and the cells are treated at the 100-cell stage, most of the initial pool of the 28-E cells will have a zero background for retrotransposition, although a few may have 1% or more background (one or more cells out of the 100).At the end of the experiment it is very obvious which set of 100 had an initial background, as the number of clones will be at least an order of magnitude greater than expected. These points can be eliminated, leaving only the experiments that initiated with zero background. This approach is the basis of the ‘fluctuation test’ and its extended studies [25, 26, 34], originally designed to determine spontaneous mutation rates in cultured bacterial and mammalian cells.Heavy metals are cytotoxic to cultured cells, thus a toxicity control was included. A set of neoR cells and a set of native (non-L1 transfected) cells were treated in parallel to evaluate the influence of toxicity on cell survival and colony formation. The average number of colonies with no treatment was used as the 100% value for both the toxicity and L1 activity (Figure 2A). A correction factor was utilized to compensate for the toxic effects of the heavy metals when comparing the L1 retrotransposition to the control at different doses. For example, if dose X of HgS reduced the number neoR colonies of the control plasmid by 20%, the number of colonies from the L1 with the same treatment was corrected to compensate for that 20% reduction (Figure 2A, gray columns). Using this approach, the plating control numbers were used to adjust the data obtained from the first part of the assay to determine the adjusted L1 retrotransposition rate and to obtain normalized results for plotting (Figure 2B).We evaluated the water-insoluble carcinogenic forms (particulate) of the cadmium and mercury (CdS and HgS). Reports in the literature demonstrate that the phagocytosis of the particles allows for a more efficient internalization of high levels of these compounds into cells leading to a higher concentration of the compound within the cell [35–37]. Evaluation of mercury and cadmium demonstrated that both metals significantly increase L1 retrotransposition in a dose-dependent manner (Figure 2). About a three-fold peak stimulus is observed on the number of neoR colonies recovered (Fig 2A) also reflected when L1 retrotransposition rate is calculated using fluctuation analysis (Fig 2B). As expected, at higher doses, the overall retrotransposition rate drops, probably due to toxicity leading to cell death. The toxicity control data were utilized to correct for toxicity (gray bar) as described above. The stimulation of retrotransposition by all the doses (0.5, 1.4, 4.6 and 13.7 ppb) of HgS and 2.9, 5.8, 11.5 ppb of CdS were significantly different relative to the no dose control, p < 0.01 (paired t-test).In this assay, HeLa cells were transiently transfected with the JM101/L1.3ΔCMV plasmid. The cells are then exposed to different doses of the different metals for 48 hours prior to the two weeks of G418 selection to obtain neoR colonies (Figure 3B). The no treatment (0 dose) was defined as 100% and the rest of the data were plotted relative to it. As toxicity control, an unrelated plasmid, containing a functional neoR gene (pIRES2-EGFP) was transfected in parallel allowing for the evaluation of the influence of toxicity on both transfection efficiency and colony formation. Adjustment for toxicity was performed in the same manner described above. Our results demonstrate that both HgS and CdS generated a 3-fold peak stimulation of L1 activity equivalent to the stimulation observed in the stable L1 assay (Figure 3A). However, the dose range where the stimulation is observed varies between both assays. The L1-stable assay presents an increased sensitivity, seen as the detection of an effect with lower doses.In addition, we evaluated other metals, and NiO also increased L1 activity (Figure 3), however soluble cobalt, magnesium and zinc chloride had no effect on L1 activity (data not shown).The mechanism by which some heavy metals induce cancer is unclear [22, 38–41]. The reactions of cells to environmental exposure are very complex, involving multiple pathways and cellular components. In this manuscript we present data supporting yet another mechanism by which heavy metals may cause disease.Most of the steps involved in the L1 retrotransposition are currently poorly understood, making the determination of the mechanism by which these compounds induce L1 activity difficult. However, not all heavy metals evaluated have such stimulatory effects, so our results suggest that nickel, cadmium and mercury likely have a specific mechanism that result in these effects. The influence of these metals on the rate of L1 retrotransposition may occur at any of the retrotransposition steps, such as transcription or insertion into the genomic sites. We are currently evaluating the potential mechanism by which these metals may be affecting L1 retrotransposition (El-Sawy, et al. unpublished).Until recently, it was impossible to measure the changes in the frequency of L1 jumping. Under the optimized conditions, both assays show reproducible, overlapping patterns of stimulatory effects of HgS and CdS (the cancerous forms) on L1 retrotransposition. The assay using a stably-integrated L1 vector better reflects the L1 elements, which are naturally present throughout the genome. There are extensive data suggesting that chromatin-level regulatory events may be important with L1 elements [31–33], which the L1-stable assay would more appropriately evaluate. However, the assay is highly complex, and demands a long time for completion (~10 weeks). As an alternative, we implemented the use of transiently transfected cells for the evaluation of different compounds on L1 activity. We realize that our transient assay does not fully mimic the natural L1 amplification. Because the L1 reporter system involves expression from a transiently transfected plasmid, the assay should not be influenced by factors such as methylation or chromatin structure. However, this alternative assay proved to be simpler, shorter in duration, highly reproducible and yielded comparable results to the stable-L1 vector assay. Although more evaluation is required, these assays may prove highly valuable in evaluating environmental compounds and their potential impact on genetic instability. Also, the assays used in this study are currently the only available that can assess the influence of environmental exposure on retrotransposon activity in humans.Our study is the first of its kind to report stimulatory actions of certain heavy metals on human retrotransposition. Our findings that different compounds can stimulate retrotransposition are particularly relevant to the possible role of chronic exposures in both germline and somatic disease. Chronic exposure to heavy metals, and other toxicants, through workplace and environmental exposures may specifically increase the damage caused by retrotransposition over a long period of time. Previous data demonstrate that many heavy metals are not mutagenic by the standard bacterial assay systems; however, they are carcinogenic when tested in animal models [23, 24]. Further, our results demonstrate that even small amounts of these metals (in the range of ppb) can cause on average about 3-fold increase in the rates of L1 retrotransposition. This is of high relevance when considering that the ‘allowable’ sewer-discharge limit for cadmium according to the official guidelines by the NEA (National Environmental Agency) is 1 ppm (http://app.nea.gov.sg/cms/htdocs/article.asp?pid=1644), and that most humans do accumulate mercury through seafood ingestion and other sources in their lifetimes. Such cells, that have increased L1 retrotransposition due to exposure of CdS and HgS, very likely become more inclined to accumulate potentially deleterious mutations that may lead to human diseases like cancer. Accumulation of this damage may contribute to initiation and progression of cancer, as well as other diseases of chronic exposure. We present data to support a novel mechanism by which metals and other compounds can cause genomic damage through the modulation of mobile element activity. Our data suggest that the genetic damage caused by the stimulation of mobile elements present in the genome needs to be seriously considered as an alternate mechanism by which environmental/external compounds generate genetic instability and disease. Thus, we propose the following model of how the heavy metals cause genetic instability through the stimulation of mobile element activity.We present a model of the role the environment and mobile elements have on the generation of diseases and cancer (Figure 4). The integrity of the cellular homeostasis is constantly being assaulted by external components including environmental agents, such as oxidants, heavy metals, UV radiation, etc. The cell presents a complex system of regulatory and protective pathways to either correct (ie. DNA repair pathways) or prevent the continuation of an unrepaired event (cell cycle arrest and apoptosis). A “healthy” or disease free state is maintained as long as there is a balance of these external effects and the cellular controls. Mobile element activity is among the factors that influence this balance. For example, reports demonstrate that methylation of mobile elements maintains their expression under control [31, 42–43]. If the regulation by methylation is eliminated, the increase in expression of repeats leads cells to have an increased genetic instability with disastrous consequences [44]. Thus, external factors, such as heavy metals, affect the regulation of mobile element activity, allowing for an increase in altered cells. Mobile element activity may continue to be high in these cells, accumulating even more defects that in conjunction with inactivation of tumour suppressors, etc. may lead to a cancerous state.The ability to directly or indirectly increase mobile element activity may represent a major contributory factor to genetic instability in somatic cells, leading to cancer initiation or progression, aging, or other diseases associated with chronic exposure to carcinogenic/toxic compounds. These data suggest that damage caused by retrotransposition may need to be considered when developing mechanistic models for genetic damage associated with environmental exposures.The L1 assay systems.A. Schematic of the L1 vector. (1) The vector contains a full length L1 element with two open reading frames (ORF1 and ORF2). The construct contains the SV40 promoter (Sv40p) in the 3′UTR in the “reverse” direction that will transcribe a neo gene containing a “forward” intron that affects proper expression of the neomycin resistance. (2) RNA transcription is driven by the internal L1 promoter (L1p) located in the 5′untranslated region (UTR). The intron interrupting the neomycin resistance gene will be removed by splicing (SD: splice donor, SA: splice acceptor) only from RNA generated from the L1 promoter. (3) In the L1 retrotransposition process the RNA is reverse transcribed, followed by integration of the DNA into the genome. (4) The new L1 copy contains a functional neo gene. (5) Only newly integrated copies that retrotransposed from the spliced L1 RNA will generate neomycin resistance. The neo gene in opposite orientation relative to the L1 gene is shown as a gray box. RNA is represented by wavy lines with arrows to show direction of transcription. Note that the scale of the figure is not proportionally accurate.B. L1 retrotransposition - L1-stable cell line assay. This assay is based on the use of a clonal cell line containing the L1 reporter vector integrated into the genome (gray cells). In these cells, the L1 cassette will have a baseline expression, which consistently generates new L1 inserts (gray with black nucleus). In other words, during the normal passage and seeding of the cells a steady background of neomycin resistant cells is being generated. Also, depending on when the event occurs (during the first division or later passage), the background level could represent from 1% up to 50% of the cells, even before the exposure to an agent. Thus, a fluctuation analysis is used to appropriately evaluate the effect of compounds in this assay.C. L1 retrotransposition-transient assay. Cells are transiently transfected with the L1 reporter plasmid. Only a portion of the cells will uptake the plasmid and express the L1 RNA (gray cells). The cells are grown in the presence or absence of the compound tested (treatment). During this period a small fraction of the cells expressing L1, will generate a retrotransposition event as observed by the acquisition of neomycin resistance (gray cell with black nucleus). Growth under selection media will generate neomycin resistant (neoR) colonies each representing at least one retrotransposition event.Effect of cadmium and mercury on L1 retrotransposition activity (L1-stable cell line assay).A. NeoR colonies from separate L1 transfections (black bar) treated with different doses of HgS are shown. The no treatment (0 dose) for each experiment was used as the 100%. The average number of colonies from the toxicity control (white bar) was used to adjust the observed numbers (gray bar) as described in the text. Three independent assays in triplicate (n=9) were performed using clone 28E and error bars indicate one standard deviation. Statistically significant differences relative to the no treatment are indicated by asterisks [t-test p<0.01(*), p<0.001(**)].B. L1 insertion rate: Fluctuation analysis was performed on the basic data (example for HgS shown in A) to obtain the adjusted L1 retrotransposition rate. Both CdS and HgS stimulate L1 retrotransposition from a genomically inserted L1-vector in a dose dependent manner with the maximum stimulus being around 3 fold. Average insertion rate is shown for HgS is 10−5 and CdS in 10−6 retrotransposition rate/cell division, with error bars representing one standard deviation. Statistically significant differences relative to the no treatment [t-test p<0.01(*), p<0.001(**)] are shown.C. Schematic of time line of the L1-stable cell line assay: One or 2 cells of the L1 stable cell line were initially seeded and grown to 100 cell stage where they were exposed to the treatment with the selected agent. After treatment and recovery cells were reseeded (250,000 in a T75) and grown under selection for 2 weeks. NeoR colonies were counted after staining. The total duration of the assay is on average 10 weeks.Effect of different compounds on L1 retrotransposition activity (transient transfection assay).A. Metals stimulate L1 retrotransposition in a transient transfection assay: NeoR colonies from separate L1 transfections (black bar) treated with different doses of HgS, CdS or NiO (X axis) are shown. An unrelated plasmid with neomycin resistance was used as a transfection and toxicity control (white bar). The no treatment (0 dose) for each experiment was used as the 100%. In a similar manner as used for the L1-stable assay described in the text, the data were adjusted for toxicity (gray bar). Three independent assays in triplicate (n=9) were performed in HeLa cells and error bars indicate standard deviations. Statistically significant differences are indicated relative to the no treatment [t-test p<0.01(*), p<0.001(**)]. All the metals tested with this assay show a stimulation of L1 retrotransposition comparable to that observed in the L1-stable assay.B. Schematic of time line of the L1 transient assay: The L1-vector is transfected into cells that were seeded the previous day (150,000 in T75). Immediately after transfection (3 hour) cells are treated with the metal for 48 hours. Treatment is removed and cells are grown under selection for 2 weeks before staining. Total duration of the assay is 2.5 weeks.Proposed model for how metal stimulation of mobile elements may impact genetic stability.The balance between normal cells and altered or mutated cells is depicted as an equilibrium. External and internal components can affect the outcome of this equilibrium. Normal cells are consistently being mutated or altered by external factors, such as UV light, etc. The damage generated can be either repaired or progress to two outcomes: (1) Apoptosis to eliminate the damaged cell or (2) Disease/cancerous cell. We propose that heavy metals shift the equilibrium through the increase of retroelement activity potentially leading to the accumulation of more mutated cells.This research was supported by National Institutes of Environmental and Health Sciences ARCH grant, 1S11ES09996 (SPK and PLD) NIH RO1 GM45668 (PLD), NSF EPS-034611 (PLD), the State of Louisiana Board of Reagents Support Fund (PLD) and NIH P20 RR020152 (AMR-E and PLD).Unpublished: El-Sawy, M.; Kale, S. P.; Dugan, C; Nguyen, T. Q.; Perepelitsa-Belancio, V.; Bruch, H.; Roy-Engel, A. M.; Deininger, P. L: Nickel stimulates genetic instability through LI retrotransposition.
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+ Ultra–wideband (UWB) technology has increased with the use of various civilian and military applications. In the present study, we hypothesized that low-dose UWB electromagnetic radiation (UWBR) could elicit a mitogenic effect in AML-12 mouse hepatocytes, in vitro. To test this hypothesis, we exposed AML-12 mouse hepatocytes, to UWBR in a specially constructed gigahertz transverse electromagnetic mode (GTEM) cell. Cells were exposed to UWBR for 2 h at a temperature of 23°C, a pulse width of 10 ns, a repetition rate of 1 kHz, and field strength of 5–20 kV/m. UWB pulses were triggered by an external pulse generator for UWBR exposure but were not triggered for the sham exposure. We performed an MTT Assay to assess cell viability for UWBR-treated and sham-exposed hepatocytes. Data from viability studies indicated a time-related increase in hepatocytes at time intervals from 8–24 h post exposure. UWBR exerted a statistically significant (p < 0.05) dose-dependent response in cell viability in both serum-treated and serum free medium (SFM) -treated hepatocytes. Western blot analysis of hepatocyte lysates demonstrated that cyclin A protein was induced in hepatocytes, suggesting that increased MTT activity after UWBR exposure was due to cell proliferation. This study indicates that UWBR has a mitogenic effect on AML-12 mouse hepatocytes and implicates a possible role for UWBR in hepatocarcinoma.UWB technology involves the radiation, reception, and processing of wide bandwidth radio frequency emissions [1]. The global use of UWB technology has drastically increased in recent years. Exposure to UWB devices has incited public concern over potential health risks. UWBR (also known as “nanopulses”) is used in radar technology and electronic warfare to illuminate objects for wideband spectral analysis [2]. UWB devices used by civilian and military organizations include: tactical handheld and network low-probability interference/detection (LPI/D) radios; LPI/D wireless intercom systems; intelligent transportation systems, electronic tags, and smart appliances; intrusion/detection radar; precision geo-location systems; and high speed data links [3]. UWB devices can transmit data over a wide spectrum of frequency bands with very low power. Moreover, low power short range microwave devices operating on various frequencies have been used in ground-penetrating radar for obtaining the images of objects buried under ground or behind surfaces; wireless communications, particularly for short-range high-speed data transmissions; home safety intrusion detection systems; and medical devices for monitoring biological systems [3].UWBR is a form of non-thermal electromagnetic fields (EMFs), and has photon energy less than 10eV; a level not sufficient to produce ions by ejection of orbital electrons from atoms [4]. The energy of non-thermal EMFs does not break chemical bonds directly; the effects are caused by indirect mechanisms [4]. Genetic effects of extremely low UWBR are well documented [5–7]. Exposure to extremely low UWBR has been closely linked to an increased risk of cancer and frequency of micronuclei, DNA-damage, mitogenic activity, up-regulation of heat shock proteins (hsp), and alterations in sleep patterns [8–12].The potential risk of cancer remains a primary concern of extremely low UWBR exposure. The non-thermal activity of UWBR has been considered too weak to damage or mutate DNA [13, 14], however some studies indicate mutagenic or comutagenic activity [15, 16]. In addition, UWBR has the ability to increase proliferation in various cell types [17, 18]. It is has been suggested that extremely low UWBR contributes to the promotional phase of cancer [18]. Extremely low UWBR has been shown to mimic the tumor-promoter agent, 12-O-tetradecanoylphorbol-13-acetate, by inhibiting Friend erythroleukemia cell differentiation, and thereby stimulating cell proliferation [18]. In the present study, we hypothesized that low-dose UWBR could elicit a mitogenic effect in AML-12 mouse hepatocytes, in vitro.Dulbecco’s Modified Eagle’s Medium (DMEM) was purchased from Hyclone (Lot No. ANK19799; Logan, Utah) and tissue culture supplements were purchased from American Type Culture Collection (ATCC) (Manassas, VA). Dulbecco’s phosphate buffered saline (Lot No. 1163547) was obtained from Invitrogen Corporation (Grand Island, New York). Thiazolyl blue trazolium bromide (CAS 298-93-1; purity 97.5%), β-mercaptoethanol, and dimethyl sulfoxide were purchased from Sigma-Aldrich (St. Louis, Missouri). ITS (insulin-transferrin-selenium) was obtained from Cambrex Bio- Science Baltimore, Inc. (Baltimore, MD). Twelve percent SDS-PAGE gels were obtained from ISC BioExpress (Kaysville, UT). Cyclin A primary antibody was purchased from Calbiochem (La Jolla, CA). Alkaline phosphatase conjugated goat-anti-mouse IgG secondary antibody, and BCIP/NBT color development substrate were purchased from Promega (Madison, WI). Reagents for protein determination, gel electrophoresis, and Western analysis were obtained from Bio-Rad (Hercules, CA).The UWBR exposure facility at Louisiana Tech University, Ruston, LA, is illustrated in Figure 1. The large room was shielded with copper and steel plating and contained a GTEM cell and pulser. A D-dot probe was used to determine pulse characteristics in the GTEM cell. The output of the probe was proportional to the time derivative of the electric field. The pulser was operated from a second copper mesh, RF-shielded enclosure. A 6 GHz bandwidth digital storage oscilloscope (Tektronix Model TDS 6400) was used to monitor pulse experiments. Cable termination panels on both rooms allowed room-to-room electrical connections.Alpha mouse liver 12 (AML 12) hepatocyte cultures were established from a mouse transgenic for human transforming growth factor α (ATCC CRL-2254, Manassas, VA). The cells were stored in liquid nitrogen in the laboratory until use. The contents of each vial were transferred to a 75 cm2 tissue culture flask diluted with DMEM, supplemented with 10% fetal bovine serum (FBS) and 1% streptomycin and penicillin (hepatocyte growth medium; HGM), and incubated at 37°C under an atmosphere of 5% CO2 in an incubator with humidified air to allow the cells to grow and form a monolayer in the flask. Subsequently, cells grown to 80–95% confluence were washed with phosphate buffer saline (PBS), trypsinized with 5 mL of 0.25% (w/v) EDTA, diluted, counted, and seeded in two 96-well microtiter tissue culture plates (5 × 105 cells/well).In all experiments, cells were grown in HGM for 24 h prior to UWBR treatment. On the day of the experiment, medium was replaced with fresh HGM or serum-free growth medium (SFM). In some experiments, medium was supplemented with ITS at the following concentrations: .625, 1.25, 2.5, μg ITS/mL. For UWBR exposure, microtiter plates were placed in a horizontal position inside the GTEM cell. Samples were exposed to UWBR for 2 h at a temperature of 23°C. The pulse width was 10 ns, the repetition rate 1 kHz, and the applied field strength was in the range, 5–20 kV/m. Pulses were triggered by an external pulse generator for exposure or not triggered for sham exposure.Following a post-exposure period of 8- to 24 h, cell viability was evaluated using a colorimetric assay in which the reduction of a tetrazolium salt [3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide] (MTT) by mitochondrial dehydrogenases of living cells was detected. In this assay, metabolically active cells were able to convert MTT to water-insoluble dark-blue formazan crystals. Viable cells were quantified by dissolution in 100% dimethyl sulfoxide and measured by absorbance with the wavelength set at 540 nm, using an EL 800 Model ELISA plate reader (Bio-Tek Instruments Inc., Winooski, Vermont) [19].Cells grown to 80–95% confluence were washed with PBS, trypsinized with 5 mL of 0.25% (w/v) EDTA, diluted, counted, and seeded in two 96-well microtiter tissue culture plates (5 × 105 cells/well). Cells were exposed to UWBR as described above. Following a post-exposure period of 24 h, an equal volume of sample buffer (0.2 mol/L Tris, pH 6.8, 1% SDS, 30% glycerol, 7.5% β-mercaptoethanol, 0.1% bromophenol blue) was added to each well. Cells were mechanically dislodged, transferred to microcentrifuge tubes, and heated at 95°C for 10 min. Samples were then frozen until future use. The Bradford protein assay in a microtiter plate format was used for the determination of protein concentrations in samples. The total protein concentrations for cell lysates were quantitatively measured at 540 nm absorbance; using the Multiskan Ascent microplate reader (Labsystems, Beverly, MA).Whole cell extracts from AML-12 mouse hepatocytes were heated at 100°C for 10 min and electrophoresed on a 12% SDS-polyacramide gel. Separated proteins were transferred onto a nitrocellulose membrane in 20 mM Tris base, 150 mM glycine, 20% methanol (pH 8.0). Subsequently, the nitrocellulose membrane was blocked (10 mL of Tris-buffered saline 0.1 Tween-20 [TBST] with 5% nonfat dry milk) for 1 h at room temperature. Cyclin A protein was detected using cyclin A (1:200) bovine primary monoclonal antibody that was then detected with a 1:750 dilution of alkaline-conjugated goat anti-mouse IgG, secondary antibody. BCIP/NBT color substrate was incorporated to develop protein bands. Immunoblot protein bands were assessed for abundance by TotalLab computer software (Nonlinear USA Inc. Durham, NC)Results are given as the mean ± SDs for the indicated number of independently performed experiments. Differences between the mean values were evaluated by Student’s t test.The effects of UWBR on the viability of AML-12 mouse hepatocytes are shown in Figure 2. In this experiment, we exposed hepatocytes to low-dose UWBR for 2 h, as indicated in the methodology section. Results from this experiment demonstrated a statistically significant (p<0.05) time-related increase in cell viability within the post-exposure periods of 8–24 h. The maximum increase in cell viability was 38% at the 24 h post-exposure period.We also investigated whether UWBR could cause a similar response under the same conditions for SFM-treated hepatocytes for a post-exposure period of 24 h. The comparison of SFM- and HGM-treated hepatocytes exposed to low-dose UWBR is shown in Figure 3. Upon the post-exposure period of 24 h, UWBR had a smaller but significant (p<0.05) stimulatory effect on cell viability in SFM-treated cells. These results suggest that the low field strength (5–20 kV/m) of UWBR elicits a greater proliferative effect in HGM-treated AML-12 mouse hepatocytes.Next, we examined the effect of UWBR on AML-12 mouse hepatocytes maintained in HGM- and SFM containing ITS, Figure 4. A dose-dependent response was demonstrated in HGM-treated hepatocytes with ITS; showing a statistically significant (p<0.05) increase in cell viability within the dose range of .625–2.5 μg ITS/mL. For example, cell viability percentages of 119 ± .16%, 127 ± .18%, and 147 ± .22%, were recorded for .625, 1.25, and 2.5, μg ITS/mL, respectively. The greatest observed effect was a 1.5-fold increase in proliferation at 2.5 μg ITS/mL. With regard to SFM-treated hepatocytes containing ITS, a dose-dependent effect was also demonstrated within the dose range of .625–2.5 μg ITS/mL. Statistically significant (p<0.05) cell viability percentages were recorded as 108 ± .23%, 103 ± .17%, and 105 ± .19%. Under SFM conditions, the highest increase in cell viability was observed at .625 μg ITS/mL; representing 1-fold increase.To determine whether the increase in MTT activity resulted from an increase in proliferation, we tested the ability of UWBR to stimulate the induction of the cyclin A protein in AML-12 mouse hepatocytes. A qualitative identification of cyclin A protein was made by Western blot analysis and quantitative assessment by densitometric analysis. Densitometric analysis of cyclin A expression in control and UWBR-treated mouse hepatocytes is shown in Figure 5. The mitogenic activity of UWBR was evidenced by the induction of the 60-kDa cyclin A protein in UWBR-treated mouse hepatocytes. The overexpression of the cyclin A protein in hepatocytes constitutes an indication of the mitogenic activity of UWBR.Recent investigations from our laboratory have demonstrated that PCP, a tumor-promoter agent, has the ability to induce cell proliferation in AML-12 mouse hepatocytes [20]. We chose the AML-12 mouse hepatocyte cell line to study the mitogenic effect of UWBR, in vitro, because hepatocarcinoma is the primary form of liver cancer that arises from hepatocytes. Data presented in this study show that UWBR induced an increase in cell viability of AML-12 mouse hepatocytes, measured by MTT incorporation (Figure 2). It has been suggested that extremely low UWBR plays a role in the carcinogenic process by eliciting tumor-promoting effects, and therefore, is likely to be involved in the promotional phase of cancer [18]. Moreover, it has been shown that the tumor-promoting effects of extremely low EMFs inhibited differentiation and stimulated proliferation of Friend erythroleukemia cells [18]. Our results are similar to those of Chen et al. [18], who demonstrated that extremely low EMFs stimulated cell proliferation 50% above the sham-treated Friend erythroleukemia cells.In the present study, we used the GTEM cell for exposing our cells to UWBR. The uniformity of GTEM field exposure is critical to quantifying the biological response versus the electromagnetic dose [21]. Moreover, conducting medium, such as culture medium in vitro, in combination with GTEM field exposure are determinant factors for dose-dependent effects of EMF on biological cells [21]. Various studies have shown that EMFs have a proliferative effect on cells in culture medium containing growth additives [22–24]. In our experiments, we used HGM-and SFM conditions to compare the effects of UWBR on AML-12 mouse hepatocytes. We observed a higher level of MTT activity in HGM-grown cells compared to those grown in SFM, but both showed increased MTT activity compared to control (Figure 3). Typically, serum-free conditions cause cells to withdraw from the cell cycle and/or differentiate. A previous investigation exposed human chrondocytes to low-frequency pulsed EMFs in medium supplemented with 10% or 0.5% fetal calf serum and in serum-free medium [24]. As a result, 3H-thymidine incorporation showed an increase of cell proliferation in cultures exposed to low-frequency pulsed EMFs when serum was present in the culture medium, whereas no effect was observed in serum-free conditions [24]. The same study concluded that the proliferative response of chondrocytes to the low-frequency pulsed EMF exposure was dependent on the amount of serum in the culture medium [24]. It was further suggested that increased DNA synthesis of chondrocytes was dependent on the amount of fetal calf serum which is similar to the basic mechanism of low-frequency pulsed EMF stimulation of fracture healing [24]. Data obtained from our experiments demonstrated a similar pattern in SFM-treated hepatocytes exposed to UWBR; indicating that UWBR has very little effect on hepatocytes under serum-free conditions. Our results show that the greatest mitogenic response of mouse hepatocytes to UWBR exposure is dependent on the presence of FBS in the culture medium.We have shown that the addition of ITS (insulin, transferrin, and selenium), a mitogen, causes a dose-dependent response in HGM- and SFM-treated cells. Cytokinetic analysis using fluorescence flow cytometry and microscopic techniques indicated that selenium treatment increased the duration of G1, S, and G2 phases of the cell cycle [25]. It has been suggested that the presence of growth factors and mitogens in culture medium is necessary to obtain an induction of cell proliferation after low-frequency pulsed EMF exposure; indicating that the interaction between cell membrane receptors and mitogens is one of the molecular events affected by extremely low EMFs [24, 26]. Cossarizza and colleagues [26] demonstrated that low-frequency pulsed EMF exposure induced an increase of cell proliferation and IL-2 receptor expression only in phytohemagglutinin (PHA) -stimulated lymphocytes, whereas no effect was observed when PHA was not added to the culture medium.Cell proliferation and progression through the cell cycle are regulated by the sequential events of various cyclin-dependent kinases (cdks) [27, 28]. Cyclin A is required for cell cycle S phase entry, and its overexpression has been strongly linked to tumorigenesis [29]. Moreover, cyclin A is considered to be a rate-limiting component required for both the initiation of DNA synthesis and entry into mitosis [30]. In the present study, we have established that UWBR causes the upregulation of cyclin A (Figure 5). It has been suggested that cyclin A is a direct transcriptional target of JunB, thus, driving cell proliferation [30].Biological membranes have an appreciable electrical capacity to store and separate charges. Exposure to EMFs induces currents into tissues, and therefore, bioeffects that are dependent on frequency, wave shape, and intensity. Cellular stress caused by exposure to EMFs may trigger a change in metabolic activities, cell cycle reaction rates, cell morphology, electrical charges across the plasma membrane, and gene expression in the nucleus [31–34]. In addition, EMFs can have a regulatory affect on free radical- and enzymatic reactions in biological systems [35, 36]. One study found that treatment of rats before exposure to EMFs with free radical scavengers, melatonin and N-tert-butyl-α-phenylnitrone, blocked the effects of EMF on DNA single- and double-stranded breaks in their brain cells [37]; thus suggesting that EMF enhances free radical activity in cells, which in turn lead to DNA-damage. Another investigation exposed a mast cell line, RBL-2H3, to an EMF of 835 MHz and showed an increased rate of DNA synthesis and cell replication [34]. The same study associated alterations in actin distribution, cell morphology, and calcium ionophore to EMF exposure [34]. A recent investigation demonstrated that in vitro exposure of human peripheral blood lymphocytes to continuous 830 MHz EMFs caused an increase in chromosomes (aneuploidy), a phenomenon known to increase the risk for cancer [38]. Results from the same study suggest that the genotoxic effect of the EMF is elicited via a non-thermal pathway [38]. EMF ability to stimulate cell proliferation has been demonstrated in human astrocytoma cells exposed to EMFs of 60-Hz [17]. It has been suggested that cells with high rates of iron intake, especially proliferating cells, and cells with high metabolic rates would be more sensitive to the effects of EMFs [4]. Well-documented studies show that extremely low EMFs, such as UWBR, can alter the transcription and translation of various genes [8, 39].Various studies suggest that the interaction site for extremely low UWBR is the plasma membrane, since exposure can alter Ca2+ influx [40, 41]. Increases in intracellular (I) Ca2+ can be mediated by Ca2+ influx across the plasma membrane and/or by Ca2+ release from internal compartments [41]. ICa2+ plays an intimate role in regulating cell migration, differentiation, proliferation, and signal transduction pathways. Hepatocytes are target cells for biotransformation; however, they lack voltage-gated Ca2+ channels in their membrane, and maintain Ca2+ at greater levels outside the cell with the exception of ICa2+ compartments. Hepatocytes used calcium pumps to direct the flow of calcium ions through the plasma membrane. This event causes Ca2+ ions to move from the extracellular fluid, and from intracellular storage sites, into the cytoplasm. In a previous study that used a mammalian hepatoma cell line (supplemented with 10% fetal calf serum), the application of a 1 or 10 Hz electric field to human hepatoma (Hep3B) cells induced a fourfold increase in ICa2+[40]. The consequence of increased cytosolic Ca2+ is activation of a class of enzymes known as protein kinases, positive regulators of cell cycle progression. Some members of the protein kinase family play a role in the control of cell proliferation and survival. In the present study, a possible explanation for increased cell viability in HGM-treated hepatocytes is that UWBR activates Ca2+ pumps; thereby releasing ICa2+ and increasing cytosolic Ca2+.It can therefore be concluded from the findings of this research that: 1) Low-dose UWBR caused a mitogenic response in AML-12 mouse hepatocytes. 2) During the post-exposure period of 8–24 h, a dose-dependent increase in cell viability was demonstrated in HGM-treated hepatocytes. 3) In the presence of HGM- and SFM-treated hepatocytes coupled with ITS, a dose-dependent response was also demonstrated with regard to cell viability. 4) Low-dose UWBR caused 1.5-fold (p<0.05) increase in cell viability at 2.5 μg ITS/mL in HGM-treated AML-12 mouse hepatocytes. 5) Cell proliferation activity in AML-12 mouse was confirmed by UWBR ability to up-regulate the 60-kDa cyclin A protein.UWBR Exposure Facility at LA Tech University. Nanopulse facility was designed by Donald T. Haynie. The large room is shielded with copper and steel plating and contains a GTEM cell and pulser. A D-dot probe is used to determine pulse characteristics in the GTEM cell. The ouput of the probe is proportional to the time derivative of the electric field. The pulser is operated from a second copper mesh, RF-shielded enclosure. A 6 GHz bandwidth digital storage oscilloscope (Tektronic Model TDS 6400) is used to monitor pulse experiments. Cable termination panels on both rooms allow room-to-room electrical connections.Mitogenic effect of UWBR on HGM-treated AML-12 mouse hepatocytes. HGM-treated AML-12 mouse hepatocytes were exposed to UWB pulses for 2 h at a temp of 23°C, pulse width of 10 ns, repetition rate of 1 kHz, and applied field strength of 5–20 kV/m. Cell viability was assessed following a post-exposure incubation period of 8–24 h. In this assay, viable cells converted MTT to a water-insoluble formazan dye, as indicated in the methodology section. Bars are means ± SDs, n=3. *Significantly different from control, p ≤ 0.05.Effect of UWBR on HGM and SFM AML-12 mouse hepatocytes. HGM and SFM AML-12 mouse hepatocytes were exposed to UWB pulses for 2 h at a temp of 23°C, pulse width of 10 ns, repetition rate of 1 kHz, and applied field strength of 5–20 kV/m. Cell viability was assessed following a post-exposure incubation period of 24 h. In this assay, viable cells converted MTT to a water-insoluble formazan dye, as indicated in the methodology section. Bars are means ± SDs, n=3. *Significantly different from control, p ≤ 0.05.Effect of UWBR on HGM- and SFM-treated AML-12 mouse hepatocytes in the presence of ITS. HGM- and SFM-treated AML-12 mouse hepatocytes in the presence of ITS were exposed to UWB pulses for 2 h at a temp of 23°C, pulse width of 10 ns, repetition rate of 1 kHz, and applied field strength of 5–20 kV/m. Cell viability was assessed following a post-exposure incubation period of 24 h. Data in this figure indicate a biphasic relationship with regard to cell viability in serum-treated cells. The mean levels of cell viability were: 119 ± .16%, 127 ± .18%, and 147 ± .22% in .625, 1.25 and 2.5 μg ITS/mL, respectively. Bars are means ± SDs, n=2 with 32 replications per concentration. *Significantly different from control, p ≤ 0.05.Western blot for cyclin-A expression and densitometric analysis in UWBR-treated AML-12 mouse hepatocytes. AML-12 mouse hepatocytes were exposed to UWB pulses for 2 h at a temp of 23°C, pulse width of 10 ns, repetition rate of 1 kHz, and applied field strength of 5–20 kV/m. Cyclin A protein identification was assessed following a post-exposure incubation period of 24 h. Inset shows representative Western blot analysis. Densitometric analysis shows an increase in cyclin A abundance at 43,221.This research was financially supported in part by a subcontract from LA-Tech University (Ruston, LA) under the Air Force Office of Scientific Research (Contract No. F49620-02-1-0136), and in part by a grant through the Title III Research Excellence Fund at Grambling State University, Grambling, LA. We thank Dr. Matthew Ware, Chairman of Physics Department, Grambling State University (Grambling, LA) for his valuable support and advice on this research.
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+ Nanotechnology and nanomaterials have become the new frontier world-wide over the past few years and prospects for the production and novel uses of large quantities of carbon nanotubes in particular are becoming an increasing reality. Correspondingly, the potential health risks for these and other nanoparticulate materials have been of considerable concern. Toxicological studies, while sparse, have been concerned with virtually uncharacterized, single wall carbon nanotubes, and the conclusions have been conflicting and uncertain. In this research we performed viability assays on a murine lung macrophage cell line to assess the comparative cytotoxicity of commercial, single wall carbon nanotubes (ropes) and two different multiwall carbon nanotube samples; utilizing chrysotile asbestos nanotubes and black carbon nanoaggregates as toxicity standards. These nanotube materials were completely characterized by transmission electron microscopy and observed to be aggregates ranging from 1 to 2 μm in mean diameter, with closed ends. The cytotoxicity data indicated a strong concentration relationship and toxicity for all the carbon nanotube materials relative to the asbestos nanotubes and black carbon. A commercial multiwall carbon nanotube aggregate exhibiting this significant cell response was observed to be identical in structure to multiwall carbon nanotube aggregates demonstrated to be ubiquitous in the environment, and especially in indoor environments, where natural gas or propane cooking stoves exist. Correspondingly, preliminary epidemiological data, although sparse, indicate a correlation between asthma incidence or classification, and exposure to gas stoves. These results suggest a number of novel epidemiological and etiological avenues for asthma triggers and related respiratory or other environmental health effects, especially since indoor number concentrations for multiwall carbon nanotube aggregates is at least 10 times the outdoor concentration, and virtually all gas combustion processes are variously effective sources. These results also raise concerns for manufactured carbon nanotube aggregates, and related fullerene nanoparticles.While nanotechnology and especially the development of nanoparticulate materials has emerged as the new global focus for a wide spectrum of basic sciences and applied engineering, it has caused a corresponding concern for the potential health risks both in the manufacturing microenvironments and the ambient air; both indoor and outdoor. Some groups have called for a moratorium on nanotechnology-related research, especially nanoparticulate production and use [1, 2]. Carbon nanotubes, especially single-wall carbon nanotubes having extraordinary electrical, thermal, and mechanical properties (purported to be ~100 times stronger than steel at only 17% of its weight [3]), are especially promising nanomaterials. Correspondingly, these and related nanoparticulates have raised concerns for the implications for health effects. A critical factor here involves not only the potential risk (implicit in demonstrated toxicity), but especially the exposure.It is now becoming apparent that in general ultrafine or nanoparticulate matter (PM) designated PM0.1, indicative of particulates smaller than 0.1μm in diameter, poses a greater health risk than fine PM or larger (>PM2.5 which is the standard designation for PM larger than 2.5 μm in mean geometric or aerodynamic diameter) [4–7]. In addition to the nano PM risk, Momarca, et al. [8] also demonstrated that fine crystalline silica (SiO2) had a more detrimental effect on lung epithelial cell damage than fine amorphous silica PM. In addition, recent outdoor airborne particulate analyses have shown that more than 90% of the PM less than 1 μm in diameter (PM1) is crystalline, and 80% of this PM1 range are aggregates ranging from as few as 2 primary particles to more than 1000 primary particles; with morphologies ranging from dense, symmetric aggregates to complex, fractal-like, branched, cluster geometries [9–11]. Warheit [12] has recently noted that the total lung toxicity database for comparing the effects of nano PM (<0.1 μm) with fine PM (<2.5 μm) is composed of studies on only three particle types: titanium dioxide (TiO2 – rutile), black carbon (BC), and diesel particulate matter (DPM) [7, 13, 14].Only two recent studies have begun to explore the toxicological features of single wall carbon nanotubes (SWCNTs). Lam, et al. [15] concluded that SWCNTs are much more toxic to mouse lungs than black carbon, and possibly more toxic than silica; in the nanoparticulate regime. Warheit, et al. [16] correspondingly demonstrated that high dose, intratrachially instilled SWCNTs in rat lung produced transient inflammatory cell injury in contrast to ultra-fine silica particulate exposures. Warheit [12] has more recently cautioned that although rat studies can provide potential hazard estimates, they do not necessarily represent human risk or exposure. Furthermore, the single wall carbon nanotubes in these studies were essentially uncharacterized, and it is unclear whether they were indeed unaggregated, single wall nanotubes, aggregated fibrils or so-called nano-ropes or mixtures (bundles) of nanotubes and other fullerenic polyhedra and metal catalyst nanoparticles. PM sizes or morphologies were unspecified, and comparison of presumed nanotube or fibril morphologies with aggregated carbon spherules of carbon black (as demonstrated in the recent work of Murr, et al. [17]) in the work of Lam, et al. [15] and Warheit, et al. [16] is also absent.Humans have been exposed to nanoparticulates, especially natural environmental nanoparticulates such as mineral species, combustion products, and other anthropogenic nanoparticulates for millennia, if not longer. It has recently been observed in fact that aggregated multi-wall carbon nanotubes (MWCNTs) and other fullerenic polyhedra are not only ubiquitous in the contemporary, outdoor environment, but homes with gas cooking ranges (which are sources of these aggregates) may contain two orders of magnitude more of these aggregates (in number concentration) than the ambient, outdoor air [11, 17]. Furthermore, and as intimated above, Esquivel and Murr [18] have observed aggregates of MWCNTs and other fullerenic nanoparticles in a 10,000 year-old Greenland ice core sample; indicative of the fact that these carbon nanoforms have been a component of the natural atmospheric combustion product regime in antiquity as well.In this research we performed viability assays on a murine lung macrophage cell line to assess the comparative cytotoxicity of commercial, manufactured SWCNTs and two different MWCNT samples; utilizing chrysotile asbestos nanotubes and commercial black carbon as toxicity standards. These nanotube and nanoparticulate samples were fully characterized by transmission electron microscopy. We also collected and analyzed very preliminary clinical (epidemiologic) data relating asthma classifications and prevalence of home gas cooking stove exposure in a mixed patient group ranging in age from 11 to 89 years. In addition, cytokine production (IL-10 and IL-12) was also investigated for the corresponding cytotoxicity assays. Finally, we have examined the implications of this preliminary data and the carbon nanotube studies in the context of potentially new etiological approaches to environmental health issues involving anthropogenic carbon nanotubes in particular and manufactured nanoparticulate materials, especially carbon nanomaterials, in general.Three different, manufactured carbon nanotube materials were examined in this study. A single wall carbon nanotube powder was purchased from Carbon Nanotechnologies, Inc. in Houston, Texas. This material, designated SWCNT-1 contained roughly 5% to 10% iron catalyst as an impurity. Two different commercially manufactured multi-wall carbon nanotube materials were purchased from Rosseter Holdings, Ltd., Limassol, Cyprus, (designated MWCNT-R), and from Nanolab, Inc., Newton, Massachusetts (designated MWCNT-N). Additionally, a high quality chrysotile asbestos (Mg3Si2O5(OH)4) mineral specimen was obtained as a toxicity comparator because fiber or fibrous particulates are characteristically different in their aerodynamic behavior in airway systems as well as their interactions with cells and tissue as a consequence of their generally large aspect ratios [19]. It is also well established that chrysotile asbestos exhibits predictable effects ranging from a cell irritant to severe toxicities involving carcinogenic responses in humans [20,21]. Consequently since carbon nanotubes are generally fibrous with correspondingly large aspect ratios, controls or comparisons with other carbon nanoparticles such as black carbons seemed morphologically unrepresentative. Furthermore, it has already been demonstrated by Murr and Soto [22] that chrysotile asbestos is essentially indistinguishable, microstructurally, from common forms of multiwall carbon nanotubes. Service [23] has earlier raised the prospects for carbon nanotube toxicity considering this resemblance.Finally, a black carbon (BC) (Vulcan XC-72) manufactured by Cabot Corporation, Billerica, Massachusetts, was also examined and tested in this program because its distinctly different nanostructure allows for a meaningful comparison with the nanofiber/nanotube morphologies described above. In addition, and as noted previously, BC is also one of the three particle types, in addition to TiO2 and diesel particulate matter (DPM) that compose a database for comparing total lung toxicity effects for fine and ultrafine particles [13].Each of the experimental nanomaterials described above were fully characterized by examination in the TEM by sprinkling very small and representative quantities onto either carbon/formvar-coated, 3 mm (200 mesh Ni) TEM grids (especially the chrysotile asbestos), or silicon monoxide/formvar-coated, 3 mm (100 mesh Cu) TEM grids. An identical grid was placed on top of these sample grids to form a grid sandwich which constrained the nanomaterials sample within the TEM.Observations of the experimental nanomaterials were made in a Hitachi H-8000 analytical TEM operated at 200 kV accelerating potential, and fitted with a goniometer-tilt (±20°) stage, and a Noran, light-window (Be) energy-dispersive (X-ray) spectrometer (EDS). The general crystallinity or quasi-crystalline features of these nanomaterials samples was also evaluated from selected-area electron diffraction (SAED) patterns [24]. Normal, direct image magnifications ranged from 20,000 to 100,000 times, and enlargements of TEM negatives by 10 times produced image magnifications up to 1,000,000 times.All materials were suspended in a stock solution at 5μg/mL in DMSO. Murine alveolar macrophages (RAW 267.9 cells) were cultured in 96-well flat-bottom plates (50,000cells/well) in the presence of decreasing concentrations of compound (starting at 10μg/mL with 11 doubling dilutions thereafter). Controls were incubated with equivalent dilutions of vehicle (DMSO) and with neither vehicle nor compound. The cells were cultured in DMEM, 10% FCS, 5 × 10−5 M 2-Me, penicillin, streptomycin, and 2mM glutamine at 37°C, 5% CO2. After 48 hours of incubation, 20μL of MTT (3-(4,-dimethylthiazol-2-yl)-2,5-diphenyl-tetrazolium) (5 μg/mL in H2O) (Sigma-Aldrich Co., St. Louis, MO) was added and the cells were incubated for an additional 4–6 hrs. at which time 180 μL of supernatant was removed and 50 μL of lysis buffer, containing 10 N HCl in isopropanol, was added. After several minutes, the MTT crystals formed were solubilized with gentle pipetting and the content of dissolved MTT crystals was measured with a Molecular Devices VersaMax tunable microplate reader set at 570nm. Cell viability assessments or mitochondria/activity of living cells were made by measuring the relative absorbance or optical density (O.D.) for mitochondrial dehydrogenase-transformed formazan (or color product). The data were graphically presented as means ± standard errors of the means.For detection of IL-12 and IL-10 in culture supernatants, viability assays were performed as previously described. At 48 hr. supernatants were collected. IL-12 and IL-10 were detected by sandwich ELISA using the BD Pharmingen cytokine detection kits (BD Pharmingen). Briefly, microtiter ELISA plates were coated overnight with capture antibody at 4°C. Plates were blocked at room temperature with 3% bovine serum albumin in PBS for 2–3 hours. After washing, samples were added neat, in duplicate, for 2 h at room temperature. The plates were later incubated with biotin-conjugated anti-cytokine antibody and were subsequently incubated with horseradish peroxidase-labeled avidin (Vector Laboratories). The enzyme substrate O-phenylenediamine (Sigma) was utilized for color development. Cytokine concentrations were calculated against murine recombinant cytokines (BD Pharmingen).Asthma patient data was collected by having clinical patients fill out a 2-page questionnaire randomly selected at the time of routine clinic/office visits during the Fall and Winter months of 2003–2004. The randomness of those completing the questionnaire also contributed to the overall sample randomness. Principal questions included whether or not patients were currently and continuously exposed to kitchen gas stoves. There may be some intrinsic bias in the data collected since all clinic patients were assumed to have some respiratory health problem. The data selected involved those with some recognized asthma symptoms. Asthma categories were designated severe, moderate, and mild based upon standard office spirometer readings, or forced expiratory volume (of exhaled air) in 1 second: FEV1[25]. The focus on kitchen gas stove use and exposure stems from the collection and observation of carbon nanotubes and related carbon nanoparticulate aggregates in kitchen stove exhausts, and fuel-gas combustion sources in general, in both the indoor and outdoor environments [11, 17].Previous research as noted above has demonstrated that carbon nanotube aggregates, particularly multiwall carbon nanotubes aggregated with other fullerenic nanoparticulates, are ubiquitous in the environment; especially in indoor environments with gas cooking stoves (either natural gas or propane) [17, 26]. As a consequence of these observations, we colleted representative carbon nanotube aggregates in a few kitchens using either natural gas (~96% methane: CH4) or propane gas (C3H8) in order to compare these anthropogenic carbon nanotube aggregate observations with the characteristic carbon nanotube materials utilized in the cytototoxicity studies performed herein. The collections were made using a thermal precipitation device described previously [27, 28]. This device utilizes a thermal gradient effect to adsorb airborne nanoparticulates on the surface of silicon monoxide/formvar-coated, 3 mm, TEM grids which are placed on a cold block inside the device. The grids, with adsorbed nanoparticulates, can be observed directly in the TEM. While the number concentrations of carbon nanotube aggregates collected in indoor environments has been shown to be one or two orders of magnitude greater than the general outdoor environment the number concentrations are nonetheless low; ~104–105/m3.Figure 1 compares the typical microstructures for the SWCNT-1 sample (Figure 1(a)) and MWCNT-N sample (Figure 1(b)). Figure 1(a) shows the SWCNT microstructure to consist of bundles or so-called ropes of single-wall carbon nanotubes which are aggregated with residual iron catalyst nanoparticles and other nanoparticle contamination especially observable in the magnified insert in Figure 1(a). These ropes form tangled aggregates which range in mean (or geometric) diameters from 2μm to 20 μm; with an average aggregate diameter of ~10 μm. The ropes are composed of several to tens of SWCNTS (arrow in Figure 1(a) insert) and range in thickness (or diameter) from 10 nm to about 50nm.Figure 1(b) shows, in contrast to the complex, rope aggregates of SWCNTs in Figure 1(a), more regular MWCNTs which are also aggregated with a small fraction (~10–15%) of other carbon (fullerenic) nanopolyhedra. These MWCNT aggregates (a portion of which is represented by the insert in Figure 1(b)) range in size from about 1μm to 3μm in diameter; with individual MWCNT diameters (Figure 1(b)) ranging from 5nm to 30nm. The MWCNTs range in length from roughly 30 nm (representing short concentric cylindrical shells) to more than 3 μm; with corresponding aspect ratios ranging from ~4 to >100. The MWCNTs in Figure 1(b) are not straight, but they are more regular than the SWCNT nanorope aggregates represented in Figure 1(a) as recognizable, concentric carbon nanotubes.The complex, interwoven ropes or bundles of SWCNTs shown in Figure 1(a) with a large number of iron catalyst nanoparticles attached are produced by the catalytic decomposition of acetylene (C2H2) or methane (CH4) in the presence of the metal catalyst nanoparticles. In contrast, hydrocarbon gas (usually C2H2) pyrolysis produces the complex, kinked, and interwoven aggregates of MWCNTs shown in Figure 1(b). These two different carbon nanotube forms (Figure 1) therefore represent two different manufacturing routes to produce carbon nanotube aggregates with relatively similar sizes or geometric diameters ranging from 1 to 10 μm, or PM10[29].Figure 2 shows a magnified TEM view of the MWCNT-R material which is distinct from the MWCNT-N material in Figure 1(b): the MWCNTs are generally very straight, and the aggregates of MWCNTs and other, related fullerene polyhedra, shown in the lower magnification insert in Figure 2, are composed of essentially equal parts of MWCNTs and fullerenic nanopolyhedra. This material was produced by carbon arc evaporation which is a different manufacturing route than that for the SWCNT-1 or the MWCNT-R materials [29]. These aggregates range in size from 0.1 μm to 3 μm in mean (geometric) diameter, and the MWCNTs range in size from ~10 nm to 30 nm in diameter. MWCNT fiber lengths range from roughly 50 nm to 1 μm, creating aspect ratios ranging from ~3 to >50. As shown in Figure 2, the fullerene nanoparticles, consisting of many concentric graphene shells, range in size from 20 nm to 100 nm; and there is a nearly equal volume fraction of these fullerene polyhedra and multiwall nanotubes.It can be noted on comparing Figures 1 and 2 that the experimental carbon nanotube samples represent very diverse structural or microstructure regimes ranging from complex, interwoven SWCNT bundles or nanoropes to more regular and identifiable MWCNs to very straight, fiber-like MWCNTs. Each of these carbon nanotube sample materials is aggregated and the aggregates have similar geometric (or mean) sizes. Correspondingly, and as noted, these materials represent a similar diversity in manufacturing processes.Figure 3 shows high and low magnification TEM images for the chrysotile asbestos. It is apparent that the asbestos occurs as more regular or more recognizable nanotube bundles than the MWCNT-R material in Figure 2, but the individual asbestos nanotubes are morphologically indistinguishable from the MWCNTs shown in Fig. 2(b). It can be noted that like the MWCNTs shown in Figure 1(b) and Figure 2, the asbestos nanotubes are capped except when broken, and because of the very long length of asbestos nanotubes they are easily broken. The asbestos bundles, although fibrous in morphology as evident in Figure 3(a) have mean, geometrical sizes (bundle or aggregate sizes) ranging from about 0.5μm to 15μm. The individual asbestos nanotubes range in diameter from 15 nm to ~40 nm (Figure 3(b)) with lengths from ~0.5 μm to more than 15 μm; corresponding to fiber aspect ratios ranging from 50 to >1500.Figure 4 illustrates, for comparison with the three experimental carbon nanotube aggregate test materials shown in Figures 1 and 2, the general appearance of carbon nanotube aggregates collected in kitchen stove-top burner exhausts. It can be noted that these carbon nanotube aggregates are very similar to the MWCNT-R test material shown typically in the TEM images of Figure 2. Propane (Figure 4(a)) and natural gas (Figure 4(b)) combustion sources produce variations of these aggregate mixtures [11, 17, 26], but the indoor MWCNT aggregates represented in the TEM images of Figure 4 always closely resemble the MWCNT-R test material (Figure 2) (compare Figure 2(b) and Figure 4(b)). It can be noted in Figure 4(a) that the aggregate sizes and morphologies for propane and natural gas-produced carbon nanotube and related nanoparticulate aggregates are also characteristic of the arc-evaporation produced MWCNT material, implicit on comparing Figures 2(a) and 4(a).Figure 5(a) shows the BC (Vulcan XC-72) test material to be characterized by complex, branched aggregates of carbon spherules ranging in size from about 10 nm to 50 nm. The SAED pattern insert shows the same graphite diffraction rings shown in the SAED pattern insert in Figure 4(b) for aggregates of carbon nanotubes and fullerene polyhedra, but the diffraction ring intensity profiles exhibit a more diffuse amorphous-like microstructure consistent with the turbostratic, amorphous carbon and fullerene nanoparticle compositions of soots illustrated in recent high-resolution TEM studies by Grieco, et al. [30] and Vander Wal and Tomasek [31]. Correspondingly, Fig. 5(b) shows for comparison with Figure 5(a) typical diesel (truck) soot aggregate characterized by a similar, branched, fractal-like, carbon spherule aggregate; with spherule dimensions also ranging from about 10 nm to 50nm in diameter.The cytotoxic effect of the various nanoparticulate materials on murine alveolar macrophages is shown in Figure 6. As shown in Figure 6 cells exposed to media or DMSO alone are viable, with O.D. readings of approximately unity. The macrophages were exposed to increasing concentrations of each of the nanoparticulate materials, starting at 0.005μg/mL and increasing to 10μg/mL in doubling concentrations. The nanoparticulate materials, including the chrysotile asbestos, began to induce cellular death at a threshold of 2.5μg/mL. Essentially all of the carbon nanoparticulate materials exhibited a cytotoxic response (as relative macrophage cell viability) similar to the chrysotile asbestos.Figure 7 shows the relative cell viability measured as O.D. readings at 450 nm and a concentration of 5 μg/mL indicative of the toxicities of the carbon nanoparticulate materials (relative to the media and DMSO controls) and the chrysotile asbestos. The BC nanoparticulate aggregates are observed to be slightly less toxic than the carbon nanotube aggregates, and the chrysotile asbestos.Supernatants from the cell cultures treated with the nanoparticulate materials were obtained to assess cytokine production by the alveolar macrophage exposure. Macrophages produce two primary cytokines, interleukin (IL)-12 and IL-10, depending upon the insult experienced by the cells. IL-12 is typically produced topromote an inflammatory response to intracellular microbes, and is a key inducer of both innate and cell-mediated immune responses to these microbes. IL-10, on the other hand, is produced typically to inhibit activated macrophages and, correspondingly, maintains homeostatic control of innate and cell-mediated immune reactions. IL-10 is commonly coupled with the induction of T helper type 2 (Th2) responses that mediate removal of helminthes, and is associated with allergic and asthmatic macrophage responses. After several replicated trials, no IL-10 or IL-12 response was observed. Consequently, while the test concentrations above about 2μg/mL caused cell death, the exposure time or concentrations did not stimulate cytokine production.Table 1 shows two dozen patient data entries colleted in this study from clinic questionnaires. These questionnaires were randomly distributed to office visit patients over a roughly 6 month period. The actual response (Table 1) was roughly 15% of all patients. The response in Table 1 illustrates that 75% of the respondents were women and half of these women, ranging in age from 11 to 54, were Hispanic. The women respondents ranged from mild to severe asthma based on FEV1 category values. Eighty-three percent of all women in the study are currently exposed to gas cooking stoves in the home while 67% of these women have been continuously exposed to gas kitchen stoves over an extended period of time.All (100%) of the Hispanic women are currently exposed to kitchen gas cooking stoves while 67% of the Hispanic women have been continuously exposed to gas kitchen stoves; the same as for all women included in the sample shown in Table 1. It is interesting to note in Table 1 that none of the Hispanic women exhibit severe asthma. Peters, et al. [32] have recently demonstrated an association between the number concentration of ultrafine or nanoparticulate matter and peak expiratory flow rate (PEFR) or the forced expiratory flow between 25% and 75% of vital capacity among asthmatic adults. While the FEV1 value recorded in Table 1 is a more reliable indication of obstruction than the PEFR value [25], specific associations with number concentrations of particulates were not studied specifically. The gas stove or gas combustion product exposure indicated in Table 1, and as illustrated in the TEM images of thermophoretically collected, gas-stove-generated carbon nanotube aggregates shown in Figure 4, may imply some association with asthmatic adults, especially women, since the multiwall carbon nanotube material examined in this study (Figure 2: MWCNT-R), which closely resembles the kitchen gas-stove-produced multiwall carbon nanotube aggregates (Figure 4), exhibited a cytotoxic response similar to chrysotile asbestos. This is significant because this manufactured nanomaterial is essentially a surrogate for anthropogenic carbon nanotube aggregates (compare Figs. 2 and 3).Figure 8(a) shows a bar graph representation of the gas exposure response for women, including Hispanic women, from the data in Table 1, while Figure 8(b) shows for comparison the total natural gas use in the United States from 1985 to 1995 in contrast to asthma deaths per hundred thousand population for the same period. While the asthma deaths illustrate a reduction in contrast to the total gas us in 1995, there is no specific association for gas use and asthma, or asthma deaths. Nonetheless, Figure 8 suggests, especially in the context of the toxicity assay for multiwall carbon nanotube aggregates (Figures 6 and 7), that anthropogenic carbon nanoparticulate aggregates in particular may pose a respiratory health response. Certainly further research, including epidemiological studies, directed toward these issues is needed to assess the implied health effects.Standard cytotoxicity assays in this study have demonstrated that various carbon nanotube aggregates characterizing very different particulate morphologies exhibit cytotoxic responses essentially identical to nanoparticulate black carbon aggregates, as well as chrysotile asbestos fibril aggregates. Although it is somewhat speculative to generalize these results in terms of specific health risks to humans, concerns for potential occupational exposure hazards should at least be acknowledged. While the physiological relevance of these results remains to be determined, the observations that MWCNT aggregates identical in nanostructure and composition to those nanoaggregates produced ubiquitously in the environment exhibit relative cytotoxicity identical to that for chrysotile asbestos, and nanoparticulate black carbon aggregates, certainly raises concerns for environmental health effects not previously considered either epidemiologically or etiologically. This is true especially for the escalating incidences of respiratory diseases, particularly asthma.While murine alveolar macrophages can provide an in vitro model of potential in vivo effects upon inhalation of the nanoparticulate materials, and also allow for the in vitro experiments to be translated to an in vivo model of exposure (a murine in viro model), cell culture studies in general do not include any effects related to the interaction of the respiratory tract and nervous system with the nanoparticulates. Even experiments that simulate physiologically relevant phenomena simulate only a fraction of the biochemical reactions that take place in a living organism. Correspondingly, and as Warheit [12] (among others) has pointed out, mouse or rat pulmonary response to inhaled particulates is considerably different from larger mammals, particularly humans. Nonetheless, the comparative cytotoxicity assays reported herein can begin to provide some preliminary assessments for the potential pulmonary responses for nanoparticulate aggregates. Correspondingly, it would appear that inhalation threshold limits for airborne chrysotile asbestos should be applied to BC and carbon nanotubes, both single wall and multiwall, regardless of the specific aggregate or primary particle morphologies.Because of the diversity of nanoparticle aggregates and their morphologies (Figures 1–5), especially amongst the carbonaceous nanomaterials, it does not seem likely that the cytotoxicities and related insults for murine macrophage cells are morphologically specific since the BC response is essentially the same as the MWCNT-R material as well as the chrysotile asbestos. Indeed, amongst the common asbestos forms, chrysotile asbestos is the least toxic, and demonstrates the least potential to induce mesothelioma [21, 33]. This is due mainly to its chemistry since other asbestos forms are also nanofibril aggregates; although chrysotile is a nanotube as shown in Figure 4. Like the synergistic interactions for cigarette smoking and asbestos exposure [34, 35], other nanoparticulate materials such as anthropogenic carbon nanotube aggregates and cigarette smoke or diesel nanoparticle aggregates (Figure 5(b)) and PAHs may exhibit similar synergistic interactions in the context of respiratory tract insults or asthma.Of the fibrous particulate matter associated with general toxicity in the environment, especially occupational environments, asbestos is probably the most studied and most notorious. Epidemiologic as well as animal studies have shown that inhaled asbestos can result in pulmonary fibrosis, lung cancer, mesothelioma, and other pleuropulmonary disorders [21, 36]. Chrysotile asbestos is the most common type of asbestos worldwide and represents roughly 95% of fibrous asbestos uses in thousands of commercial products worldwide [37]. As illustrated in Fig. 3, chrysotile is a crystalline, nanotube, serpentine mineral (Mg3Si2O5(OH)4). Commercial amphibole asbestos mineral fibrils, namely amosite ((Fe, Mg)7 Si8O22(OH)2) or crocidolite (Na2Fe5Si8O22(OH,F)2), which account for the other 5% of fibrous asbestos use, have a relatively higher toxicity and pathogenicity than chrysotile because of the iron content [21]. These fibers are not nanotubes like the chrysotile and the exact mechanisms of injury and disease development caused by asbestos fibers remain unclear, but generation of oxidants due to cell uptake and subsequent interactions has been shown to be of fundamental importance in many cellular responses to asbestos [21,38]. To the extent that iron content in asbestos nanofibers plays a pathogenic role, the iron nanoparticulates aggregated with the carbon nanoropes in the SWCNT-1 material (Fig. 1(a)) may also contribute to the characteristic cytotoxicity, although this would not be the case for the two MWCNT samples.The occupational exposure limits for asbestos have varied worldwide between about 1 and 10 fibers/mL (106 to 107 fibers/m3) ingested over an 8h period [38, 39]. MWCNT and other fullerene nanoform aggregates collected on kitchen stove tops as illustrated in Fig. 5 produce particle number concentrations of ~104 to 105/m3 for these aggregates. However, since the larger aggregates can contain up to 103 loosely bound MWCNT fibers, the potential home kitchen exposure could exceed these limits if only 10% of the individual carbon nanotubes were released into the deep lung regime during respiration for about the time it takes to cook dinner (~30 min; × 0.01m3/min. of respired air × 100 MWCNTs × ~104 aggregates/m3 ≅ 0.3 × 106 carbon nanofibers/m3). Furthermore, in home kitchens, the point of emission and point of exposure are little changed in contrast to many particulates emitted into the outdoor atmosphere, including carbon nanotube aggregates. Other than occupational exposures to asbestos and crystalline silica, another cytotoxic particle type classified as a human carcinogen [40], the long term effects of nanoparticulate materials are little known. Correspondingly little is known of the long term effects of low concentrations or levels of specific air pollutants, especially nanoparticulates. Heinrich, et al. [41] have discussed the increasing respiratory symptoms in children with generally declining ambient (outdoor) air pollution while Jedrychowski, et al. [42] have shown that low levels of ambient air pollution can have adverse effects on lung function growth in children. However, specific particulate matter was not considered.There are advantages and disadvantages in using a murine macrophage cell line in the present studies. The main advantage is the consistency of the assay and the simplicity, independent of interconnected metabolic and pulmonary functionality. The biochemistry of single cells, especially cell lines, can differ from the responses occurring in specific animal tissue, and especially in the normal human body. Correspondingly, the disadvantage is that cell culture studies do not include any effects related to the interaction of the specific particulates being studied with the respiratory tract and associated systems and organs. In addition, and as noted recently by Warheit [12], rat and mouse lung instillation studies also have their limitations when toxicity assessments are made for human lung responses to particulate matter. Nikula, et al. [43] have shown that rat lungs process inhaled particulates very differently from larger mammals, especially humans. Consequently, while some relative hazard or figure of merit for potential toxicity may be determined as we have illustrated herein, the issue of exposure or the exposure component, including the exposure length and concentration, is largely unknown, and the risk for specific particulates is difficult to establish, especially for humans predisposed to respiratory or related effects.As pointed out by Warheit [12] risk assessment involves the actual hazard (or toxicity) plus the exposure (or exposure time). A risk paradigm, illustrating the intervening steps between sources of toxic PM and adverse health effects, was recently developed by the National Research Council (USA) in the form [44]:A corresponding and widely used exposure model is also given by Spengler, et al. [45]:where, Ê is the mean exposure over N microenvironments, fj is the fraction of time spent in the jth microenvironment, and Cj is the (jth) microenvironment concentration (of the particulate or toxic agent).While concentrations are often linked statistically by epidemiology to health effects, the sequential linkage across the risk paradigm requires understanding the intervening relationships, especially concentrations of toxic PM and the time of human exposure in specific microenvironments, especially the home; since the average adult spends roughly 70% of the day indoors and as much as 90% of the day in the combined home and work microenvironments. Those with health problems and the elderly would normally spend an even greater percentage of the day indoors [46–48].It may be important to collect sufficient quantities of MWCNT nanoaggregates in home (kitchen) environments and to perform cytoxicity assays directly on these collected samples. Techniques are now available to allow for the selective collection and concentration of the nanoparticulate fraction of sampled air [49–51]. In addition, more statistically significant and more comprehensive clinical data linking respiratory diseases and asthma to sources and actual exposure to MWCNT aggregates will be necessary in establishing exposure-response relationships. More extensive pulmonary bioassays especially utilizing human macrophage cell lines may also provide necessary support for the proposed source linkages, and the health effects hypothesis implicit in these preliminary findings.This study has shown that nanoparticulate aggregates of BC, SWCNT’s as aggregated, complex bundles or ropes containing nanoparticles of Fe catalyst, and MWCNT’s, especially aggregates of MWCNT’s and other fullerene nanoparticles similar to the nanoparticle aggregates produced by many common fuel-gas/air combustion sources such as kitchen gas stove-top burners, exhibit the same toxicity to murine alveolar macrophage cells as chrysotile asbestos nanotube bundles; at concentrations above about 2μg/mL. The chrysotile asbestos nanotube material along with the BC and the MWCNT material which emulates (as a surrogate) anthropogenic MWCNT aggregates in both the indoor and outdoor air inhibited survival of murine macrophage cells. Preliminary collections and analysis of clinical data for asthmatic patients revealed that 83% of women ages 11 to 54 and with mild to severe asthma are currently exposed to kitchen gas cooking stoves in the home while 67% have been continuously exposed. Correspondingly, 100% of Hispanic women are currently exposed to kitchen gas cooking stoves while 67%, the same as for all women in the study, have been continuously exposed. The implications suggest that the anthropogenic occurrences of carbon nanotubes and related nanoparticulate aggregates and exposures in various microclimates may contribute to allergies and/or asthma in humans, especially for long-term exposure. In addition, at sufficiently high concentrations or for very long exposure times, carbon nanoparticulate materials may cause other pulmonary health effects. The corresponding toxicity contrast for chrysotile asbestos also raises concerns for manufactured carbon nanomaterials as well. While murine lung macrophage cell lines as a toxicity assay basis are a long way from representing human lung function, the similarity of the cytotoxic response referenced to chrysotile asbestos may provide some compelling rationale for potential health effects in humans given the significant volumes of evidence for human pulmonary health effects of asbestos [21, 33].TEM bright-field images of SWCNT-1 and MWCNT-N carbon nanotube aggregates.(a) SWCNT-1 sample showing low-magnification view of typical nanostructure and complex aggregated morphology. The insert shows a magnified view of nanoropes and attached Fe catalyst nanoparticles.(b) MWCNT-N structure showing individual, kinked, aggregate structure of capped, multiwall carbon nanotubes and other fullerenic nanoforms. The insert shows a lower magnification view to provide an aggregate size context.TEM bright-field image of MWCNT-R test sample aggregate(a) Magnified view of a section (marked by arrow) illustrating the multiwall carbon nanotube structure, especially tube end caps and other fullerene nanopolyhedra(b) The SAED pattern insert attests to the crystalline-like diffraction features for these aggregated nanoparticles. The bright, lowest-order diffraction ring is graphite (002).TEM bright-field images of MWCNT aggregates collected form the exhaust streams of kitchen stove-top burners by thermal precipitation.(a) Propane-air flame exhaust product showing aggregated MWCNTs and other fullerene nanopolyhedra.(b) Natural gas-air flame exhaust product showing magnified views of MWCNTs and other fullerene polyhedra along the aggregate edge.TEM bright-field images of chrysotile asbestos nanotube bundles (From Murr and Soto [22]).(a) Individual nanotube structure and end caps.(b) SAED pattern shows diffraction streaks and spots (perpendicular to the nanotube axes)(c) A magnified view of the nanotube end caps in (a).TEM bright-field image for black carbon (BC) (VULCAN-XC-72) (After Murr, et al. [17]).(a) Aggregates composed of complex, branched, turbostratic carbon spherules. The SAED pattern insert shows diffuse diffraction (graphitic) rings characteristic of the turbostratic (mostly amorphous) spherule structure.(b) A typical aggregate of carbon spherules characteristic of diesel (bus) soot particulates collected in the outdoor air by thermal precipitation.Cytotoxicity (murine macrophage cell relative viability) versus concentration (μg/mL) for carbonaceous nanoparticulate aggregates and chrysotile asbestos nanotube aggregates.Comparative cytotoxicities of carbonaceous nanoparticulate aggregates and chrysotile asbestos nanotube aggregates to murine macrophage cells referenced to the media and DMSO solutions at a concentration of 5μg/mL. The relative cell viability was measured by optical densitometry at 570 nm.Asthma and Natural Gas Use. U.S. natural gas use data in (a) is from the U.S. Department of Energy, Energy Information Administration: http//www.eia.doe.gov/neic/infosheets/natgasconsumption.htm. The asthma data in (b) includes both male and female of all ages, age adjusted to the 1940 U.S. standard population; from the U.S. National Center for Heath Statistics Annual Summary of Vital Statistics.(a) Current and continuous gas use and exposure in the home (kitchen stove) for asthmatic females and Hispanic females (as a subpopulation) from Table 1.(b) Natural gas use in the U.S. (in trillion cubic feet/year) and asthma deaths in the U.S. per hundred thousand population/year.Clinical Response – Asthma Incidence Vs. Gas Stove (Kitchen) Exposures (2003/2004)FEV1 and ΔFEV1 indicate forced expiratory air volume per second [25] and the patient recorded change in this value, respectively. FEV1 values generally above 80 are designated mild, and those below 60 are designated severe as an asthma classification. Those values between 60 and 80 are designated moderate.This research was supported by a number of programs, grants, and contracts as follows: a Mr. and Mrs. MacIntosh Murchison Endowed Chair (LEM) a Southwest Center for Environmental Research and Policy (SCERP) Grant (Projects A-02-5 and A-04-1) (LEM, DM, PAG, DAL), Research Centers at Minority Institutions (RCMI) Grant (G12RR08124) (KMG, AC, TGP) a University of Texas System Louis Stokes Alliance for Minority Participation (LSAMP) Bridges to Doctorate Fellowship (KFS) and a Dodson Scholars Program at the University of Texas at El Paso (DAR, PAG, DAZ).
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+ Cyanobacterial toxins have caused human poisoning in the Americas, Europe and Australia. There is accumulating evidence that they are present in treated drinking water supplies when cyanobacterial blooms occur in source waters. With increased population pressure and depleted groundwater reserves, surface water is becoming more used as a raw water source, both from rivers and lakes/reservoirs. Additional nutrients in water which arise from sewage discharge, agricultural run-off or storm water result in overabundance of cyanobacteria, described as a ‘water bloom’. The majority of cyanobacterial water-blooms are of toxic species, producing a diversity of toxins. The most important toxins presenting a risk to the human population are the neurotoxic alkaloids (anatoxins and paralytic shellfish poisons), the cyclic peptide hepatotoxins (microcystins) and the cytotoxic alkaloids (cylindrospermopsins). At the present time the only cyanobacteral toxin family that have been internationally assessed for health risk by the WHO are the microcystins, which cause acute liver injury and are active tumour promoters. Based on sub-chronic studies in rodents and pigs, a provisional Guideline Level for drinking water of 1μg/L of microcystin-LR has been determined. This has been adopted in legislation in countries in Europe, South America and Australasia. This may be revised in the light of future teratogenicity, reproductive toxicity and carcinogenicity studies. The other cyanobacterial toxin which has been proposed for detailed health risk assessment is cylindrospermopsin, a cytotoxic compound which has marked genotoxicity, probable mutagenicity, and is a potential carcinogen. This toxin has caused human poisoning from drinking water, and occurs in water supplies in the USA, Europe, Asia, Australia and South America. An initial health risk assessment is presented with a proposed drinking water Guideline Level of 1μg/L. There is a need for both increased monitoring data for toxins in drinking water and epidemiological studies on adverse health effects in exposed populations to clarify the extent of the health risk.Cyanobacterial toxins are well recognized as a cause of livestock poisoning, which has been extensively reported in the Americas, Europe, Asia and Australasia [1]. Livestock are inevitably vulnerable to poisoning as they are restricted in access to water by topography and by fences, and hence may have no choice but to drink water infested by toxic cyanobacteria. Most stock deaths have resulted from the formation of cyanobacterial water-blooms in ponds and lakes on farms, but several major poisoning events were through water blooms on rivers and drinking water reservoirs [2, 3].Human poisoning has also occurred, but the reports are less well documented. The symptoms of poisoning by the main toxic cyanobacteria in drinking water reservoirs overlap with a range of other gastrointestinal illnesses, largely caused by infectious disease organisms. As a consequence during an outbreak of enteric disease the pathogens are investigated first, as the most probable cause, and only after exhaustive exploration are toxins of any type evaluated. Agricultural chemicals and industrial pollutants such as heavy metals are likely to be next suspected, with cyanobacterial toxins ignored until well after the event [4].Epidemiological data for human poisoning by cyanobacterial toxins only exists for a small number of events. The most well characterized case was the poisoning of renal dialysis patients in a clinic in Caruaru, Brazil, in 1996. In this instance the patients treated in a dialysis clinic during one week suffered severe illness following perfusion, with hepatic failure and, in more than 50 cases, death. Investigation of the water treatment unit at the clinic found contamination of the filters by two types of cyanobacterial toxin, microcystins and cylindrospermopsins [5, 6]. Microcystins were detected in the blood and liver of poisoned individuals [7]. Because of the severity of the poisoning a thorough investigation was carried out, which showed up major defects in the operation of the water treatment unit at the clinic. Exposure to toxins through renal dialysis is a particularly potent route of poisoning, equivalent to an intravenous injection in the case of water soluble toxins. The volume of water used in perfusion is large, about 120L, so that the total amount of toxin to which a dialysis patient is exposed is much greater than possible through drinking water.Exposure to cyanobacterial toxins through consumption of contaminated drinking water has however also resulted in poisoning. The earliest demonstration of this was in 1983, when the population of a rural town in Australia was supplied with drinking water from a reservoir carrying a dense water bloom of a toxic species of cyanobacterium, Microcystis aeruginosa. The toxicity of this water bloom was being monitored in the reservoir. The controlling authority dosed the reservoir with copper sulphate to destroy the cyanobacteria, which caused the cells to lyse and release toxin into the water. Epidemiological data for liver injury in the affected population, a control population and comparison of the time periods before the bloom, during the bloom and lysis, and afterwards, showed clearly that liver damage had occurred only in the exposed population and only at the time of the water bloom [8]. In another less well characterized event, about 140 children and 10 adults were hospitalized, after the water supply authorities treated a cyanobacterial bloom in a small drinking water supply reservoir with copper sulphate, to resolve taste and odour problems. Within a week severe hepatoenteritis was apparent in the population, with about 20 cases requiring intravenous therapy. No-one died though several children were placed in intensive care [9]. Subsequent investigation demonstrated a “new” toxic cyanobacterial species in the reservoir, with a potent general toxin [10]. Later work on this strain of cyanobacterium lead to the identification of an alkaloid cytotoxin, with considerable liver toxicity [11].Cyanobacteria are a normal component of the worldwide biota, with a wide tolerance of climatic conditions and environment. As a very ancient life-form they occupy every conceivable ecological niche, and their abundance is limited by nutrient and light availability [12]. In aquatic systems cyanobacteria are always present, through the population density varies from very small numbers to more than 106 organisms/mL. There is a strong relationship between phosphorus concentration in the water and cyanobacterial numbers and also a similar though less linked relationship between dissolved nitrate/ammonia and cyanobacteria [13]. Thus in general toxic cyanobacterial species will be presentin all water bodies, with numbers dependent on the available nutrients and light.As human population density rises the inflow of nutrients into water bodies increases through agricultural fertilizer use, urban run-off and sewage discharge. This increase in aquatic nutrients is termed eutrophication, and it is observed worldwide. Phytoplankton in general becomes more abundant, and among these organisms are the cyanobacteria. Cyanobacteria can utilize nutrients competitively with eukaryotic phytoplankton, and will proliferate more successfully at lower nutrient concentrations than the green algae. As a result many rivers, lakes and reservoirs worldwide develop high cyanobacterial cell concentrations, especially in the summer months, which appear as greenish suspensions in the water. Some species float to the surface under warm, still conditions, forming scums with extreme cell concentrations above 1×106 cells/ml. Dried scums often appear blue-green or red through liberation of phycocyanin pigment, leading to the common name of these organisms - blue-green algae.The scum-forming cyanobacterial species are largely toxic, and the majority of domestic animal poisonings have occurred from the animals drinking scum [12]. Cell populations carry over from year to year, and once a reservoir or lake has an established water bloom of cyanobacteria in summer, it is very difficult to reverse this phenomenon. Cyanobacteria proliferate in warmer weather, and often form extensive blooms in late summer. With an increase in global temperatures, cyanobacterial populations are likely to increase also. With the growth in human populations, demand for drinking water has resulted in water is being drawn from water bodies carrying substantial cyanobacterial populations, thus presenting a risk to human populations.Annual or even permanent blooms of toxic cyanobacteria are becoming increasingly common in drinking water reservoirs. To give an illustration the three main reservoirs supplying Brisbane in Australia all carry substantial populations of the toxic Cylindrospermopsis raciborskii. This cyanobacterium forms dense layers 5–10m below the surface, so that the first indication of the proliferation of the organism may be the blocking of filters in the drinking water treatment plant. Other examples are the main drinking water supply reservoirs for the cities of Sao Paulo in Brasil and Lodz in Poland, which contain heavy blooms of the toxic Microcystis aeruginosa in summer.Neurotoxins form one of the major groups of cyanobacterial toxins. They are produced by several genera of cyanobacteria growing in freshwater which have the capacity to form dense waterblooms and floating scums at the edge of lakes and rivers. The neurotoxins are alkaloid compounds, fast acting and have caused many deaths of dogs and livestock [1]. Three types of alkaloid have so far been described (Figure 1).The first to be characterized was anatoxin-a (Fig. 1), a neuromuscular blocking agent which causes death by respiratory paralysis [14]. This toxin has been found in three common genera of cyanobacteria, Anabaena, Aphanizomenon and Planktothrix, all filamentous planktonic organisms capable of high cell concentrations and potential scum formation. There has been no clear evidence of human poisoning from these organisms, though a coroner in Wisconsin in 2003 resolved that the death of a male teenager who was diving and playing in a pond containing neurotoxic Anabaena had died as a consequence of ingestion of these cyanobacteria [15] To verify this cause of death, evidence of toxin in the gastrointestinal tract or tissues was required, however there has been no published report of the presence of toxin. By contrast, dogs poisoned by anatoxin-a have shown the toxin in stomach contents [16]. The compound is stable in the environment, as exhibited by the dogs having died after eating decaying lumps of cyanobacteria on the lakeside. Anatoxin-a has been identified in the water of lakes in North America and in Europe, which are largely used for recreation. There is the possibility of consumption of moderate quantities of water during swimming and especially water skiing, and hence a risk exists for anatoxin-a poisoning of recreational water users [17]. Many authorities in developed countries have warning procedures for cyanobacterial blooms at popular recreational areas, to reduce risk to water users [18]. There has been little attention paid to the assessment of risk to drinking water consumers from anatoxin-a, largely because of the rapid excretion of the toxin from the body, no evidence of residual effects and low free-water concentrations in lakes.Anatoxin-a(s) is much less common in cyanobacterial waterblooms, though it was first identified following cattle deaths in the USA[19]. The alkaloid closely resembles an organophosphorus insecticide (Figure 1), and acts as an anticholinesterase. The characteristic feature of this poisoning is excessive salivation, which is the reason for the designation (s). The compound is highly unstable and unlikely to persist in water supplies, and as a result is also unlikely to present any risk.The saxitoxin-type neurotoxins are well known as the cause of paralytic shellfish poisonings, which have resulted in many hundreds of human deaths worldwide [20]. As a result, legislation controls the allowable concentration of saxitoxins in shellfish harvested for human consumption (80μg/100g fresh shellfish tissue) and there is a substantial monitoring program in many countries. Saxitoxins are however not limited to marine waters, and also occur in freshwater cyanobacteria. Anabaena, Aphanizomenon and Lyngbya genera of cyanobacteria have species that produce saxitoxins [21–23]. The massive waterbloom of Anabaena circinalis on 1,000km of the Darling River in Australia in 1991 killed a large number of sheep and cattle, and also resulted in detectable neurotoxicity in town water supplies [24]. Saxitoxins are heat-stable molecules, which are not easily removed in conventional water treatments unless pH and chlorine residuals are carefully controlled, but can be effectively removed by ozone or activated carbon [25].The toxicity of saxitoxin is considerable, as the alkaloid blocks sodium conduction in axons preventing nerve impulse transmission, leading to paralysis. The oral LD50 in mice is about 260μg/kg bodyweight [26]. Acute poisoning in humans is unlikely to occur from contaminated water supplies, as the human body can tolerate about 100μg of saxitoxin without ill effect [20], which translated to drinking water is 50μg/L assuming 2L water drunk per day. No cumulative effects have been demonstrated, though there is limited evidence of resistance to toxicity in exposed human populations [26]. New Zealand is considering a Maximum Acceptable Value in drinking water of 3μg/L of saxitoxin equivalents in their new drinking water guidelines, which should be ratified shortly [27]. This value has also been proposed in Australia [28]. The data on which this value was based were the intraperitoneal toxicity of saxitoxin to mice, and incorporated a safety factor of 1,000. This issue will be discussed further in the section on hepatotoxins (microcystins). There are no data for the concentrations of saxitoxin-type neurotoxins in drinking water, and there were no reports of neurotoxic symptoms in the town population when neurotoxicity was detected in the drinking water supply [29].These toxins have received the greatest attention, as they are the source of the most likely risk to consumers of drinking water. The predominant genera of cyanobacteria forming the peptide toxins called microcystins are Microcystis, Planktothrix and Anabaena. Species from these genera are common in Europe, the Americas, Africa and Asia, and poisoning of domestic animals has been widely reported [12]. Only two epidemiological investigations have so far shown human injury from microcystin in drinking water, one in Australia [8] and one in China (unpublished). As a consequence of the frequency of cyanobacterial blooms containing hepatotoxins in drinking water reservoirs, the WHO carefully examined the need for the major toxins, the microcystins, to be included in the drinking water guidelines. An ‘expert group’ was established to examine the whole issue of cyanobacterial toxins in drinking water, which resulted in a comprehensive assessment of the risks involved [30]. The outcome was a recommendation that the microcystins should be included among the chemicals for which Guideline Values be determined. These peptide toxins are cyclic, and contain a majority of D-amino acids (Figure 2). The positions shown as [X] and [Y] are L-amino acids, and are variable between species and strains of cyanobacteria. The most abundant variant has L-leucine (L) and L-arginine (R) respectively at [X] and [Y] (microcystin-LR). The amino acid at the left of the molecule is unique, is connected into the ring through an amino group at the β-carbon atom, and has the trivial name of ADDA.Microcystins are resistant to digestion in the gastrointestinal tract of eukaryotes, as peptide bonds linking to the D-amino acids are not susceptible to normal hydrolytic enzymes. The toxins are concentrated into the liver by an active transport system, similar to the bile acid transporter [31]. Microcystins specifically inhibit protein phosphatases 1 and 2A, which have a vital role in cell control and in intracellular structure [12]. Acute poisoning is through destruction of the liver architecture, leading to blood loss into the liver and hemorrhagic shock [32]. Later death is through liver failure with massive destruction of hepatocytes, seen in large animal deaths and human fatalities [5, 33]. Chronic exposure to these toxins in drinking water led to ongoing active liver injury in mice [34].There is experimental evidence for tumour promotion by microcystins, and limited data for carcinogenesis in rodents [35]. In rural areas in Southern China some villages showed hyper-endemic rates of hepatocellular carcinoma, which have been shown to be linked to hepatitis, aflatoxin in the food, and drinking surface water. Microcystins in the ponds and ditches used as water sources were suspected of contributing to the cancer rates [36, 37].These toxins are highly stable in water and are resistant to boiling. Hence they present a risk to consumers in less developed regions and countries who are collecting water from surface sources to drink. Many lakes, ponds, ditches and streams in rural and outer urban areas suffer from eutrophication through excessive nutrient leaching from housing, sewage, and intensive agricultural use, leading to cyanobacterial proliferation. In tropical and temperate regions of the world the genus Microcystis is the most abundant cyanobacterium forming toxic blooms, with toxin concentrations sufficient to poison domestic animals. If these contaminated water sources are used for human consumption, there is a risk of human poisoning. Conventional Western drinking water treatment may not be effective under bloom conditions in removing microcystins from drinking water and hence there is a risk to consumers. Advanced water treatment using ozone and activated carbon will reliably remove microcystins [12].WHO have carried out an assessment of the safe level of microcystins in drinking water, based on data from a subchronic toxicity trial in mice, with supporting data from growing pigs [38]. The calculation used the No Observed Adverse Effect Level for male mice during a 13 week oral toxicity trial, of 40μg of microcystin-LR/Kg/day [39]. This was used to calculate a Tolerable Daily Intake (TDI) for safe human consumption, by the incorporation of uncertainty or safety factors. While these are subjective, a factor of 10 for interspecies uncertainty between rodents and humans, a further 10 for variability in sensitivity between people, and an uncertainty of 10 for inadequate data, possible tumour promotion and lack of lifetime exposure are generally accepted.Thus the:From this value the Guideline Value (also called the reference dose and the maximum acceptable concentration) was calculated from the standard bodyweight of 60Kg, an assumption of the proportion of the dose from drinking water of 0.8 (some may come from food and particularly blue-green algal diet supplements) and a standard water consumption of 2L/day.This value was determined from the toxicity of microcystin-LR, so the WHO Chemical Safety Committee set the Guideline Value for microcystin–LR. Since there are some 60 variants of the molecule, and some highly toxic blooms do not contain any of the –LR variant, it is necessary to interpret this as toxicity equivalent to microcystin-LR. Where this Guideline Value has been adopted as the basis for national legislation, the need for monitoring of all the toxin variants has been recognized and the equivalent total toxicity calculated [40]. Individual countries have also adopted a higher standard bodyweight, and a different proportion of the consumption from drinking water. All of the Guideline Values adopted so far lie between 1 and 2μg/L of microcystin equivalents, which for practical purposes are the same.A larger potential adjustment to this value may result from re-classification of microcystin as a carcinogen, rather than a non-carcinogenic poison. While there is experimental evidence for tumour promotion by microcystin in liver, skin and colon, the only data indicating carcinogenesis have been obtained by continued very high intraperitoneal doses of toxin in mice which cause extensive liver damage [1]. In China there is ongoing investigation into the relationship between surface water consumption and cancer of liver and colon [41]. This issue is discussed in detail elsewhere [12], concluding that there is insufficient evidence at present to determine that microcystin is a probable carcinogen but the possibility requires continual evaluation.These alkaloid cytotoxins were relatively recently discovered, following the widespread human poisoning at Palm Island, Australia due to contamination of the water supply [9]. Cyanobacteria from the supply reservoir were collected, cultured and evaluated for toxicity, showing potent toxicity to liver, kidney, adrenals, lymphoid cells and other tissues in mice [10]. Subsequent investigation of the oral toxicity of the cyanobacterium responsible, Cylindrospermopsis raciborskii, further demonstrated the tissue damage caused by the toxin [42, 43]. The toxic alkaloid was isolated and identified as a potent inhibitor of protein synthesis [11, 44]. There are on-going investigations into the mechanism of cylindrospermopsin toxicity, which may involve activated metabolites of the alkaloid [44, 45]. The alkaloid has several reactive groups, including a hydroxymethyl uracil, which may be vulnerable to biological oxidation (Figure 3).Cylindrospermopsin has been found in water bodies that have blooms of the cyanobacterial species Aphanizomenon ovalisporum[46], and Umezakia natans[47] as well as those with Cylindrospermopsis[48, 49], and recently in Germany, in the absence of any of those species [50]. It is apparent from this data that cylindrospermopsin is likely to occur widely in freshwater sources, and that only a beginning has been made in identifying species producing this toxin. Monitoring of water supplies for cylindrospermopsin has found concentrations in natural water bodies, reservoirs and in drinking water that are above 10μg/L, which is a cause for concern [51, 52].On the basis of the experimental toxicity of cylindrospermopsin, the reported human poisoning associated with Cylindrospermopsis and the toxin concentrations measured in water bodies, it was apparent that risk assessment for this toxin in drinking water was required. A subchronic oral exposure trial of cylindrospermopsin in male mice provided a No Observed Adverse Effect Level of 30μg of cylindrospermopsin/Kg/Day. On the assumption of standard uncertainty factors and the total intake arising from drinking water, a Guideline Value of 1μg/L resulted [53].Examination of the molecular structure of cylindrospermpsin indicated that it may be able to interact with DNA or RNA in cells, through the uracil group, assisted by the planar shape of the molecule. If this proves to be the case, then evaluation of the possible carcinogenicity of the molecule is required. Preliminary data indicated that cylindrospermopsin may form DNA adducts [54], and there is evidence for clastogenicity and micronucleus formation in a cultured human white cell line incubated with cylindrospermopsin [55]. A preliminary trial of carcinogenicity in mice indicated the presence of excess tumors, providing support for a more definitive carcinogenicity trial [56]. It is premature at present to attempt to classify cylindrospermopsin as a possible human carcinogen, because of the very limited current data. Further experimental and epidemiological research on this toxin is required to clarify these issues, and cylindrospermopsin is now on the ‘Candidate Contaminant List’ of the US Environmental Protection Agency.There are two areas in which more data is necessary to make a clear case for national action on minimizing health risks from cyanobacterial toxins. The first is the need for widespread monitoring for the presence of toxic cyanobacterial species and toxins in drinking water sources, to identify the abundance of locations of potential risk. This is in progress in Europe, and an initial survey has been carried out through the American Water Works Association in the USA. From data arising from these surveys, the extent of the problem has become apparent, and the location of a proportion of the populations at most risk.The second and more difficult aspect is the need for epidemiological studies on at-risk populations to quantify the adverse health effects. Exposure biomarkers will have to be developed, in addition to quantitating the toxin concentrations in tap water. The commonly used clinical measures of liver and kidney function, and clinical records for hepatoenteritis, provide relevant health information. The earlier data on population injury from microcystins indicates the clinical parameters of particular interest [8].The cyanobacterial toxins provide a risk to human health when the population of toxic cyanobacteria in drinking water sources rises to bloom proportions. The present assessments of Guideline Values for these toxins as chemical, non-carcinogenic contaminants indicate that a safe concentration in drinking water is in the region of 1μg/L, a concentration that has been exceeded in numerous water storages. Carcinogenicity of these toxins is not yet established, though both microcystins and cylindrospermopsin have caused excess tumors in rodent experiments. With the increased eutrophication of water supplies and global warming, cyanobacterial populations and hence toxic risks are likely to rise in the immediate future. The extent and potential severity of the risks need further evaluation.Structures of cyanobacterial neurotoxic alkaloids.Structure of the peptide hepatotoxin microcystin, first isolated from the cyanobacterium Microcystis aeruginosa.Molecular structure of the cyanobacterial alkaloid toxin cylindrospermopsin. The bridging hydroxyl group may be in either stereochemical position, and may also be replaced by hydrogen.
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+ Ultraviolet (UV)-induced cataracts are becoming a major environmental health concern because of the possible decrease in the stratospheric ozone layer. Experiments were designed to isolate gene(s) affected by UV irradiation in rabbit cornea tissues using fluorescent differential display-reverse transcription-polymerase chain reaction (FDDRT-PCR). The epithelial cells were grown in standard medium for 2 or 4 hours post treatment. Cornea epithelial cells were irradiated with UVB for 20 minutes. RNA was extracted and amplified by reverse transcriptase-polymerase chain reaction using poly A+ specific anchoring primers and random arbitrary primers. Polyacrylamide gel electrophoresis revealed several differentially expressed genes in untreated versus UV irradiated cells. Complimentary DNA (cDNA) fragments resulting from fluorescent differentially expressed mRNAs were eluted from the gel and re-amplified. The re-amplified PCR products were cloned directly into the PCR-TRAP cloning system. These data showed that FDDRT-PCR is a useful technique to elucidate UV-regulated gene expressions. Future experiments will involve sequence analysis of cloned inserts. The identification of these genes through sequence analysis could lead to a better understanding of cataract formation via DNA damage and mechanisms of prevention.The human eye and skin are the only tissues directly exposed to ultraviolet radiation (UV) and visible radiations. The full spectrum of UV radiation can be classified into three groups, based on wavelength, ultraviolet A (UVA) (400 – 315 nm), ultraviolet B (UVB) 315-280 nm, and ultraviolet C (UVC) (280-100 nm). UVB and UVA can reach the surface of the earth causing biochemical and physiological effects depending upon radiant exposure and wavelength [1]. Over 98% of solar UV radiation exposure is in the form of UVA. It penetrates the skin more deeply than UVB or UVC, but is less associated with DNA damage. UVB accounts for less than and 2% of our solar UV radiation exposure, as much of it is absorbed in the upper atmosphere. Although UVB is responsible for most of the DNA damage within skin cells that might lead to the promotion of cancers [1–2], however UVC is considered the most lethal form of UV radiation. The ozone layer effectively blocks most UV. The risk of exposure to harmful UV radiation is gradually increasing due to continuous erosion of the stratospheric ozone layer [3–4]. Studies have shown that excess exposure to ultraviolet radiation can lead to severe abnormalities within radiation-exposed tissues. UV radiation has been shown to damage ocular tissues [5–6] and UVB is known to reach the cornea. Chronic exposure of eyes to UV is heavily implicated in the development of cataracts (for example, opacities of the lens), and can also cause phototoxic effects to the retina [7–9]. Age-related cataract is a multifactorial eye disease, the leading cause of blindness, and is becoming an increasing global problem. An estimated 1% decrease in ozone thickness will enhance the rate of cataracts by 0.7% [10]. Epidemiological studies have shown that cataract formation is associated with prolonged exposure to sunlight. DNA is a major target of UV-induced cellular damage. In addition, several risk factors for cataract formation include diabetes, alcohol, smoking, and steroid use [11–13]. All three wavelengths of naturally occurring UV light, UVA, UVB, and UVC, may directly induce pyrimidine and thymine dimer formation [7,14], as well as DNA strand breaks and DNA-protein cross-linking [15]. It is thought that UV-induced skin cancer may largely result from such DNA damage [16, 17]. UV irradiation is known to induce programmed cell death or apoptosis in the cornea. Apoptosis is a mechanism associated with corneal cell death after UV irradiation [15, 18, 19]. Exposure of mammalian cells to UV radiation triggers an alteration in gene expression [16, 20]. The molecular mechanisms are still unclear and the effects on expression of UV-induced genes involved in DNA damage and cataract formation have not been determined.We employed the fluorescent differential display polymerase chain reaction (FDD-PCR) technique to analyze changes in gene expression between untreated and UVB irradiated corneal epithelial cells. The mRNA differential display (DD) technique developed by Liang and Parde [21] allows the isolation of unknown genes without prior knowledge of their sequence, just on the basis of their cellular abundance. The technique also permits simultaneous identification of up-regulated and down-regulated genes between two groups of cells, tissues, or conditions. It has been successfully used to identify differentially expressed genes in numerous systems including ocular tissues [22–25]. UV light induces the expression of a wide variety of genes. The prevalence of cataracts approximately doubles with each decade after the age of thirty. However, the effects of UV exposure in corneal epithelial cells at the molecular level have not been elucidated. If we are able to delay the onset of age-related cataracts by ten years, the number of cataract operations could be decreased tremendously. The goals of this study were to identify differentially expressed genes in the corneal epithelial cells in response to UVB irradiation and to analyze the changes in gene expression related to DNA damage with subsequent cataract formation.Eyes were purchased from eight to twelve weeks old New Zealand White rabbits. The method of Cubitt et al. [26] was used (with some modification) to culture the corneal epithelial cells. Briefly, corneas were dissected at the limbus, and digested overnight at 4°C by placing the cornea on top of 60 – 80μl of Dispase II (Boehringer Mannheim). The epithelial layer was gently dissected from the stroma using a small spatula, and the epithelial sheets were digested with 1 ml of trypsin (0.05)-EDTA (0.02) (Sigma) and was incubated at 37°C for 10 minutes. Trypsinization was stopped by adding soybean trypsin inhibitor (Sigma) (1.5ml). Single-cell suspensions were made by passing the suspensions four to five times through a syringe with a 23-gauge needle and 15% rabbit serum was added. The cells were centrifuged at 1000 rpm for 3 minutes and the supernatant was discarded. The cells were seeded in three tissue culture flasks (Falcon Primaria positively charged 25 cm2) per two corneas. Three milliliters of media were added to each flask. Cells were cultured in media (Cascade Biologics) supplemented with rabbit corneal growth factors (Cascade Biologics) and gentamicin (Sigma) in a humidified incubator at 37°C with five percent CO2. The epithelial cells were sub-cultured at a passage ratio of 1:3. The majority of the experiments were carried out using confluent cells from second and third passages, although primary cells and fourth passage cells were used in the indicated experiments. However, it should be noted that all cells had the characteristic epithelial cobblestone appearance, and the results were not dependent on passage number.The corneal epithelial cells were grown to 90% confluence. The confluent cells were incubated in fresh medium (4mls) at 37°C for 30 minutes before UV irradiation exposure. The corneal epithelial cells were cultured in Medium 500 and MEM containing fifteen percent rabbit serum, respectively. Two XX-15B lamps (Spectronics) were placed side by side and used as an UVB source. The UVB lamps had a maximum at 310 nm with a wavelength range of 280 – 365 nm. The radiant energy at the level of the cells was measured with a DRC-100H radiometer (Spectronics) with the sensor, DIX-300 for UVB. During UVB irradiation, the lids of the Petri dishes were removed and replaced with quartz plates. The UV dose was administered for 20 minutes for UVB exposure. The irradiated and untreated corneal epithelial cells were incubated at 37°C for 2 or 4 hours, after which they were harvested.Total RNA was isolated from untreated (UT) and UVB exposed corneal epithelial cells using the RNApure reagent (Genhunter, Nashville, TN). The RNA samples were treated with RNase-free DNase I and incubated for 30 minutes to eliminate chromosomal DNA contamination prior to proceeding to the differential display system. The concentration of the RNA was determined spectrophotometrically (absorbency at 260 nm and A260/A280 ratios of 1.7 or higher) [27]. The size distribution and the integrity of the purified total RNA was analyzed by denaturing formaldehyde agarose gel electrophoresis.Fluorescent Differential Display (FDD) using rhodamine (GenHunter, Nashville, TN) was performed according to the manufacturer’s instructions. The RNA was quantitated as previously described and diluted to the appropriate amount with diethylpyrocarbonate water (DEPC-H2O). The mRNA Differential Display method was performed routinely by using 0.1μg of total RNA for reverse transcription reaction. Three reverse transcription reactions for each RNA sample was carried out in thin-walled PCR tubes, and each containing one of the three different rhodamine labeled anchored oligo dT primers (RH-T11M, where M was G) were prepared for each RNA sample isolated from untreated and UV-B induced corneal epithelial cells. Each reaction mix contained RNase-free H2O (9.4μl), 5X reverse transcriptase buffer [125mM Tris-HCl, pH 8.3, 188mM KCl, 7.5mM MgCl2, and 25mM DTT] (4μl), dNTPs [250μM] (1.6μl), RNA [0.1μg/μl] (2μl), and RH-T11M primer (2μl). The thermocycler was programmed as follows: 65°C for five minutes, 37°C for 60 minutes, 75°C for five minutes, followed by a cooling at 4°C. One microliter of MMLV reverse transcriptase (100units/μl) was added to each tube 10 minutes after incubation at 37°C. Throughout the FDD experiments, control RNA (positive control) isolated from transformed rat embryo fibroblasts provided by the kit (GenHunter Corporation) was run simultaneous to compare the efficiency of the system. The fluorescent dyes were light sensitive and were kept in the dark until used.Following the reverse transcription of the RNA, the resulting cDNAs were PCR amplified using various combinations of the rhodamine labeled arbitrary anchored oligo dT primer and an arbitrary decamer. The PCR reactions consisted dH2O (10.2μl), 10X PCR buffer (2μl), dNTP mix (1.6μl), H-AP primer [2μM] (2μl), rhodamine labeled RH-T11M primer (2μl), RT-mix (2 μl) and Taq DNA polymerase (0.2μl) for a total volume of 20μl. The thermocycler was programmed as follows: 94°C for 30 seconds, 40°C for 2 minutes, 72°C for 60 seconds for 40 cycles, followed by 72°C for 5 minutes, and stored at 4°C until further use.PCR products were resolved in parallel lanes on a 6% denaturing polyacrylamide gel in TBE buffer. Gels were pre-run for 30 minutes and 3.5μl of each sample with 2μl of FDD loading dye were mixed and incubated at 80°C for two minutes immediately before loading onto 6% DNA sequencing gels. The gels were electrophoresed for 2–3 hours at 60 watts constant power, and scanned on a Fluorescence Imager (Hitachi FMBIO® II) using 585 nm filters, following manufacturer instructions.Differentially expressed bands were excised from the gels, boiled for 30 minutes and re-amplified with the same primer combinations and PCR conditions, except unlabeled H-T11M anchor primer was used instead of the rhodamine labelled primer. The re-amplification PCR reactions consisted of dH2O (23.3μl), 10X PCR buffer (4μl), NTP mix (0.3μl), H-AP primer [2uM] (4μl), H-T11M [2uM] (4μl), cDNA template (from the gel) (4μl) and Taq DNA polymerase (Qiagen) (0.4μl) for a total volume of 40μl. The re-amplified PCR products were run on 1.5% agarose gels stained with ethidium bromide before cloning. These products were cloned into the PCR-TRAP® cloning system (GenHunter, Nashville, TN) according to manual protocol. The size standards used for colony PCR screening was NEB 100 base pair (bp) ladder (Gibco-BRL, Gaithersburg, MD). The clones selected for sequencing were restreaked onto Luria Broth plates containing tetracycline (LB-tet) for single-colony isolation.For northern blots, total RNAs (5 g) isolated from untreated and UVB irradiated corneal epithelial cells were fractionated on 1.2% agarose gel and transferred onto Hybond-N+ membrane (Amersham) using standard procedures (30). The membrane was hybridized with 32P dCTP-labeled cDNA probe (JS6) at 42°C for 16 hours and extensively washed with 2xSSPE/0.1% SDS, 1xSSPE/0.1% SDS and 0.1xSSPE/0.1% at 42°C. Autoradiography was performed at −70°C overnight.By phase contrast microscopy, the morphology of the cornea epithelial cells appeared changed after UVB exposure. These cells were visualized at a 40X magnification. The untreated corneal epithelial cells (Figure 1A) were healthy and attached with a cobberstone appearance. These cells have been growing for 3–4 days and a few cells have shown signs of apoptosis. After exposing the cells to UVB for 20 minutes (Figure 1B), the cells became separated and elongated due to a loss of membrane integrity. This event may be a result of primary necrosis or secondary apoptosis [28]. Four cells exhibited apoptosis (Figure 1B). Corneal epithelial cells were exposed for 45 minutes (Figure 1C), which resulted in more separation with some rounding of the cells. Some cells were detached and formed dark bodies within the cells. This may be due to necrosis and other forms of DNA damage such as single strand breaks. Twenty-five cells exhibited apoptosis (Figure 1C).The FDD allowed for parallel analysis of four RNA populations. The RNA populations compared were untreated versus UVB irradiated corneal epithelial cells with 2 hours post-treatment and untreated versus UVB irradiated corneal epithelial cells with 4 hours post-treatment. The FDD analysis of the untreated and UVB irradiated corneal epithelial cells indicated a number of similarities in gene expression between untreated and UVB irradiated cells (Figure 2). A large number cDNAs were present in both untreated and UVB irradiated corneal epithelial cells, however, the majority of these genes were not affected by UV radiation exposure. As a result, these cDNAs represent the house keeping genes found in corneal epithelial cells. Figure 2 represents a typical fluorescent image of differentially expressed cDNAs. Several differentially expressed bands were detected in the differential display gels. The eleven differentially expressed bands with the strongest intensities and best resolutions were excised from the gel and re-amplified using the same primer set. Eight of the differentially expressed bands were down-regulated and the other three differentially expressed bands were up-regulated in response to UVB exposure.The differentially expressed bands were electrophoresed on a 1.5% agarose gel and ranged in size from 200 base pairs (bp) to 800 base pairs (Figure 3).The sizes of the differentially expressed bands chosen for reamplification were as follows: JS1, 300 bp; JS2, 550 bp; JS3, 400 bp; JS4, 300 bp; JS5, 280 bp; JS6, 300 bp; JS7, 280 bp; JS9, 200 bp; JS10, 800 bp; JS11, 200 bp; and JS12, 300 bp. The eleven re-amplified bands were cloned into the PCR-TRAP Cloning System. Four colonies for each band were checked for inserts by colony-PCR (Figure 4).Two colonies from each band showing the correct size were re-streaked on LB agar plates containing tetracycline. Differential expression of one of the eleven cDNAs was confirmed by Northern blot analysis (Figure 5) following standard procedures; as indicated in the material and methods section. The cDNA (JS#6) was used as a probe; a distinct band approximately 1.2 kb (Figure 5, lane 1) was detected in the untreated cells whereas, no band was detected in the UVB irradiated corneal epithelial cells (Figure 5, lane 2), suggesting a down-regulation of the gene in the corneal epithelial cells following UVB exposure.The corneal epithelial cells were grown to confluency and irradiated for 20 minutes. Rogers et al. [29] have shown that these conditions have been effective in corneal epithelial cells. The cells were incubated for 2 or 4 hours post-treatment after which the cells were harvested to allow cellular recovery after UV irradiation. Some cells had undergone apoptosis following UVB exposure. Rogers et al.[29] have performed tunnel assays to study the induction of cell death due to broadband UVA and UVB to the cornea epithelial cells; furthermore, they have shown that UVA and UVB, at low dose, induced cell death in corneal epithelial cells.In this study, we used the FDD system to identify differentially expressed genes induced by UVB irradiation in rabbit corneal epithelial cells. DD has proven to be a powerful technique for the detection and isolation of differentially expressed genes. The DD technique provides a novel method for identifying those mRNAs that are induced or repressed at the gene level and is applicable in a variety of biological systems [30–32].The DD has been used as an alternative to the conventional differential or subtraction hybridization techniques for detecting differences in gene expression between closely related populations of eukaryotic cells. The major obstacle of differential display is not the technique itself however it is the post-differential display issue of discriminating between false positives and the truly differentially expressed mRNA. Therefore, to overcome this obstacle, and to achieve a reduced number of false positives, a variety of modified primers and altered cycling conditions have been implemented [32]. By generating reproducible cDNA expression data, it is possible to compare gene expression in two or more cell types, or developmental stages or tissues associated with diseases, and a technique to isolate unknown differentially expressed genes. For example, this method has been quite successful in identifying differentially expressed genes in normal versus tumor-derived human mammary epithelial cells [33], isolating the gene D2-2 that was over expressed in glioblastoma multiforne tissue as compared to normal brain tissue [34], as well as identifying and characterizing differentially expressed genes in various stages of prostate cancer development [35] and isolating light activated genes differentially expressed in Coprinus congregatus[36]. The differential display technique allows side-by-side comparison of two different cell populations and therefore helps to identify known genes as well as unidentified new genes. FDD methodology has many advantages in that it is based on PCR and DNA sequencing gel electrophoresis and is more sensitive and reproducible in screening new genes. Importantly, this technology has been revolutionized by the use of fluorescent detection of PCR products instead of radioactivity. The use of fluorescent dyes is significantly comparable to gamma [32P] isotopic labelling [37]. The goals of this study were to identify differentially expressed genes in response to ultraviolet exposure in corneal epithelial cells. In this study, the FDD was used to identify and isolate differentially expressed genes that are associated with UVB response in corneal epithelial cells.Total RNA isolated from untreated or UVB irradiated corneal epithelial cells was used as a template to synthesize cDNA. Three, one-base-anchored oligo-dT primers, labelled with 5′ rhodamine served as the primers. The use of denaturing sequencing gels showed several distinct bands from normal and UVB irradiated cells. After the gel was electrophoresed and scanned, eleven differentially expressed bands were selected and one was used as a probe for gene expression of differential display. It was demonstrated by Northern blot analysis that the gene (JS6 as a probe) was differentially expressed found in untreated and UVB irradiated cornea epithelial cells. In this case some genes were either turned on or off to UVB exposure (Figures 2–6). Gene regulation is manifested after UV radiation. This manifestation is seen as a down regulation of the gene of interest and is associated with a 1.2kb down regulated gene (Figure 5). Furthermore, this gene may be associated with cataract development involved in damaging the eye tissue.The findings of this study indicate that the fluorescent differential display-reverse transcription-polymerase chain (FDD-RTPCR) reaction is a useful method for identifying genes that are differentially expressed in response to corneal epithelial cells exposed to UVB irradiation. The literature has not elucidated which genes are activated or deactivated in association with cataract development due to UVB irradiation, which is the impetus for further ongoing studies in our laboratory. The genes differentially expressed in the untreated and UVB irradiated corneal epithelial cells will be used to study the mechanisms of action associated with cataract development. Therefore, we will address the identification through sequence analysis of genes activated or deactivated by UVB irradiation that may reveal molecular mechanisms underlying UVB exposure to cataract development. In addition, our laboratory also plans to use other cDNAs as probes to determine if there are other genes differentially expressed. Further, we would like to use the DD techniques on varying times of UVB exposure to the corneal epithelial cells and to determine if these genes are down-regulated or up-regulated, as well as exposure of cells to UVA. Therefore, the results of these studies should lead to a better understanding in the prevention of UVB-induced lens opacity (cataracts) with subsequent DNA damage to the cornea.Morphology of Corneal Epithelial Cells by Phase Contrast Microscopy. Morphologic changes in corneal epithelial cells exposed to UVB by phase contrast microscopy (40X magnification).A. Untreated confluent corneal epithelial cells (no exposure to UVB); a few cells were sensitive to apoptotic death.B. Morphologic changes in corneal epithelial cells exposed to UVB (0.6J/cm2) for 20 minutes; four apoptotic cells were visualized.C. Morphologic changes in corneal epithelial cells exposed to UVB (0.6J/cm2) for 45 minutes; twenty-six apoptotic cells were visualized.Fluorescent Differential Display. A typical fluorescent image of differentially expressed cDNAs using a 6% denaturing polyacrylamide electrophoresis. Differentially expressed bands obtained from untreated or UVB irradiated cornea epithelial cells are marked by arrows.Re-amplification of differentially expressed genes. 1.5% agarose gel of the eleven differentially expressed bands chosen for re-amplification. Lane 1 is 100 bp ladder (Gibco-BRL) and lanes 2–12 are the selected differentially expressed bands (JS1–JS11).Cloning of reamplified PCR products. Four colonies for each differentially selected band reamplified were checked for inserts by colony PCR. 100 bp ladder (NEB) was used as a size standard in lanes 1A, 1B and 1C for gels. JS represents selected differentially bands from normal and UVB irradiated corneal epithelial cells. Gel A: Lanes 2–5 (Inserts from JS #1; Lanes 6–9 (Inserts from JS #2); Lanes 10–13 (Inserts from JS #3); Lanes 14–17 (Inserts from JS # 4). Gel B: Lanes 2–5 (Inserts from JS #5; Lanes 6–9 (Inserts from JS #6); Lanes 10–13 (Inserts from JS # 7); Lanes 14–17 (Inserts from JS #8). Gel C: Lanes 2–5 (Inserts from JS #9; Lanes 6–9 (Inserts from JS #10); Lanes 10–13 (Inserts from JS #11).Down Regulation of Corneal Epithelial Cells after UVB Exposure. The differentially band, JS#6, was used as a probe; a 1.2 kb band was detected in the untreated (UT) cells (lane 1), whereas no band was detected in the UVB irradiated corneal epithelial cells (lane 2).This publication was made possible by the National Institutes of Health (NIH) Grant Number G12 RR13459 from the National Center for Research Resources. The project described used instrumentation found in the Molecular and Cellular Biology Core Laboratory and Analytical Core Laboratory. These laboratories were also supported by the NIH Grant Number G12 RR130459. Special thanks to Jeff Fisher for his technical help with the Fluorescent Differential Display System at GenHunter Corporation. Special thanks to Kathy Williams for assistance in the Northern blot analysis. We want to thank Mrs. Lynette Ekunwe, MCB Core Laboratory technician, for all of her tireless efforts. The authors also would like to thank the STARGE student, Tanya Robinson for her hard work. Special thanks to the External and Internal Advisory Board Committees for their advice and support during our research endeavours. We also extend our appreciation to Mr. Reginald Woodard in the Department of Computer Science within the College of Science, Engineering and Technology (Jackson State University) for his technical help.
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+ Over the past several years, a great deal of interest has been focused on the harmful effects of ultraviolet (UV) radiation to human skin. UV light has been implicated in aging, sunburn and skin cancer. Few studies, however, have been done to determine the effects that UV light, in conjunction with other environmental contaminants, may have on human skin. Polycyclic Aromatic Hydrocarbons (PAHs) are a class of compounds that have been reported to be toxic, mutagenic and carcinogenic to many eukaryotic organisms. UV light is also known to increase the toxicity of PAHs through photo-activation and photo-modification. The purpose of this study was to assess the effects of UV-A irradiated pyrene (Pyr), 1-aminopyrene (1-AP) and 1-hydroxypyrene (1-HP) on human keratinocytes, the skin primary site of UV irradiated PAH exposure. Our findings indicate that simultaneous treatment of human keratinocyte cell line, HaCaT, with 1.0μg/ml pyrene, 1-AP or 1-HP and 3.9 J/cm2/min UV-A light resulted in significant inhibition of cell proliferation. Approximately 100% of the cells died in the case of UV-A irradiated 1-AP and 1-HP. In the case of UV-A irradiated pyrene, more than 70% of the cells died, indicating that UV-A is able to transform these PAHs into more harmful intermediates.Over the past several years, a great deal of interest has been focused on the harmful effects of ultraviolet (UV) radiation to human skin. UV light has been implicated in aging, sunburn and skin cancer. Few studies, however, have been done to determine the effects that UV light, in conjunction with other environmental contaminants, may have on human skin. The ubiquitous environmental contaminants known as Polycyclic Aromatic Hydrocarbons (PAHs) are a class of compounds that have been reported to be toxic, mutagenic and carcinogenic to many eukaryotic organisms [1–3]. UV light is also known to increase the toxicity of PAHs through photo-activation and photo-modification [4, 5]. The purpose of this study was to assess the effects of UV-A irradiated pyrene (Pyr), 1-aminopyrene (1-AP) and 1-hydroxypyrene (1-HP) on human keratinocytes, which are the primary site of UV-A irradiated PAH exposure to humans.Dulbecco’s Modified Eagle Medium (DMEM) 1X with high glucose, L- glutamine, sodium pyruvate, pyridoxine hydrochloride, Fetal Bovine Serum (FBS) and 0.25% Trypsin were purchased from Gibco Invitrogen, Corporation, USA while Penicillin/Streptomycin (P/S) antibiotic 10,000 I.U./mL was purchased from Mediatech, Inc. USA. Pyrene, 1-aminopyrene, 1-hydroxyprene and benzo[a]pyrene, HPLC grade N, and N-Dimethylformamide (DMF) were purchased from Sigma-Aldrich Chemical Co., (St. Louis, Missouri).Dulbecco’s Modified Eagle Medium (DMEM) supplemented with 10% Fetal Bovine Serum (FBS) and 1% Penicillin/Streptomycin (P/S) antibiotic served as the complete growth medium (CGM). DMEM plus 1% P/S is the serum free medium. Stock solutions (1.0 mg/mL) of pyrene, 1-aminopyrene, 1-hydroxyprene and benzo[a]pyrene were prepared in HPLC grade N, N-Dimethylformamide (DMF).Spontaneously immortalized human keratinocyte cell line, HaCaT, was the cell model used in the cytotoxicity studies. They were a generous gift from Dr. Norbert Fusenig of the Division of Differentiation & Carcinogenesis of the German Cancer Research Center (DKFZ). HaCaT cells were prepared from the surgical excision of full thickness adult human skin from the distant periphery of a melanoma [6].Initially, a vial of frozen HaCaT cells containing 106cells/mL was rapidly thawed in room temperature water. The entire 1.0mL cell suspension was plated in 4.0 mL of complete growth medium in a T-25 canted-neck culture flask. The cell culture was allowed to grow at 37°C in a humidified (5% CO2, 95% air) incubator until cells were completely confluent, usually 7–9 days, with change of growth medium every 48 hours. Cells were trypsinized, counted using a hemocytometer, and passaged at 1:10 split ratio.DNA synthesis is used as an indirect measure of cell proliferation. DNA synthesis by HaCaT cells is assessed with [3H]thymidine incorporation assay. Briefly the assay involves growing cells to sub-confluence (≈60%), synchronizing cells by overnight serum starvation, followed by labeling with [3H]thymidine at 1μCi/mL for 4–6 hours. Cells are then fixed in 10% trichloroacetic acid (TCA) and solubilized with 2.0mL/well of 0.5M NaOH solution. One milliliter of the solubilized cells in 5.0mL scintillation cocktail is counted in Packard Tri-Carb TR 2700 Liquid Scintillation Analyzer.HaCaT cells were plated in Primaria™ 6-well plates at 105 cells/mL in CGM and allowed to grow at 37°C in a humidified (5% CO2, 95% air) incubator until sub-confluent (≈60% confluent). HaCaT cells were serum starved for 24 hours and then exposed to only UV-A light (3.9 J/cm2/min) for 20, 60, 80, 100, 120, 140, 160, 180 and 200 minutes at room temperature. Except for a thin coating of medium to prevent cells from drying up, all medium was withdrawn from the cells during UV irradiation. Fresh medium was added after UV-irradiation and cells were incubated for 18hours before being labeled with 3H[thymidine] at 1μCi/mL for 4–6 hours. One milliliter of the solubilized cells in 5.0 mL scintillation cocktail is counted in Hewlett Packard Tri-Carb 2700 TR Liquid Scintillation Analyzer. Cells treated with CGM and serum free DMEM served as the positive and negative controls, respectively. Each treatment was done in triplicates.HaCaT cells were assessed for survival after exposure to various concentrations of pyrene, 1-aminopyrene and 1-hydroxypyrene. Cells were plated in Primaria™ 6-well plates at 105cells/2ml in CGM and allowed to grow until sub-confluent. Cells were serum starved for 24 hours and then treated with pyrene, 1-AP or 1-HP at 0.01, 0.1, 1.0 and 10.0μg/ml for 18 hours before being labeled with 3H[thymidine] at 1μCi/mL for 4–6 hours. Cells treated with CGM and serum free DMEM served as the positive and negative controls, respectively. Each treatment was done in triplicates.HaCaT cells were assessed for survival after exposure to UV-A irradiated pyrene, 1-HP and 1-AP. Cells were plated in Primaria™ 6-well plates at 105 cells/2ml in CGM and allowed to grow until sub-confluent. Cells were serum starved for 24 hours and then simultaneously exposed to pyrene, 1-AP or 1-HP (1.0μg/ml) and UV-A light (3.9J/cm2/min) for 60 minutes and then incubated at 37°C for 18 hours prior to treatment with 3H[thymidine] at 1μCi/mL for 4–6 hours. Cells treated with CGM and serum free DMEM served as the positive and negative controls, respectively. Each treatment was done in triplicates.The results show that HaCaT cells responded to UV-A light in a biphasic manner (Figure 1). The growth of HaCaT cells exposed for 120 and 140 minutes was significantly different from the unexposed control. It has been established that UV-A light causes damage to cellular DNA, resulting in cell death [7–8]. The present study, however, did not show significant cell death after UV-A exposure. It is reasonable to say, therefore, that the UV-A exposure periods at an energy of 3.9 J/cm2/min were not adequate enough to stop DNA synthesis. This could be because the DNA was not damaged in a way to cause breaks in the DNA, although significant mutations could have occurred.PAHs are also known to be toxic to human cells, causing lung cancer when inhaled and skin cancer when exposed by skin contact [9]. Results obtained in this set of experiment suggest a biphasic response to PAH exposure. At low concentrations, pyrene caused stimulation of HaCaT proliferation while 1-HP and 1-AP inhibit cell proliferation; and at 10μg/mL, pyrene, 1-AP and 1-HP significantly inhibit HaCaT proliferation (Figures 2–4). It is an established fact that substituted PAH derivatives are more soluble than their parent compounds, and are therefore more accessible and toxic to the cells [10]. This could explain why, at low concentrations, the substituted PAHs were more toxic to the HaCaT cells than the parent compound pyrene.Other reports have also demonstrated PAH toxicity to eukaryotic cells. Comet assay profiles of blood cells collected from Ctenomys torquatus (rodent) captured from a coal strip mine, an area highly contaminated with PAHs, verify considerable DNA damage when compared to blood cells from the same type of rodent captured from coal free sites [11].To assess the response of human keratinocytes to UV-A irradiated PAHs, HaCaT cells were exposed to UV-A irradiated pyrene and two of its substituted forms, 1-AP and 1-HP. The exposure time (60 minutes) and PAH concentration (1.0μg/ml) were selected because in the single factor experiments, proliferation of HaCaT cells in either case was not significantly different from that of the unexposed control. In this group of experiments our findings suggest that pyrene, 1-AP and 1-HP are considerably more toxic to HaCaT cells when irradiated with UV-A light. Cell growth was almost completely prevented in the case of UV-A irradiated 1-HP and 1-AP, while in the case of UV-irradiated pyrene, cell growth was inhibited by more than 70% (Figure 5).Our findings indicate that simultaneous treatment of human keratinocyte cell line, HaCaT, with 1.0μg/ml pyrene, 1-AP or 1-HP and 3.9J/cm2/min UV-A light resulted in significant inhibition of cell proliferation. Approximately 100% of the cells died in the case of UV-A irradiated 1-AP and 1-HP. In the case of UV-A irradiated pyrene, more than 70% of the cells died, indicating that UV-A is able to transform these PAHs into more harmful intermediates. These results are consistent with other reports that discuss cell death caused by the interaction between UV light and PAHs. The photo-modification of PAHs results in the formation of diones and quinones, which can bind to cellular macromolecules and cause problems in signalling and proliferative pathways, which ultimately lead to cell death. With the binding of large macromolecules to DNA, the cell becomes severely challenged to replicate such DNA molecules with these large adduct. The DNA damage repair system could be ineffective in repairing such DNA damage. The end result is cell death. Cytotoxic effect, such as lipid peroxidation of the cell membrane leads to compromise of cell membrane integrity. Once more, the end point is cell death, either by necrosis or apoptosis. It has been found that UV-A irradiated benzo[a]pyrene causes cellular reactions to occur that heighten DNA damage [10]. Such exposure to human epidermoid carcinoma cells and human keratinocytes results in a 5-fold increase in the production of H2O2, which is known to cause significant DNA damage [12–13]. The classic Haber-Weiss-Fenton reaction involving H2O2 in the presence of Fe+2 leads to double strand-break of DNA strands. Either could be a possible reason for the cell death observed in this study. These results are significant because in the South and South-eastern US, exposure of the population to UV-A is considerably higher than in the North. Exposure to combination of UV-A and PAHs from petrochemicals, creosote use in wood treatment facilities and tractor exhaust gases in the cotton fields of the Delta, makes this a vexing environmental health concern. The results obtained in this study do not only increase our understanding of the problem but also provide us the opportunity to seek appropriate solutions for the problem.From the results in this study, the following conclusions can be made: (1) UV-A light inhibits growth and proliferation of HaCaT cells (2) UV-A irradiated PAHs are highly toxic to HaCaT cells (3) UV-A irradiated substituted pyrenes are more toxic than their parent compound. This agrees with what is already known about UV-A light causing damage to cellular DNA, resulting in cell death [7–8]. The toxic effects of UV-A irradiated PAHs to human cells, is established. Drastically reduced exposure of humans to this combination of UV-A and PAHs is necessary to avoid lung and skin cancer induced by exposure to UV-A irradiated PAHs [9].Effects of UV-A light on HaCaT proliferation. HaCaT cells were grown in 35mm plates to subconfluent in CGM. Cells were serum starved for 24 hours and then exposed to UV-A (3.9J/cm2/min) for 20, 40, 60, 80, 100, 120, 140, 160, 180, and 200 minute intervals. Thymidine incorporation assay was then performed as described in materials and methods. Results represent the mean +/− SD values of experiment performed in triplicate. [* indicates that treatment mean is significantly different from negative control according to Dunett test (p<0.05)].Effects of pyrene on HaCaT proliferation. HaCaT cells were grown in 35mm plates to subconfluent in CGM. Cells were serum starved for 24 hours and then treated with pyrene at 0.01, 0.1, 1.0 and 10.0μg/ml concentrations for 18 hours. Thymidine incorporation assay was then performed as described in materials and methods. Results represent the mean +/− SD values of experiment performed in triplicate. [* indicates that treatment mean is significantly different from negative control according to Dunett test (p<0.05)].Effects of 1-hydroxypyrene on HaCaT proliferation. HaCaT cells were grown in 35mm plates to subconfluent in CGM. Cells were serum starved for 24 hours and then treated with pyrene at 0.01, 0.1, 1.0 and 10.0μg/ml concentrations for 18 hours. Thymidine incorporation assay was then performed as described in materials and methods. Results represent the mean +/− SD values of experiment performed in triplicate. [* indicates that treatment mean is significantly different from negative control according to Dunett test (p<0.05)].Effects of 1-aminopyrene on HaCaT proliferation. HaCaT cells were grown in 35mm plates to subconfluent in CGM. Cells were serum starved for 24 hours and then treated with pyrene at 0.01, 0.1, 1.0 and 10.0μg/ml concentrations for 18 hours. Thymidine incorporation assay was then performed as described in materials and methods. Results represent the mean +/− SD values of experiment performed in triplicate. [* indicates that treatment mean is significantly different from negative control according to Dunett test (p<0.05)].Effects of UV-A irradiated pyrene, 1-aminopyrene and 1-hydroxypyrene on HaCaT proliferation. HaCaT cells were grown in 35mm plates to sub-confluent in CGM. Cells were serum starved for 24 hours and then treated with each PAH at 1.0μg/ml and immediately irradiated with UV-A (3.9J/cm2/min). PAH treatment remained on cells for a total of 18 hours, including time during UV-A exposure. Thymidine incorporation assay was then performed as described in materials and methods. Results represent the mean +/− SD values of experiment performed in triplicate. [* indicates that treatment mean is significantly different from negative control according to Dunett test (p<0.05)].This research was supported in part by a grant from the Army Research Office (Grant# DAAD 19-01-1-0733), awarded to Jackson State University, and in part by a grant from the National Institutes of Health Research Centers in Minority Institutions, NIH-RCMI (Grant #1G12RR13459) awarded to Jackson State University. Part of this work was used by Rochelle Hunter to satisfy the Masters Degree thesis requirement at Jackson State University.
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+ Living in an environment that has been altered considerably by anthropogenic activities, fish are often exposed to a multitude of stressors including heavy metals. Copper ions are quite toxic to fish when concentrations are increased in environmental exposures often resulting in physiological, histological, biochemical and enzymatic alterations in fish, which have a great potential to serve as biomarkers. Esomus danricus was chosen as model in the present study and the metabolic rate, gill morphology, total glycogen, total protein, superoxide dismutase and catalase were critically evaluated. The 96h LC50 value was found to be 5.5mg/L (Cu as 1.402mg/L). Fish groups were separately exposed to lethal (5.5mg/L) and sub lethal concentrations (0.55 mg/L) of copper sulphate over a period of 96h to examine the subtle effects caused at various functional levels. Controls were also maintained simultaneously. Significant decrease in the metabolic rate (p<0.001) of the fish was observed in both the concentrations studied. Studies employing Automated Video Tracking System revealed gross changes in the architecture of gill morphology like loss, fusion, clubbing of secondary gill lamellae, and detachment of gill rakers following softening of gill shaft in fish under lethal exposures indicating reduced respiratory surface area. Biochemical profiles like total glycogen and total protein in gills and muscle of fish exposed to 5.5 mg/L showed appreciable decrease (p<0.05 to 0.001) from control. Significant inhibition of superoxide dismutase (60.83%), catalase (71.57%) from control was observed in fish exposed to 5.5 mg/L at the end of 96h exposure only. Interestingly, in fish exposed to 0.55 mg/L enzyme activity is not affected except for catalase. Toxic responses evaluated at various functional levels are more pronounced in fish exposed to 5.5mg/L and these can serve as potential biomarkers for rapid assessment of acute copper toxicity in environmental biomonitoring.Environmental pollutants such as metals, pesticides and other organics pose serious risks to many aquatic organisms including fish [1]. Heavy metals constitute a core group of aquatic pollutants [2]. The crux of the problem lies in the fact that these metals not only accumulate in waters and sediments but also concentrate in the tissues of the fish causing alterations at various functional levels of the organism. Copper is one of 26 essential trace elements occurring naturally in plant and animal tissues and its availability is influenced by physico-chemical, hydrodynamic and biological factors. It makes its way into the receiving waters by extensive use in agriculture apart from usage in various industries like textile, tanneries, paints, battery, laundry, photography, copper ware and piping for water distribution systems. Copper ions are quite toxic to fish at various functional levels when environmental concentrations are increased [3].Acute toxicity of a xenobiotic often can be very helpful in predicting and preventing acute damage to aquatic life in receiving waters as well as in regulating toxic waste discharges [4]. Gill architecture of fish is an important index in understanding the effect of heavy metals on the structural integrity of vital organs at cellular level. Respiratory distress is one of the earliest manifestations of heavy metal toxicity [5]. The toxic effects may result from the bioconcentration of metals and their consequent binding with biologically active constituents of the body such as lipids, amino acids, enzymes and proteins [6]. Despite the essential role of copper in a number of vital biochemical processes including cellular respiration, the metal has the potential to exert adverse toxicological effects [7]. The metal is known to impair glycolysis in freshwater fish [8]. Copper also acts as an enzyme activator as it is incorporated into enzymes like cytochrome oxidase, superoxide dismutase [9]. The activity of these enzymes is dependent on the adequate supply of metal but excess copper can also inhibit the activity of enzymes. Copper due to its redox potential generates free radicals leading to oxidative stress, which damages cellular components like lipids by causing lipid peroxidation, DNA and proteins if not quenched by the antioxidants. The occurrence of such alterations in biochemical profiles has a great potential to serve as ‘biomarkers’.Fish has a great potential to serve as sensitive indicators, signalling exposure and understanding the toxic mechanisms of stressors in aquatic ecosystems. Less frequently studied are the implications of interrelated toxic effects on the survival, histology, physiology, biochemical constituents and behaviour caused by aquatic pollutants for fish populations. To establish these effects a systematic evaluation was carried out on the above parameters in the present study using Esomus danricus exposed to lethal and sub-lethal concentrations of copper so that these responses can be employed as potential indicators of its toxicity in freshwaters.Esomus danricus were collected from Undasagar fish farm (Long 17°27′-78°27′15″ and Lat 17°18′-17°18′15″) located nearby Hyderabad for toxicity tests. The fishes were transported from the farm in oxygenated polythene bags to the laboratory and immediately transferred into glass aquaria of 100 l capacity containing well-aerated unchlorinated ground water. Fish were screened for pathological signs, if any and acclimated for a fortnight in the glass aquaria of 50 l capacity before the experiments. Rice bran was fed to the fish ad libitum during the acclimation period. Healthy fish that showed active movements were only considered for the experimentation.Prior to the acclimation of fish, the physico-chemical characteristics of the dilution water were analyzed adopting standard protocols [4] owing to their role in determining the toxic potential of heavy metals. Renewal bioassay was adopted in the present study due to its advantages over other bioassay techniques. This method has the advantage of replacing the toxicant solution afresh every 24h so that metabolic waste (ammonia) which itself is highly toxic can be removed, sustains copper bioavailability besides replenishment of dissolved oxygen. Desired concentrations of copper were derived by adding aliquots of 1% CuSo4.5H2O stock solution (prepared in double distilled water) and the dilution water in the test chambers was renewed and fresh solution of same concentration was added every 24h. Pilot experiments were conducted to choose concentrations that resulted in mortality of the fish within the range of 5% to 95%. Fish measuring from 4.6cm with a weight of 1.45g are used. The fish were starved 24h prior to and also during the course of the experiment. Glass aquaria of 50 l capacity were used as test chambers and 30 fish were tested in each concentration. The loading of fish in the test chambers was according to the recommendations [10]. No distinction was made between sexes.Definitive tests were later conducted using four concentrations of copper sulphate i.e. 2.5mg/L, 5.0mg/L, 7.5mg/L and 10.0mg/L, which resulted in the mortality of the fish within the range of 5% to 95%. Thirty fish were exposed to each concentration separately. Controls without toxicant were also run simultaneously. Behavioural manifestations and condition of the fishes were noted every 24h up to 96h. Between the experiments, the chambers were carefully washed to eliminate residual metal adsorption to walls. The fish that failed to respond even to strong tactile stimuli were considered dead and removed immediately. The mortalility of fish was recorded for each concentration of the toxicant and the data was used to find the median lethal concentration (LC50) adopting probit analysis and the corresponding results were generated with computerized program [11]. The regression equations were calculated by the method of least squares and 96h LC50 value was derived from the equation.Esomus danricus (length 4.6cm and weight 1.45gm) were exposed to 96h LC50 concentration of copper sulfate (5.5 mg/L) and sub lethal concentration of copper sulfate (0.55mg/L) for 96h. At the end of exposure period the fishes that survived were sacrificed and dissected carefully to isolate gill III tissue. The gills of control and exposed fishes were placed in saline and rinsed for 3 to 4 times thoroughly. They were placed on a glass slide and observed under a microscope (Polyvar, Reichert-Jung light microscope) attached to Ethovision-version 2.3 (Noldus Information Technology, Netherlands) through a CCD camera (Sony CCD IRIS, Model No: SSC-M370CE). Instantly, the digital photographs were stored in the computer system. The magnification of the snaps was calibrated with the aid of ocular and stage micrometers (ERMA, TOKYO, JAPAN).Fish (n=30) were exposed both to 5.5mg/L; 0.55mg/L and a range of concentrations of copper sulphate (1.5mg/L, 2.5mg/L, 5.0mg/L, 7.5mg/L and 10mg/L) that were used to determine the median lethal concentration and examined for any change in the metabolic rate and ventilation frequency over a period of 96h. Controls without toxicant were also run simultaneously. The oxygen consumption of the fish was estimated by using a DO probe (Century, CD 501, India). In this method, oxygen probe containing an electrode is connected to the electronic meter, which displays the dissolved oxygen in the sample water at a specific temperature (25°C). Respiratory measurements were carried out as per the method [12]. A BOD bottle (300 ml) chamber was used as respiratory chamber, which was filled with tap water. The initial DO of the water was measured by introducing the probe into the respiratory chamber without any air-bubbles. Fish were introduced into the bottle taking care to avoid air bubbles. The bottles were stoppered airtight and kept aside, undisturbed for an hour. At the end of one hour, the final DO of the water was measured using the DO probe and the fish were replaced into their corresponding aquaria, after each interval of exposure. This process was repeated for 5.5mg/L, 0.55mg/L and a range of concentrations viz., 1.5mg/L, 2.5mg/L, 5.0mg/L, 7.5mg/L and 10.0mg/L. The experiments were repeated thrice and the arithmetic mean was considered. After the experiment, the wet weights of the fish were taken to calculate the metabolic rate. The amount of oxygen consumed by the fish in 1 h was expressed as mg O2/h and the metabolic rate as mg O2/g/h. The respiratory measurements were made in diffused daylight and the time of experiment was kept constant (11.00am to 3.00pm) to avoid the effect of time of day on the respiration of the fish.The numbers of ventilation movements (opercular beats) were counted per 1 minute in the BOD bottle during the respiratory measurements. Performing the experiment in the BOD bottle aided in simultaneous determination of both the oxygen consumed and the ventilation rate of that particular fish.Thirty fish were exposed both to 5.5mg/L and 0.55mg/L concentrations of copper and examined for any change in the biochemical constituents over a period of 96h. Controls without the toxicant were also run simultaneously. The fishes were sacrificed at the end of every 24h, blotted dry and weighed to the nearest mg. They were later dissected to isolate the gill and muscle tissues. The tissues were dried for 24h, at 50 °C to eliminate water. The dried tissues were weighed to the nearest mg. The tissues of control were also processed similarly for biochemical analysis. The important biochemical biomarkers viz., total glycogen and total protein were analyzed as per the universally accepted protocols [13–14], respectively. The data are subjected to Unpaired Student’s ‘t’ test [15]. P<0.05 was selected as the criterion for statistical significance.In a separate set of experiments, fish (n=30) were exposed to 5.5 mg/L and 0.55 mg/L for a period of 96h. At the end of 24 and 96h of exposure viscera was isolated to carry out the enzyme assays. The viscera are immediately homogenized (10% w/v) in 0.1 Mphosphate buffer (Ph 7.5) using Potter-Elvehjam Homogenizer fitted with a Teflon pestle. The homogenates were centrifuged at 10,000-x g for 10 minutes. The resultant supernatant of viscera was used as the enzyme source for the estimation of enzymes (antioxidants). All the enzyme preparations were carried out at 4°C.SOD activity in the viscera was estimated at the end of 24 and 96h of exposure [16]. A typical run for all the in vivo experiments in 96 well plates consisted of 220μl of Pyrogallol (C6H6O2) and 30μl of enzyme sample for each well. The colour that was developed was recorded continuously for 10 minutes in kinetic mode at 420 nm using a molecular device UV-spectrophotometer supported by soft max-pro-3 software. The percent change in SOD activity of exposed organisms was calculated based on the control values. SOD activity was calculated as O.D/min/mg protein.Catalase activity in the viscera was estimated at the end of 24 and 96h of exposure [17]. A typical run for all the in vivo experiments in 96 well plates consisted of 1.0ml Hydrogen peroxide (H2O2 0.059M) in phosphate buffer, 1.95ml of distilled water and 50μl of enzyme for each well (extracted from control and Lc50 exposed organisms). The decrease in absorbance was measured immediately at 240nm using kinetic mode against distilled water blank with 10 seconds intervals for 2 minutes in a molecular device UV-spectrophotometer supported by soft max-pro-3 software. The percent change in Catalase activity of exposed organisms was calculated based on the control values. CAT activity was calculated as O.D/min/mg protein.The concentrations of CuSo4.5H2O tested in the present study were 1.5mg/L (Cu as 0.382mg/L), 2.5mg/L (Cu as 0.637mg/L, 5.0 mg/L (Cu as 1. 275mg/L), 7.5mg/L (Cu as 1.912 mg/L) and 10.0mg/L (Cu as 2.549mg/L). The mortality in different concentrations ranged from 6.66% (2.5mg/l) to 93.33% (10.0 mg/l) and is dependent on both time and concentration.Table 1 shows the physico-chemical characteristics of water used for acclimation, control and experiment. The percentage mortality of Esomus danricus over 96h exposure at different concentrations of CuSo4.5H2O and the regression equation of the expected Probit (Y) and log concentration (X), 96h LC50 value, 95% confidence limits are presented (Table 2). The 96h LC50 value for CuSo4.5H2O is found to be 5.5mg/L (Cu as 1.402mg/L) and the 95% fiducial limits are 4.781mg/L (Cu as 1.219 mg/L) to 6.229mg/L (Cu as 1.588 mg/L). The sub-lethal concentration 0.55mg/L (Cu as 0.1402mg/L) was derived as 1/10th of LC50 concentration.The reaction and survival of aquatic animals depend on not only the biological state of the animals and physico-chemical characteristics of water but also on kind, toxicity, type and time of exposure to the toxicant [18]. In the present study, the mortality increased with an increase in the concentration of the toxicant and also the duration of exposure. This is in agreement with earlier studies explaining the relationship between exposure duration, tissue residues, growth and mortality in rainbow trout (Oncorhyncus mykiss) juveniles’ sub-chronically exposed to copper [19]. Behavioural manifestations of acute toxicity like copious mucus secretion, loss of scales, grouping, loss of equilibrium was observed in Esomus danricus exposed to copper. Loss of swimming performance was also observed in brown trout, Salmo trutta exposed to sub lethal concentration of copper [20]. The effects of copper and zinc salts on Cyprinus carpio and Ctenopharyngodon idellus showed that the body and the gills of dead fish seemed to be covered by a veil-like film which looked like coagulated mucus and which was formed by the heavy-metal ions reacting with some constituents of the mucus secreted by the gill [21]. Heavy metal induced changes in physiology and survival of aquatic organisms is complicated because such changes differ from metal to metal, species to species and from one experimental condition to another which may also account for the differences in the 96h LC50[22–23]. The exact causes of death due to heavy metal poisoning are multiple and depend mainly on time - concentration combinations.Figure 1 shows the histological structure of the normal gill characterized by the presence of primary lamellae along with secondary lamellae, shaft and rakers confirming the general architecture of the tissue. Gills exposed to 5.5 mg/L of copper have shown that the metal affected the primary and secondary lamellae, rakers and shaft. Figure 2 illustrates the degeneration of secondary gill lamellae with loss of original shape due to the onset of necrosis. Fish gills exposed to 0.55 mg/L shown that copper exposure produced hyper secretion of mucous (Figure 3).The fish gill has very little protection other than the body cover – the operculum- and is susceptible to both physical and chemical damage. A common response of the gill to irritation from any irritant is hyperplasia due to the chemical or physical irritation as a form of protection. Secondary lamellae clumped together affecting gaseous exchange and respiration. No gaseous exchange can take place and the fish literally suffocates. Since gills are the respiratory and osmoregulatory organs of the fish, cellular damage induced by the metal in terms of atrophy, bulging, hyperplasia of interlamellar epithelia, separation of epithelial layer might have probably impaired the respiratory function of the gill by reducing surface area. A thick coat of mucus covering entire gill filaments and lamellae was observed after 96h exposure to LC50 concentration of copper sulfate and the direct deleterious effect of the toxicant in the form of necrosis and abnormalities of gill lamellae is evident when compared to the control. The observed epithelial necrosis and rupture are direct responses induced by the action of copper sulfate.Whether the changes in gill lamellae observed in the present study are secondary or primary effects of copper sulfate is difficult to establish, atrophy, bulging and desquamation of secondary gill lamellae can be due to the metal induced toxicity. A perusal of the available literature revealed that copper potentially damages the gill epithelium in a variety of fishes and it was indicated that the study of histology is a successful tool capable of revealing sensitively and selectively even the sub lethal effects of heavy metals on the environment and aquatic biota [24–28]. The changes in gill epithelia of Esomus danricus caused by copper may represent a defense response initially because these changes increase the distance across which copper must diffuse to reach the blood stream. They also increase the water-blood distance for oxygen diffusion whereas lamellar fusion reduces the respiratory area. Ironically, it was reported that the mucus layer on the gill surfaces creates a microenvironment, which may act as an ion trap, concentrating trace elements from the water [29]. Although Esomus has a reasonably large respiratory surface the changes observed in its gill tissue probably impaired branchial gas transfer, generating an internal hypoxia.Table 3 shows the effect of lethal and sub-lethal concentrations of copper sulphate on the metabolic rate of Esomus danricus. It is clear from the results that the metabolic rate decreased while the percentage decrease from control increased with an increase in the exposure period from 24 to 96h to copper sulphate. The results of the ‘t’ test show that the decrease in the metabolic rate is highly significant with p<0.01 and p<0.001 at the end of each exposure period.The present investigation revealed a significant depression in the metabolic rate of Esomus danricus exposed to CuSO4.5H2O with respect to time. With increase in time duration of the experiment, the metabolic rate in each concentration showed a decrease. At the end of 24h, 48h, 72h and 96h, the metabolic rate showed significant decreases with increase in duration of exposure. Decrease in oxygen consumption is attributed to the onset of hypoxia and gill damage while the ventilation frequency increased to flush out the toxicant from the body. Reduced oxygen consumption can be attributed to increased mucous secretion, a specific property of copper to precipitate proteins and increase mucous secretion. As a result of gill damage, there is onset of acute hypoxia under metallic stress. The drop in the oxygen consumption also appears to be a protective mechanism to ensure that there is low intake of CuSO4. Architectural changes in the gills reduced their surface area, making it harder for the fish to extract oxygen from the water. Copper causes damage to gill tissues and induces fish mortality by the excessive cell proliferation and mucous production. Fish may asphyxiate due to excessive mucus [30]. Previous studies also reported that gills are vital respiratory and osmoregulatory organs and cellular damage induced by metals might impair the respiratory function of the fish by reducing the gill surface area [24–26, 31, 32].Fish exposed to 5.5 mg/L and 0.55 mg/L of CuSO4, respectively, showed a decrease in metabolic rate as a function of time. But the metabolic rate of sub-lethal exposed fish showed lesser percentage decreases from the control than compared to that of the fish exposed to lethal concentration. This phenomenon can be accounted due to the fact that the sub-lethal concentration is very low and the stress induced is less when compared to that of the lethal concentration.Table 4 shows the effect of lethal and sub-lethal concentrations of copper sulphate on the ventilation frequency of Esomus danricus in both the control and exposed fish at the end of 24h, 48h, 72h and 96h along with the percent increase from control. The ventilation rate and the percentage decrease from the control plotted against time is shown (Figure 4). The ventilation rate increased from control with an increase in the exposure period of 24h to 96h to fish exposed to lethal, sub-lethal and lower concentrations of 1.5 mg/L, 2.5mg/L and 5.0mg/L of CuSO4. Interestingly, in higher concentrations of 7.5mg/L and 10.0mg/L there was an initial increase in the ventilation frequency up to 48h and exhibited a significant drop at the end of 72 and 96h.As a result of gill damage, the oxygen consumption decreased which caused the fish more stress. In order to make more oxygen available from the water, the fish increase the rate at which they ventilate their gills [27–28]. When fish pump water across their gills at a faster rate, more oxygen is available for uptake. Therefore, a link between the ventilation rate and oxygen consumption was observed. Because of gill damage, there was a decrease in oxygen consumption with an increase in ventilation frequency to overcome the metal induced stress. The increase in ventilation frequency can also be attributed to physiological stress induced by the toxicant. Because CuSO4 induces mucous secretion, a layer of thick mucous reduced the uptake of oxygen by covering the gills and as a result, the ventilation rate increased. The decrease in the ventilation frequency after an initial increase in the higher concentrations (7.5 mg/L and 10 mg/L) can be attributed to the fact that the stress induced in the fish due to the toxicant at the end of 96h exposure was so high that the fish could not tolerate any longer and perished. Decreased ventilation frequency in higher concentrations was due to the failure of the homeostatic mechanism to cope up with the increasing metal load (Figure 5)The total glycogen and total protein contents expressed as mg per gram dry weight of the tissues, gills and muscle of Esomus danricus exposed to lethal and sub lethal concentrations of copper and control at the end of every 24h up to 96h are presented (Tables 5–8).Table 5 shows the total glycogen content in gills of the fish exposed to lethal and sub lethal concentrations of copper. The total glycogen content in the gills of the fish exposed to lethal (5.5mg/L) concentration was 4.57mg, 4.47mg, 4.33mg and 4.12 mg and sub-lethal (0.55mg/L) concentration was 4.78mg, 4.54mg, 4.53 mg and 4.27 mg, showing a decrease of 7.863%, 9.879%, 12.702% and 16.935% in gills of fish exposed to 5.5mg/l and 3.629%, 8.467%, 8.669% and 13.911% in gills of fish exposed to 0.55mg/L from control at the end of 24, 48, 72 and 96h, respectively. The‘t’ test values show that the decrease of glycogen content from control in gills of fish exposed to 5.5mg/L is significant at the end of 48h (p<0.05) 72h and 96h (p<0.01) exposure periods whereas the decrease of the same in gills of fish exposed to 0.55mg/L is significant only at the end of 96h (p<0.01).Decrease in total glycogen content was also observed in muscle of the fish exposed to lethal, sub lethal and other concentrations of copper. Table 6 shows that glycogen in muscle of fish exposed to 5.5 mg/L was 11.49 mg, 11.46 mg, 10.90 mg and 10.82 mg and that of the same in muscle of fish exposed to 0.55 mg/L was 11.70 mg, 11.59 mg, 11.54 mg and 11.38 mg at the end of every 24h exposure and up to 96h showing a decrease of 2.213%, 2.468%, 7.234% and 7.915% in 5.5 mg/l and 3.629%, 8.457%, 8.669% and 13.911% in 0.55 mg/L from control at the end of 24, 48, 72 and 96h, respectively. Statistical analysis show that the decrease in total glycogen content in muscle of fish exposed to 5.5 mg/L is significant (P<0.01) at the end of 72 h and 96 h while the decrease of the same in muscle of fish exposed to 0.55 mg/L is insignificant.The total protein content in gills of the fish exposed to lethal and sub lethal concentrations of copper are presented (Table 7). In gills of fish exposed to 5.5 mg/L, the total protein content was 13.75 mg, 12.30 mg, 11.76 mg and 11.50 mg and that of gills of fish exposed to 0.55 mg/L was 14.75 mg, 14.50 mg, 13.75 mg and 13.50 mg at the end of every 24h exposure period and up to 96h. A decrease of 7.594%, 17.338%, 20.967% and 22.715% in 5.5 mg/l and 0.874%, 2.554%, 7.594% and 9.274% in 0.5 mg/l was observed from control at the end of 24, 48, 72 and 96h, respectively. The ‘t’ test results show that the decrease in the gills of the fish exposed to 5.5 mg/L is significant (p<0.01) at the end of 48, 72 and 96h while this decrease is significant only at the end of 96h in gills of fish exposed to 0.55 mg/L.Table 8 shows the protein content in muscle of fish exposed to 5.5 mg/L and 0.55 mg/L of copper. Protein content in muscle of fish exposed to 5.5 mg/l was 22.8 mg, 22.46 mg, 20.84 mg and 20.12 mg and that of fish exposed to 0.55 mg/L was 23.3 mg, 22.95 mg, 21.50 mg and 21.39 mg at the end of 24, 48, 72 and 96h, respectively. A decrease of 2.979%, 4.426%, 11.319% and 14.383% in 5.5 mg/l and 0.8511%, 2.340%, 8.511% and 8.979% in 0.55 mg/L was observed at the end of 24, 48, 72 and 96h, respectively. The results of the ‘t’ test show that the decrease in total protein content in muscle of fish exposed to 5.5 mg/L is significant (p< 0.05) at the end of 96h and that of fish exposed to 0.55 mg/L is insignificant.It is evident that copper is highly toxic to the fish and the decrease in biochemical biomarkers viz., total glycogen and total protein content in gills and muscle of fish exposed 5.5 mg/L and 0.55 mg/L concentrations of copper demonstrated a linear and positive correlation with the concentration and duration of exposure indicating that the decrease in biochemical constituents is time and dose dependent.Total glycogen content of Esomus danricus was depleted due to copper toxicity. Similar trends were also observed in other investigations on copper toxicity to freshwater fish. Copper was found to impair glycolysis in Labeo rohita[8]. A depletion of glycogen and phosphocreatine in the white muscle was observed in brown trout Salmo trutta exposed to sub-lethal concentration of copper in soft acidic water [20]. Decrease in glycogen level in liver of Oreochromis niloticus was reported with an increase in copper concentration in water [33]. The levels of carbohydrates and glycogen in aquatic organisms reveal their involvement in the endogenous derivation of energy during stress. Rapid depletion of muscle and liver glycogen reserves in order to compensate the energy needs of fish under acute metallic stress was also reported [39, 40]. A consistent decrease in the tissue glycogen reserves observed in this study suggests impaired glycogenesis. Further, the decline in the glycogen content might be partly due to its utilization in the formation of glycoproteins and glycolipids, which are essential constituents of various cell and other membranes. Decline in the glycogen content of the tissues of Esomus danricus may be due to its enhanced utilization since glycogen forms the immediate source of energy to meet the energy demands under metallic stress. It might be also due to the prevalence of hypoxic or anoxic conditions, which normally enhances glycogen utilization. Our studies revealed that copper triggers the onset of hypoxia in Esomus danricus. The enhanced utilization of glycogen and its consequent depletion in the tissues, therefore suggests the initiation of anaerobic glycolytic pathway by increased glycogenolysis as has been suggested [37].There was also depletion in the total protein content in gills and muscle of Esomus danricus exposed to copper. Protein content can be taken as a biomarker of copper level. The decrease in the tissue proteins observed in the gills and muscle (Tables 5–8) could be partly due to their utilization in cell repair and tissue organization with the formation of lipoproteins, which are important cellular constituents, occurring in cell membranes and cell organelles present in cytoplasm. Direct utilization of proteins as immediate source of energy to meet the energy demands also cannot be ruled out [42]. The effects of metals may result from their binding with biological constituents of the body such as lipids, amino acids, enzymes and proteins [6]. The depletion in tissue proteins may be due to impaired or low rate of protein synthesis under metallic stress [43] or due to their utilization in the formation of mucoproteins, which are eliminated in the form of mucous. The depletion in protein may result in further modification of enzyme activity (stimulation or inhibition). Changes in protein content may also modify signal transfer in cells. Also attack on proteins can lead to the modification of amino acids, oxidation of sulphydryl groups, leading to conformational changes, cross linking, peptide bond cleavage as well as carbohydrate modification in glycoproteins. It can be substantiated that glycogen and protein contents under metallic stress in fish can be effectively employed as biomarkers for rapid assessment of heavy metal toxicity in bio monitoring of aquatic environments.Table 9 illustrates the activity of superoxide dismutase and catalase in the viscera of the fish exposed to lethal and a sub lethal concentration of CuSO4.5H2O, respectively. The results show that there was significant decrease (p<0.001) in the activity of superoxide dismutase in the viscera of the fish exposed to lethal and sub lethal concentrations of copper for 96h. There was significant (p<0.001) decrease in the Catalase activity in viscera of the fish exposed to lethal and sub-lethal concentrations of CuSO4.5H2O for both 24 and 96h. From the results it is evident that there is decrease in the antioxidant enzymes activity in the fish exposed to lethal and sub-lethal concentrations of CuSO4.5H2O.This decrease increased with duration of exposure and concentration of the metal. Appreciable decline in SOD activity in the viscera of silver fish, Esomus danricus was found under the impact of Cu toxicity when exposed to lethal and sub lethal concentrations for 96h. The decrease in the activity of the enzyme SOD could be due to its inhibition by excess production of ROS, decreased uptake in the diet, alteration in the gene expression or due to protein precipitation [45, 46].Catalase activity declined significantly in the viscera of silver fish, Esomus danricus exposed to lethal and sub lethal concentrations for both 24 and 96h. The decrease in the Catalase activity could be due to its inactivation by Superoxide radical or due to decrease in the rate of reaction as a result of excess production of H2O2. Earlier reports attributed the inhibition of catalase activity due to Superoxide radical and demonstrated the synergism between the SOD and Catalase [47]. A decrease in the activity of catalase is due to excess free radical production by Cu, which results in accumulation of O2−, and H2O2, which in addition to direct generation by Cu is also produced by dismutation of O2− by superoxide dismutase. The H2O2 produced reacts with O2− to form more reactive ROS like OH, which damages the cellular components like DNA, proteins and lipids. Our studies also revealed a significant decrease in protein profiles of the fish, which substantiates the findings of the present study. A consistent decrease in the antioxidant enzyme levels was observed in the study suggesting that there is impaired antioxidant defense mechanism due to excess generation of oxy radicals by Cu. Further, decline in antioxidant levels might also be due to overproduction of ROS which creates a net increase in the amount of oxygen free radicals present in the cell. Antioxidants like SOD act as primary preventive inhibitor by catalyzing the conversion of superoxide anion (O2−) to H2O2 and O2 and scavenge the O2 produced by rapid dismutation reaction. Catalase functions to rapidly dismutase H2O2 to water and oxygen [48]. Therefore any significant reduction in these antioxidant levels results in lipid peroxidation, as normal levels of antioxidants produced could not quench the excess free radicals generated. The findings of the present study clearly demonstrated that enzymatic biomarkers like SOD and CAT in fish can be effectively used for rapid assessment of copper toxicity in biomonitoring of aquatic environment.Copper is found to be highly toxic to Esomus danricus and induced gross alterations in the gill architecture, significant decrease in metabolic rate, biochemical constituents like glycogen and proteins and also key antioxidant enzymes superoxide dismutase and catalase suggesting that the metal exerts its effect at various functional levels of organization as a function of time with more pronounced changes occurring in the lethal exposures. These parameters could be effectively used as potential biomarkers of copper toxicity to the freshwater fish in the field of environmental biomonitoring.Control gill lamella of Esomus danricusAlterations in secondary gill lamellae exposed to 5.5mg/l of copperAlterations in gill of Esomus exposed to sub-lethal concentration (0.55mg/L) of copperEffect of lethal and sub-lethal concentrations of copper on the ventilation frequency in EsomusEffect of various concentrations of copper on Ventilation frequency in Esomus.The physico-chemical characteristics of water used for acclimation, control and experimentsLethal Concentration (LC50) of Copper to Esomus danricusEffect of Lethal and Sub-lethal concentrations of copper sulphate on the metabolic rate (mg O2/g/h) Esomus danricus at the end of 24h, 48h, 72h and 96h.Effect of Lethal and Sub-lethal concentrations of copper sulphate on the ventilation frequency (beats/min) of Esomus danricus at the end of 24h, 48h, 72h and 96h.Total Glycogen content in gills of fish exposed to 5.5mg/L and 0.5 mg/L of CuSo4.5H2O and control at the end of every 24h exposure and up to 96h.Total Glycogen content in muscle of fish exposed to 5.5mg/L and 0.5 mg/L of CuSo4.5H2O and control at the end of every 24h exposure and up to 96h.Total Protein content in gills of fish exposed to 5.5mg/L and 0.5 mg/L of CuSo4.5H2O and control at the end of every 24h exposure and up to 96h.Total Protein content in muscle of fish exposed to 5.5mg/L and 0.5 mg/L of CuSo4.5H2O and control at the end of every 24h exposure and up to 96h.SOD and CAT activities in the viscera of fish exposed to 5.5mg/L and 0.5 mg/L of CuSo4.5H2O and control at the end of every 24h exposure and 96h.
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+ Pyrrolizidine alkaloids are naturally occurring genotoxic chemicals produced by a large number of plants. The high toxicity of many pyrrolizidine alkaloids has caused considerable loss of free-ranging livestock due to liver and pulmonary lesions. Chronic exposure of toxic pyrrolizidine alkaloids to laboratory animals induces cancer. This investigation studies the metabolic activation of retrorsine, a representative naturally occurring tumorigenic pyrrolizidine alkaloid, and shows that a genotoxic mechanism is correlated to the tumorigenicity of retrorsine. Metabolism of retrorsine by liver microsomes of F344 female rats produced two metabolites, 6, 7-dihydro-7-hydroxy-1-hydroxymethyl-5H-pyrrolizine (DHP), at a rate of 4.8 ± 0.1 nmol/mg/min, and retrorsine-N-oxide, at a rate of 17.6±0.5 nmol/mg/min. Metabolism was enhanced 1.7-fold by using liver microsomes prepared from dexamethasone-treated rats. DHP formation was inhibited 77% and retrorsine N-oxide formation was inhibited 29% by troleandomycin, a P450 3A enzyme inhibitor. Metabolism of retrorsine with lung, kidney, and spleen microsomes from dexamethasone-treated rats also generated DHP and the N-oxide derivative. When rat liver microsomal metabolism of retrorsine occurred in the presence of calf thymus DNA, a set of DHP-derived DNA adducts was formed; these adducts were detected and quantified by using a previously developed 32P-postlabeling/HPLC method. These same DNA adducts were also found in liver DNA of rats gavaged with retrorsine. Since DHP-derived DNA adducts are suggested to be potential biomarkers of riddelliine-induced tumorigenicity, our results indicate that (i) similar to the metabolic activation of riddelliine, the mechanism of retrorsine-induced carcinogenicity in rats is also through a genotoxic mechanism involving DHP; and (ii) the set of DHP-derived DNA adducts found in liver DNA of rats gavaged with retrorsine or riddelliine can serve as biomarkers for the tumorigenicity induced by retronecine-type pyrrolizidine alkaloids.Pyrrolizidine alkaloids are common constituents of hundreds of plant species of different unrelated botanical families distributed in many geographical regions in the world (1–10). It has been reported that about 3% of the world’s flowering plants contain pyrrolizidine alkaloids (11). Pyrrolizidine alkaloids share a common chemical structure that consists of a necine-base and a necic acid moiety. The alkaloids with the necine base containing an unsaturated double bond at 1, 2-position, such as retronecine, heliotridine, and otonecine are the most toxic pyrrolizidine alkaloids. Among these, the retronecine-based pyrrolizidine alkaloids are abundant and their toxicities have been studied more extensively. Pyrrolizidine alkaloid per se is not toxic. Metabolic activation is required to form a pyrrolic metabolite to exert their toxicity. The high toxicity of many of pyrrolizidine alkaloid-containing plants has caused great loss of free-ranging livestock due to liver and pulmonary lesions. In addition, chronic exposure of toxic pyrrolizidine alkaloids to laboratory rats and mice induces tumors. Because of the concern of human exposure to genotoxic pyrrolizidine alkaloids, toxicity and carcinogenicity of riddelliine, a representative pyrrolizidine alkaloid, have been studied by the National Toxicology Program (NTP) conducted by National Institute of Environmental Health Sciences (NIEHS) (12); and the mechanism of the riddelliine-induced tumorigenicity in rats and mice was studied at the National Center for Toxicological Research (NCTR) (13–16). Results of the mechanistic study showed that riddelliine was metabolized to form a reactive pyrrolic metabolite, 6, 7-dihydro-7-hydroxy-1-hydroxymethyl-5H-pyrrolizine (DHP), as well as riddelliine N-oxide (13). DHP bound to DNA in vitro and in vivo generated DHP-derived DNA adducts that were determined by the 32P-postlabeling/ HPLC analyses (14, 15). The levels of these DNA adduct formations correlate with the liver tumor potency of riddelliine (16) implying that the riddelliine-induced carcinogenesis is through a genotoxic mechanism.Retrorsine is another representative retronecine-based pyrrolizidine alkaloid. Similar to riddelliine, retrorsine is a 12-membered macrocyclic diester pyrrolizidine alkaloid with an α, β-unsaturated double bond linked to the ester group at C-7 position of the retronecine base. The toxicity of retrorsine is due to the metabolic formation of the reactive pyrrolic metabolite (17, 18); however, the mechanism of retrorsine-induced tumorigenicity is not clear. Since retrorsine is structurally similar to that of riddelliine, it is important to know whether the genotoxic mechanism of riddelliine-induced tumorigenicity is also that of retrorsine. In this paper, we report results of the study on the metabolic activation of retrorsine. The detection of DHP-derived DNA adducts formation in liver DNA of rats treated with retrorsine indicates that the tumorigenicity of retrorsine is also through a genotoxic mechanism.Retrorsine, retrorsine N-oxide and troleandomycin (triacetyloleandomycin, TAO) were purchased from Sigma Chemical Co. (St. Louis, MO). [γ-32P]Adenosine 5′-triphosphate ([32P] ATP) (sp. Act. >7,000 Ci/mmol) was purchased from ICN Biomedicals, Inc. (Costa Mesa, CA). Enzymes required for DNA hydrolysis and for 32P-postlabeling/HPLC analysis were purchased and used as previously described (14). DHP and the 3′-monophosphate of 7-(deoxyguanosine-N2-yl) dehydrosupinidine adducts (DHP-3′-dGMP) were prepared in our laboratory (14, 18). The liver microsomes of untreated rats (control microsomes) were prepared as previously described (13) and liver microsomes of female F344 rats pretreated with dexamthesasone (dosed daily with 75 mg dexamethasone/kg body weight intraperitoneally, for three consecutive days) were prepared similarly. Microsomes from rat lung, kidney, and spleen were also prepared similarly.Female F344 rats (3 per group) were obtained from the NCTR breeding colony as weanlings and maintained on a 12 h light-dark cycle. At 8 weeks of age, three animals/ treatment were dosed by oral gavages with retrorsine at 1.0 mg/kg/day in 0.1 M phosphate buffer (pH, 8.0) for three consecutive days. Control animals were gavaged with 0.1 M phosphate buffer only. Twenty-four hours after final dosing, the animals were euthanized by exposure to carbon dioxide and liver tissues were excised and stored at −80 ºC.Metabolism of retrorsine by control or dexamethasone-induced rat liver microsomes was performed in a 1.0 mL incubation volume containing 100 mM sodium phosphate buffer (pH 7.6), 5 mM magnesium chloride, 1 mM NADP+, 8 mM glucose 6-phosphate, 2 units glucose 6-phosphate dehydrogenase, 2 mg control-microsomes, and retrorsine (2 μmol in 50 μL DMSO) at 37ºC for 30 min. After the incubation, the mixture was centrifuged at 105,000 g for 30 min at 4 ºC to remove microsomal proteins. The supernatant fraction was collected and the resulting metabolite mixture was separated by reversed-phase HPLC employing two columns, a sample trap column (ODS, 4.6 × 30 mm) and an analytical column (Prodigy 5 μ ODS, 4.6 × 250 mm, Phenomenex, Torrance, CA). A switching valve was equipped between the sample trap and analytical columns. The sample was first loaded on to the sample trap column and washed with 20 mM ammonium acetate buffer (buffer A) at the flow rate of 1 mL/min, so that the aqueous soluble impurities were directed to a waste bottle. After 5 min the sample trap column was switched to connect the analytical columns and the analysis was performed by eluting with linear gradient of buffer A to 50 % methanol in buffer A (buffer B) over 30 min followed by isocratic elution with buffer B for 25 min.Metabolism of retrorsine by control or dexamethasone-induced microsomes from rat lung, kidney or spleen was similarly conducted, and the metabolites were similarly analyzed by HPLC. For enzyme inhibition study, metabolism of retrorsine was conducted in the presence of 100 μM TAO.Purified calf thymus DNA (2.5 mg, 7.5 μmol) in 2.5 mL of 20 mM K2CO3 (pH 7.5) was reacted with 64 nmol of DHP at 37 ºC for 40 min. After incubation, the reaction mixture was extracted twice with 2.5 mL of a chloroform/isoamyl alcohol mixture (24/1, v/v). The DNA in aqueous phase was precipitated by adding 250 μL of 3 M sodium acetate followed by an equal volume of cold 2-propanol and washed with 70% ethanol. After the DNA was redissolved in 20 mM K2CO3 (pH 7.5), the DNA concentration and purity were analyzed spectrophotometrically. The DNA was stored at −78 ºC prior to 32P-postlabeling/HPLC analysis.The metabolism of retrorsine in the presence of calf thymus DNA (2.0 mg) was conducted in a 2 mL incubation volume with conditions similar to those for metabolism. After incubation, the reaction mixture was ultracentrifuged at 105,000 g for 30 min to remove the microsomal proteins. The clear supernatant was extracted twice with 2 ml of chloroform/isoamyl alcohol (v/v, 24/1). The DNA in the aqueous phase was precipitated and purified as described above.Liver DNA of rats gavaged with retrorsine or with only phosphate buffer was extracted using RecoverEase DNA Isolation Kit (Stratagene, Cedar Creek, TX) according to the manufacturer’s instructions. The concentration and purity of DNA isolated from rat liver were analyzed spectrophotometrically. 32P-Postlabeling/HPLC analyses of DHP-derived DNA adducts 32P-Postlabeling/HPLC analysis was conducted as described previously (13, 14). For quantitation of each sample, the two epimeric DHP-3′-dGMP synthetic standards, in an amount that closely matched the range of modification in the liver DNA samples, were also analyzed in parallel. Statistical comparisons were conducted by analysis of variance by Student’s t-test.In vitro Metabolism of retrorsine by liver microsomes Figure 1 shows the chromatographic profile of reversed-phase HPLC analysis of the in vitro retrorsine metabolism. The chromatographic peak that eluted at 43.6 min contained the recovered substrate, retrorsine. By comparison of HPLC retention times and UV-visible absorption with those of the DHP and retrorsine N-oxide standards (14, 19), the metabolites contained in chromatographic peaks eluting at 25.8 and 34.5 min were identified as DHP and retrorsine N-oxide, respectively (Figure 1).Since the chromatographic peaks eluted prior to 22 min were also detected from the incubation with liver microsomes pre-heated for 10 min, these chromatographic peaks did not contain metabolites of retrorsine. Similar HPLC profiles of the metabolism of retrorsine by liver microsomes of rats treated with dexamethasone conducted under similar conditions were also obtained. The rate of DHP formation from metabolism of retrorsine by dexamethasone -microsomes was about 1.7-fold higher than that by control-microsomes (Table 1).To determine whether or not P450 3A is the principal metabolizing enzyme that catalyzes metabolism of retrorsine to DHP and retrorsine N-oxide, metabolism of retrorsine by rat liver control and dexamethasone-microsomes was conducted in the presence TAO, a specific P450 3A inhibitor. It was found that DHP formation were 77 and 67% reduced, respectively, compared with the metabolism by control and dexamethasone-induced microsomes without TAO (Table 1). Retrorsine N-oxide formation in control and dexamethasone-microsomal metabolism was reduced by 30 and 29%.The quantifications of DHP and retrorsine N-oxide from the in vitro metabolism of retrorsine mediated by the microsomal fractions of extrahepatic tissues, lung, kidney and spleen of rats treated with dexamethasone are shown in Table 2. Compared to the retrorsine metabolism by liver microsomes from control rats, the retrorsine metabolizing enzyme activities in the extrahepatic tissues from dexamethasone-induced rats were much lower; the DHP and N-oxide formation by the microsomal fractions obtained from the extrahepatic tissues of control rats were even lower (data not shown).Rat liver microsomal metabolism of retrorsine in the presence of calf thymus DNA was assayed and the resulting DNA adducts were analyzed by 32P-postlabeling/HPLC. Metabolism of riddelliine in the presence of calf thymus DNA was performed in parallel to determine DNA adducts identification. As previously determined (13–15), female F344 rats fed riddelliine produced a set of eight DHP-derived DNA adducts in liver that were identical to the adduct peaks obtained from the 32P-postlabeling/HPLC analysis of DHP-modified calf thymus DNA (Figure 2A). These eight DHP-derived DNA adducts contained in the chromatographic peaks eluted at 47.6, 48.3, 51.4, 53.9, 55.3, 60.1, 61.0, and 62.6 min are designated as P1, P2, P3, P4, P5, P6, P7, and P8, respectively. The DNA adducts designated as P4 and P6 are DHP-3′-dGMP adducts (13, 14) and the other six DHP-derived adducts (P1, P2, P3, P5, P7, and P8) were characterized as DHP-derived dinucleotides (15). A similar HPLC profile was also obtained from the in vitro metabolism of retrorsine in the presence of calf thymus DNA (Figure 2B). Figure 2C shows the same eight DHP-derived DNA adducts formed from liver DNA of rats gavaged with retrorsine. The level of DNA adducts of retrorsine in liver of rats receiving three daily doses (1 mg/kg/day) is 110.3±18.0 adducts/107 nucleotides.Retrorsine exhibits a variety of toxic responses, including acute toxicity, mutagenicity, DNA cross-linking in cultured bovine kidney epithelial cells in the presence of an external metabolizing system (20), and clastogenic activity (21). Metabolism of retrorsine in vitro and in vivo formed isatinecic acid, pyrrolic metabolites, retrorsine N-oxide, and retronecine (17, 22, 23). Similar to riddelline, retrorsine and retrorsine-containing plants induced liver tumors in rats (24).CYP 3A was found to be the major isozyme for metabolizing monocrotaline and senecionine (25, 26). The inhibition of the DHP formation in the metabolism of retrorsine by TAO (Table 1) indicates that the metabolic formation of DHP was primarily catalyzed by CYP 3A enzyme. This finding is consistent with our previous report that metabolism of riddelliine by liver microsomes of female F344 rats is mainly catalyzed by CYP 3A. Xia et al. (27) performed a comparative study on the metabolism of riddelliine by human and rat liver microsomes and also found that the DHP and riddelliine N-oxide were the major metabolites of riddelliine in an in vitro human microsomal incubation, with the levels comparable to those obtained from rat liver microsomal metabolism. Dexamethasone is not only a potent rat liver CYP3A inducer that induces 1.7-fold rat liver retrorsine metabolizing enzyme activity but also induced microsomal retrorsine-metabolizing enzyme activity in lung, kidney and spleen. The CYP3A activity in the extrahepatic tissues of uninduced rats was very low; in fact, under our experimental conditions (2 mg microsomal protein, 30 min incubation) the DHP and retrorsine N-oxide were hardly detected. However, using dexamethasone-induced lung, kidney, or spleen microsomes, we were able to determine the formation of DHP and retrorsine N-oxide. The levels of metabolism by extrahepatic tissue microsomes from the dexamethasone-induced rats were 18 to 32-fold lower than the metabolism of retrorsine by control liver microsomes.We have previously found that the DHP-derived DNA adducts are formed in liver of male and female F344 rats fed riddelliine. The levels of the DHP-derived DNA adducts in liver DNA of rats treated with riddelliine correlated with the liver tumor incidence indicating the DHP-derived DNA adducts that are responsible for riddelliine-induced liver tumorigenicity (15); this riddelliine-induced tumorigenesis is mediated through a genotoxic mechanism. The current study shows the similar set of DHP-derived DNA adducts that was found in the liver DNA of rats treated with retrorsine and indicates the retrorsine-induced tumorigenesis may be also due to a genotoxic mechanism. The formation of these DNA adducts from retrorsine metabolism suggests that these DHP-derived DNA adducts are responsible for retrorsine-induced liver tumorigenicity as well as the other genotoxicities. Based on the results reported in this study and published previously (13), these DHP-derived DNA adducts are potential biomarkers of pyrrolizidine alkaloid-induced tumorigenicity.Based on the present findings, we propose that the metabolic activation of retrorsine leads to the formation of DHP-derived DNA adducts and liver tumors (Figure 3). The pyrrolic metabolite, dehydroretrorsine, can (i) bind to cellular DNA followed with hydrolysis to form the DHP-derived DNA adducts, or (ii) be hydrolyzed to form DHP followed by reaction with DNA to form the DHP-derived DNA adducts (Figure 3). Our results show that DHP is the common reactive metabolite generated by the retronecine-type pyrrolizidine alkaloids. Consequently, the formation of DHP-derived DNA adducts may well be important biomarkers for exposure to pyrrolizidine alkaloids.Reversed-phased HPLC analysis of metabolites formed from metabolism of retrorsine. For the conditions for HPLC analysis see Materials and Methods.32P-Postlabeling/HPLC analysis of DHP-derived DNA adducts formed from (A) DHP-modified calf thymus DNA, (B) calf thymus DNA incubated with the in vitro metabolism of retrorsine, and (C) liver DNA of rats gavaged with retrorsine. The eight chromatographic peaks eluted at 47.6, 48.3, 51.4, 53.9, 55.3, 60.1, 61.0, and 62.6 min are the identified DHP-derived DNA adducts designated as P1, P2, P3, P4, P5, P6, P7 and P8, respectively. For the conditions for 32P-postlabeling/HPLC analysis see Materials and Methods.The proposed metabolic activation and detoxification pathways of retrorsine.Quantification of DHP and PA N-oxide formation in an in vitro PA metabolism by rat liver microsomesData represent the mean ±SD (n=3). For experimental details, see Materials and Methods.Statistically significant difference (p<0.01) between groups in the same column.Statistically significant difference (p<0.05) between groups in the same row.Quantification of DHP and retrorsine N-oxide formation in an in vitro PA metabolism with lung, kidney, and spleen microsomes from the rats pretreated with dexamethasoneData represent the mean ±SD (n=3).For experimental details, see Materials and Methods.This research was supported in part by appointment (Y.W.) to the Postgraduate Research Program at the NCTR administered by the Oak Ridge Institute for Science and Education through an interagency agreement between the U.S. Department of Energy and the FDA.
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+ Arsenic is an environmental toxicant, and one of the major mechanisms by which it exerts its toxic effect is through an impairment of cellular respiration by inhibition of various mitochondrial enzymes, and the uncoupling of oxidative phosphorylation. Most toxicity of arsenic results from its ability to interact with sulfhydryl groups of proteins and enzymes, and to substitute phosphorus in a variety of biochemical reactions. Most toxicity of arsenic results from its ability to interact with sulfhydryl groups of proteins and enzymes, and to substitute phosphorus in a variety of biochemical reactions. Recent studies have pointed out that arsenic toxicity is associated with the formation of reactive oxygen species, which may cause severe injury/damage to the nervous system. The main objective of this study was to conduct biochemical analysis to determine the effect of arsenic trioxide on the activity of acetyl cholinesterase; a critical important nervous system enzyme that hydrolyzes the neurotransmitter acetylcholine. Four groups of six male rats each weighing an average 60 ± 2 g were used in this study. Arsenic trioxide was intraperitoneally administered to the rats at the doses of 5, 10, 15, 20mg/kg body weight (BW), one dose per 24 hour given for five days. A control group was also made of 6 animals injected with distilled water without chemical. Following anaesthesia, blood specimens were immediately collected using heparinized syringes, and acetyl cholinesterase detection and quantification were performed in serum samples by spectrophotometry. Arsenic trioxide exposure significantly decreased the activity of cholinesterase in the Sprague-Dawley rats. Acetyl cholinesterase activities of 6895 ± 822, 5697 ± 468, 5069 ± 624, 4054 ± 980, and 3158 ± 648 U/L were recorded for 0, 5, 10, 15, and 20 mg/kg, respectively; indicating a gradual decrease in acetyl cholinesterase activity with increasing doses of arsenic. These findings indicate that acetyl cholinesterase is a candidate biomarker for arsenic-induced neurotoxicity in Sprague-Dawley rats.Arsenic is widely distributed in nature, being found in food, the soil, water and airborne particles; it derives from both natural and human activities [1]. More than 80% of arsenic compounds are used to manufacture products with agricultural applications such as insecticides, herbicides, fungicides, algicides, sheep dips, wood preservatives, dye-stuffs, and medicines for the eradication of tapeworms in sheep and cattle [2]. Arsenical drugs are still used used in treating certain tropical diseases such as African sleeping sickness and amoebic dysentery, and in veterinary medicine to treat parasitic diseases, including filariasis in dogs and black head in turkeys and chickens [2]. Recently, arsenic has been used as an anticancer agent in the treatment of acute promeylocytic leukemia, and its therapeutic action has been attributed to the induction of programmed cell death (apoptosis) in leukemia cells [3].A large number of people are exposed to arsenic chroniclly throughout the world. Exposure to arsenic occurs via the oral route (ingestion), inhalation, dermal contact, and the parenteral route to some extent. Humans can be exposed to arsenic through the intake of air, food and water. Although food is usually the major source of arsenic exposure, most adverse effects have been associated with consumption of arsenic-contaminated drinking water. Occupational sources of arsenic to human workers include vineyards, ceramics, glass- making, smelting and refining of metallic ores, during production and use of arsenic containing agricultural products like pesticides and herbicides [4]. The gastrointestinal tract of humans and most experimental animals readily absorbs ingested inorganic arsenic (>90%). Following absorption, arsenic compounds through blood circulation are distributed in various tissues (blood, liver, kidney, lung, skin) [5].Analyzing the toxic effects of arsenic is complicated because the toxicity varies according to its oxidation state, its solubility and many different inorganic and organic forms. [6]. Several studies have indicated that the toxicity of arsenic depends on the exposure dose, frequency and duration, biological species, age, gender as well as on individual susceptibilities, genetic and nutritional factors [7].The major metabolic pathway for inorganic arsenic in humans is methylation. Arsenic trioxide is methylated to two major metabolites via a non-enzymatic process to monomethylarsonic acid (MMA), which is further methylated enzymatically to dimethyl arsenic acid (DMA) before excretion in the urine [8]. This methylation mechanism has been widely accepted, and the metabolites MMA (V) and DMA (V) have been consistently observed in human urine [9–11].Generally, the toxicity of heavy metals is largely due to their reactions with sulfhydryl groups [12]. Lipophilic organometallic compounds easily cross the blood-brain barrier, whereas inorganic metallic compound also reach the brain tissue [12]. Acute and chronic toxic effects of inorganic arsenic involve many organ systems, including the central nervous system (CNS) [13–15]. In experimental animals, arsenic has been shown to affect hepatic mitochondrial enzymes [16]. It can also pass the blood-brain barrier, be accumulated in the brain, and can exert neurochemical effects [17, 18].The inhibition of acetycholinesterase (AChE) in the nervous tissue and other target organs is generally considered to be the critical effect leading to the acute toxicity of many toxic chemicals [19–22]. Nagaraju and Desiraju [23] further indicated that AChE activity in rats was inhibited in some regions of the brain following inorganic arsenic intake. Despite its deleterious actions on the CNS, there are very few studies of the effects of chronic or acute consumption of arsenic on brain and behaviour. The present study was done to examine the effect of arsenic on the activity of acetyl cholinesterase; a critical important nervous system enzyme that hydrolyzes the neurotransmitter acetylcholine.Arsenic trioxide (As2O3) was purchased from Fischer-Scientific, Houston, TX, USA. Cholinesterase (PTC), and heparin were purchased from Sigma-Aldrich (St.Louis, MO, USA).Healthy adult male Sprague-Dawley rats (8–10 weeks of age, with average body weight (BW) of 60 ± 2 g) were used in this study. They were obtained from Harlan-Sprague-Dawley Breeding laboratories in Indianpolis, Indiana, USA. The animals were randomly selected and housed in polycarbonate cages (three rats per cage) with steel wire tops and corn-cob bedding. They were maintained in a controlled atmosphere with a 12h:12h dark/light cycle, a temperature of 22 ± 2° C and 50–70% humidity with free access to pelleted feed and fresh tap water. The animals were supplied with commercially available dry food pellets from PMI Feeds Inc. (St. Louis, Missouri). They were allowed to acclimate for 10 days before treatment.Groups of six rats each were treated with four different arsenic trioxide dose levels, 5, 10, 15 and 20 mg/kg BW. Arsenic trioxide was diluted with distilled water (as required) and intraperitoneally administered to animals at the doses of 0, 5, 10, 15 and 20 mg/kg BW, one dose per 24h given for 5 days. Each rat received a total of five doses at 24h intervals. The cumulative doses of arsenic trioxide given to rats were thus 25, 50, 75 and 100 mg/kg BW. Distilled water was administered to the 6 animals of control group in the same manner as in the treatment groups. The acute bioassay was performed following standard test protocol.At the end of the treatment period, rats were anesthetized using 95% CO2 for 70 seconds. Immediately following anaesthetization blood specimens were collected using heparinized syringes to prevent clotting, and the plasma was separated by centrifugation at low speed (2000 g) for 10 min. Plasma samples was evaluated for enzyme identification and quantification.Measurement of serum cholinesterase has been used to assess liver function, and monitor exposure to organo-phosphorus insecticides. The enzyme is depressed in acute hepatitis, hepatic metastases and alcoholic cirrhosis.The reaction for cholinesterase assay is as follows:Cholinesterase hydrolyzes Propionylthiocholine to form thiocholine that reacts with Dithio bis2-nitrobenzoic acid to yield the yellow 5-thio-2-nitrobenzoate with an absorbance maximum at 405 nm. Therefore the rate of change in absorbance at 405 nm is directly proportional to cholinesterase activity.Cholinesterase reagent (PTC) was prepared by reconstituting with the volume of deionized water indicated on the vial. The temperature of the reaction mixture was maintained at 30° C. Spectrophotometer wavelength was set to 405 nm and absorbance reading to zero with H2O as reference. The reagent was warmed in water bath to assay temperature (30° C). To cuvet labeled TEST, 1.0 ml of cholinesterase (PTC) reagent was added and placed in temperature controlled cuvet compartment. ten μl of serum was added to the above reagent, mixed immediatel by inversion and incubated at 30° C for 15 seconds. The absorbance was read and recorded as (A) of TEST at 405 nm versus water as reference. This is called Initial A. The cuvet labeled TEST was incubated again at 30° C and absorbance was recorded after exactly 30 seconds following the initial absorbance reading. This is called Final A, multiplied by 2 to obtain the change in absorbance per minute at 405nm (Δ A per minute). The cholinesterase activity (U/L) of the sample was determined by using the following formula.Where:Δ A per min = change in aborbance per minute at 405 nm.TV = Total volume (1.01 ml)SV = Sample volume (0.01 ml)13.6 = millimolar absorptivity of 5- thio- nitrobenzoic acid at 405 nm.LP = Lightpath (1-cm)1000 = conversion of units per ml to units per liter.Figure 1 show the experimental data obtained from the analysis of acetyl cholinesterase. The results indicate acetyl cholinesterase activities of 6895 ± 822, 5697 ± 468, 5069 ± 624, 4054 ± 980, and 3158 ± 648 U/L for 0, 5, 10, 15, and 20 mg/kg BW respectively. As shown in this figure there was a dose-response relationship with respect to arsenic inhibition of acetyl cholinesterase in the blood.There is increasing interest in the development of new methods for assessing chemical toxicity, and the use of knowledge of mechanisms of toxic actions to improve this assessment. In order to determine the effect on serum acetyl cholinesterase (AChE), adult male Sprague-Dawley rats were exposed for five days to four different concentrations (5, 10, 15, 20 mg/kg) of arsenic trioxide. The data obtained from this study clearly show that arsenic trioxide significantly decreased the activity of serum acetyl cholinesterase in a dose-dependent manner (Figure 1). The results agree with previous studies that demonstrated a decreased activity of acetyl cholinesterase in neuroblastoma cells of mice [24], in rat whole brain [23, 25] and in two models of fish [26].In this study, the decrease in the activity of acetyl cholinesterase was positively correlated with the chemical dose. The inhibition of acetyl cholinesterase by arsenic trioxide is a puzzling phenomenon because AChE does not contain the structural features usually associated with the inhibition of enzymes by arsenic. Trivalent arsenic compounds are potent inhibitors of a number of enzymes [27] but the mechanism of this inhibition is the reaction of the arsenical with free sulfhydryl groups, notably those of reduced lipoic acid, to form cyclic thio-arsenite diesters. However, other compounds known to be sulfhydryl reactants such as organomercury compounds or iodoacetamide did not inhibit AChE [28]. The enzyme has been found to contain cysteine only in the form of disulfide bridges and not as free thiol [29].Acetyl cholinesterase (AChE) is one of many important enzymes needed for the proper functioning of the nervous systems of humans, other vertebrates and insects. Certain chemical classes of pesticides, such as organophosphates (OPs) and carbamates interfere with or inhibit cholinesterase. Acute toxicity manifests as a cholinergic crisis with excessive glandular secretions, altered mental status, and weakness. Several delayed syndromes associated with these pesticides exposure are myasthenic-like syndrome, peripheral neuropathies, neuropsychiatric abnormalities and extrapyramidal disorders [30].AChE has been shown to be neurotoxic in vivo and in vitro; it accelerates assembly of amyloid peptide in Alzheimer’s fibrils, leading to cell death via apoptosis [31]. Brain AChE has been shown to be toxic to neuronal (Neuro 2a) and glial-like (B12) cells [32]. There are also reports that transgenic mice over-expressing human AChE in brain neurons undergo progressive cognitive deterioration [33]. Organophosphates ability to affect acetylcholine neurotransmission, poisoning commonly manifests with acute dysfunction of the autonomic, central and peripheral nervous system. Failure to recognize these manifestations can result in worsening toxicity, delayed complications, and death [33].Despite its deleterious actions on the central nervous system, there have been very limited studies of the effects of arsenic on the brain and behaviour. Therefore, the present study provided new insights on the neurotoxicity of arsenic, and indicated that acetyl cholinesterase activity can be used as a biomarker of this neurotoxicity if the specific pre-exposure conditions are characterized.Effect of Arsenic trioxide on the activity of serum cholinesterase in Sprague-Dawley rats.This research was financially supported by NIH-RCMI Grant No. 1G12RR13459. We thank President, Jackson State University, Dr. Ronald Mason Jr. and Dr. Abdul Mohamed, Dean of the College of Science, Engineering and Technology, for their support and advice in this research.
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+ Human health is a major concern when considering the disposal of large quantities of animal waste. Health concerns could arise from exposure to pathogens and excess nitrogen associated with this form of pollution. The objective was to collect and analyze health data related to selected bacterial infections associated with the use of animal waste in Louisiana. An analysis of adverse health effects has been conducted based on the incidence/prevalence rates of campylobacteriosis, E. coli O157:H7 infection, salmonellosis and shigellosis. The number of reported cases increased during the summer months. Analysis of health data showed that reported disease cases of E. coli O157:H7 were highest among Caucasian infants in the 0–4 year old age category and in Caucasian children in the 5–9 year old age category. Fatalities resulting from salmonellosis are low and increases sharply with age. The number of reported cases of shigellosis was found to be higher in African American males and females than in Caucasians. The high rate of identification in the younger population may result from the prompt seeking of medical care, as well as the frequent ordering of stool examination when symptoms become evident among this group of the population. The association with increasing age and fatality due to salmonellosis could be attributed to declining health and weaker immune systems often found in the older population. It is concluded that both animal waste and non-point source pollution may have a significant impact on human health.Animal waste from dairy and poultry operations is an economical and commonly used fertilizer in the state of Louisiana. The application of animal waste to pasture lands not only is a source of fertilizer, but also allows for a convenient method of waste disposal. One type of animal waste product that is commonly used is dairy lagoon sediment or effluent. Most dairies in Louisiana have a one or two stage lagoon that collects liquid and semi-liquid manure from loafing barns and milking parlor areas. The solid waste settles in the lagoon where it is reduced by the process of anaerobic digestion. The liquid effluent is often pumped onto fields or is recycled for other uses. Periodically, the lagoon must be emptied of the sediment build up. The sediment is typically agitated in order to suspend it into a semi-liquid state and is pumped onto the fields. The disposal of animal wastes on land is a potential non point source of water degradation. Runoff and percolation could possibly transport organic matter and nutrients to surface and ground water. Animal wastes applied to the land come from wastes that have been removed from feeding facilities, runoff from feeding areas, and waste from animals on pasture and rangeland. Proper application of animal wastes provides nutrients for crop production and also reduces surface runoff.The health of humans is of concern when considering large quantities of animal waste. Some of the main concerns include the exposure to pathogens and excess nitrogen associated with this form of pollution. Animal waste can contain pathogens, such as fecal coliform bacteria and viruses that can contaminate drinking water and cause gastrointestinal illnesses. High levels of nitrogen leaching into drinking water supplies can increase the risk of methemoglobinemia [1]. In 1996, the Centers for Disease Control linked the high nitrate levels in Indiana well water near feedlots to spontaneous abortions in humans [2].Nitrites are relatively short-lived because they’re quickly converted to nitrates by bacteria. Nitrites produce a serious illness (brown blood disease) in fish, even though they don’t exist for very long in the environment. Nitrites also react directly with hemoglobin in human blood to produce methemoglobin, which destroys the ability of blood cells to transport oxygen. This condition is especially serious in babies under three months of age as it causes a condition known as methemoglobinemia or “blue baby” disease. Water with nitrite levels exceeding 1.0mg/L should not be given to babies. Nitrite concentrations in drinking water seldom exceed 0.1mg/L [3].Nitrate is a major ingredient of farm fertilizer and is necessary for crop production. When it rains, varying nitrate amounts wash from farmland into nearby waterways. Nitrates also get into waterways from lawn fertilizer run-off, leaking septic tanks and cesspools, manure from farm livestock, animal wastes, and discharges from car exhaust. Nitrates can be reduced to toxic nitrites in the human intestine. The U.S. Public Health Service has established 10mg/L of nitrate-nitrogen as the maximum contamination level allowed in public drinking water [3]. Nitrate-nitrogen levels below 90mg/L and nitrite levels below 0.5 mg/L seem to have no effect on warm-water fish, but many cold water fish are more sensitive. The recommended nitrite minimum for salmon is 0.06mg/L [4].Ammonia is a toxic form of nitrogen. Open air lagoons emit ammonia into the air [5]. One survey of residents living in the vicinity of a 2,500-sow facility found much higher reports of respiratory problems than were recorded from the neighbourhoods of farms where no livestock was raised [6].Many regulations for water are found in the Clean Water Act. The H.R. 961, a bill to reauthorize the Clean Water Act, was approved by the House of Representatives. The bill would reverse a 1994 Federal circuit court ruling that land application of livestock manure from a concentrated animal feeding operation is a point source which is subject to permit and enforcement provisions of the CWA (Concerned Area Residents for the Environment v. Southview Farm, No. 93-9229 ((2 Cir. Sept. 2, 1994)). The Supreme Court recently declined review of the Southview Farm case [7].Drinking water quality has been improving over time. According to the Centers for Disease Control and Prevention (CDC), the proportion of reported disease outbreaks that can be attributed to problems at public water treatment systems has steadily declined, from 73% in 1989 – 1990 to 30% in 1995 – 1996. It is possible that this decrease reflects the improvements in water treatment and in operation of plants [8].Between the years of 1997 and 1998, 13 states reported a total of 17 significant illness outbreaks associated with drinking water [9]. These caused an estimated 2,038 persons to become ill. CDC keeps records on occurrences and causes of outbreaks of illness related drinking water and recreational water. Many ofthe outbreaks that occur are often missed by the public health officials because some of the illnesses that are associated with the outbreaks are not perceived to be water related [9].In 1999, E. coli contaminated water at the Washington County fairgrounds in New York State caused the death of two people and illness in over 1000 others. The source of contamination was probably cattle fecal material from a nearby barn, which was swept into the soil by storm runoff, and then leached into the aquifer [10].Drinking water health effects are not limited to gastrointestinal illness associated with microbes. Drinking water can transmit bacteria, micro organisms, and chemicals that are capable of causing disease. The symptoms can be acute, such as diarrhoea and dehydration, or they can be long term effects that include infertility and reproductive health effects, or chronic illnesses such as cancer [11].An analysis of the adverse health effects that are associated with the types of animal waste studied in this project was conducted by the usage of materials received from Louisiana Health and Hospital Systems, Infectious Disease Epidemiology Section, New Orleans, LA. Other relevant information was collected from Louisiana Department of Environmental Quality (Shreveport, LA). Illnesses having a bacterial origin were checked in regards to the number of cases in the state of Louisiana and the incidence rates as well. Patterns in regards to race, age and gender were recorded. The parishes with high incidence rates of these illnesses were identified. Data from the 2000 census were also utilized in order to identify populations at risk so that comparisons of disease incidence rates by county, area, or other characteristics could be made. An analysis of the total amounts of animal waste, cattle waste, poultry waste, amounts of nitrogen in waste and the amount of phosphorus in waste was carried out in reference to the top four parishes with high incidence rates of diseases with bacterial origins possibly associated with animal waste. Relative information found was mapped to illustrate the distribution of diseases by areas.Demographic data and data on the numbers of cases of campylobacteriosis, shigellosis, salmonellosis, and E. coli infections in Louisiana were collected. Cases of diseases were divided by the population at risk to determine the incidence/prevalence rates. These disease rates were used as the basis or end-point for comparing the health risks by parishes, gender, race, etc. Additional information was collected on the amounts of animal wastes generated by parish, and a linear regression analysis was performed to determine if there is a correlation between the level of waste and the incidence of diseases by parish.An analysis of information received from the Louisiana Department of Health and Hospitals Infectious Disease Epidemiology Section for the years 1988 – 2001 was conducted for diseases caused by campylobacter, E. coli 0157:H7, salmonella and shigella.As shown in figure 1, the majority of cases of campylobacteriosis reported occurred during the summer months of May, June, July and August. The fewest number of cases were seen during the months of December and January. This seasonal trend was noted for the years of 1988 – 2001. Between the years of 1988 and 2001, the highest number of cases of campylobacter reported occurred in 1992 with approximately 280 cases being reported. There has been a gradual decrease of the number of cases since 1992. There were 150 cases reported in 1998 and 130 cases reported during the year 2001. This shows a 50% reduction. In 1999, it was noted that the number of campylobacter cases reported was twice as high for Caucasians than that of African Americans. Caucasians had a rate of 2.6 per 100,000 people and African Americans had a lower rate of 1.3 per 100,000 people. It was found that the majority of the cases reported during this same year were in children within the 0–4 year’s age group. The lowest number of cases was seen in the 15–19 year age group. The parishes that had the highest rates of campylobacteriosis per 100, 000 people in 1999 were Red River, Jackson, and Terrebonne Parishes. The reported rates per 100,000 people for these parishes were 22, 13, and 12 per 100,000 people, respectively. In the year 2001, the highest rate was seen in Washington Parish with 1.62 per 100,000 people followed by Red River Parish with 1.08 per 100,000 people [12].An analysis of E. coli O157:H7 cases for the state of Louisiana showed that for the years 1996–2001, the highest reported cases were in 1997 followed by the year 2000.Figure 2 shows that of the years 1996–2001, the majority of the cases were reported for young Caucasian infants and children between the ages of 0 and 9 years of age. The highest number of cases reported during the years of 1996–2001 was found in St. Tammany Parish with a total of 11 cases. Figure 3 shows that the E. coli O157:H7 number of cases reported increased during the months of June, July, August and November. These months reported 12, 13, and 9 cases, respectively. During 1999, sixty-four percent of the cases reported occurred between the months of June and September. For the years of 1996–2001, 1 case was reported during the month of February [12].An analysis of salmonellosis for the state of Louisiana revealed that during the year 1999, there were a total of 718 reported cases of salmonellosis. This showed a 17% decrease in the number of cases that were reported in 1998. When looking at the years of reporting ranging from 1965–2001, the majority of the cases occurred during the late 1980s and among Caucasian infants, adolescents and the older population. The numbers of cases of salmonellosis in infants that were reported between the years of 1997–2001 were highest in the 0–1 year of age group and children in the 1–5 year of age group. According to the Louisiana Department of Health and Hospitals, the case fatality for salmonella is extremely low (0.2%). As shown in Figure 4, the case fatality increases with increasing age. Two parishes were noted as having the highest number of reported salmonellosis cases. St. Tammany Parish reported a total of 47 cases followed by Washington Parish with a total of 46 reported cases [13]. St. Tammany Parish, with a rate of 5.26 per 100,000, was found to be among the top three parishes reporting salmonellosis for the year 2000. Also among the top three during the year 2000 were Caldwell and Terrebonne Parishes with rates of 7.14 and 4.54 per 100,000, respectively. Terrebonne Parish was noted again in the year 2001 with a high rate of 4.13 per 100,000 [12].In 1999, the number of shigellosis cases that were reported was 227. This was at a rate of 5.3 per 100,000 people. The reported cases showed that shigellosis occurred most often in African American males and African American females when compared to the number of cases reported for Caucasian males and females. The sex-race specific rates were 8.3 per 100,000 for African American males and 5.9 per 100,000 for African American females. The lower occurrence rates for Caucasian males and Caucasian females were 1.5 and 2.4 per 100,000 respectively. In 1999 the parishes having the highest rate of cases of shigellosis that were reported were West Baton Rouge, Livingston, St. John and Assumption. The rates reported for these parishes were 21, 14, 13 and 13 per 100,000, respectively [13]. In the year 2000, Desoto Parish at a rate of 1.58 per 100,000 and St. Tammany Parish at a rate of 1.45 per 100,000 had the highest reported rates of shigellosis. St. Tammany Parish appeared as having one of the highest rates again in the year 2001 at 1.67 per 100,000 followed by Washington Parish with a rate of 1.16 per 100,000. The parish having the highest reported rate of shigellosis during the year 2001 was Vermillion (4.2 per 100,000).An analysis of the parishes that were reported as having the highest incidence rates of camphylobacteriosis, salmonellosis, and shigellosis for the years of 1999, 2000, and 2001 are found in Table 1.Parishes having the highest number of reported cases for these same illnesses in 1999 are shown in Table 2. A comparison of the data regarding the parishes reporting the highest number of cases by rates for campylobacteriosis, salmonellosis, and shigellosis within the years of 1999–2001 showed that four parishes on three or more occasions were noted as being among those parishes with the highest rates for the studied illnesses. These four parishes are Red River, St Tammany, Terrebonne, and Washington. Washington Parish had the highest amount of cattle waste generated in the year 1997 among these parishes [14].As seen in Table 3, Washington Parish ranked second in the state among the sixty-four parishes with 490,000 tons of cattle waste and ranked thirteenth with regards to poultry waste generated with only 64 tons. The amounts of reported cattle waste generated in Red River, St. Tammany, and Terrebonne Parishes were 200,000, 65,000, and 33,000 tons, respectively. In regards to the amount of poultry waste generated by these parishes, lower amounts were reported with Red River having less than one ton. St. Tammany and Terrebonne Parishes reported 39 and 2 tons of poultry waste generated respectively. The state of Louisiana showed an increase in the number of heads of cattle between the years of 1987 and 1997. The number of cattle in 1987 was reported as being 813,181 and in 1997 it was 905,193 cattle. This clearly indicates a 7% increase. This increase is also reflected in the 6,600,000 tons of cattle waste generated in 1987 compared to the 7,100,000 tons that were generated in 1997. Poultry numbers showed a 14% increase between the years of 1987 and 1997. In 1987, it was reported that the state’s poultry numbers were 17,068,652 and increased to 20,335,050 in 1997. The amounts of poultry waste generated for 1987 and 1997 were reported as being 540,000 and 610,000 tons, respectively. While St. Tammany, Terrebonne, and Washington reported decreases in the amount of waste generated for cattle and poultry between the years of 1987 and 1997, Red River showed a 101 % increase in cattle waste generated [14].Educational, socioeconomic and racial makeup of Red River, St. Tammany, Terrebonne, and Washington Parishes were analyzed from the 2000 census. Based upon the census, of these four parishes Terrebonne had the largest percentage of residents having less than a ninth grade education as well as the lowest percentage of residents having a high school education or higher. In regards to the poverty level, 22.23% of the residents of Red River parish lived below this level. The percentage of residents living below the poverty level for St. Tammany, Terrebonne, and Washington Parishes were 8.86, 17.2, and 21.68 %, respectively.Table 4 shows that although Red River Parish has the smallest total population (9,622 residents) in comparison to the other three parishes, it has the largest percentage of African Americans (40.9 %) compared to 57.9 % Caucasian residents. Eighty-seven percent of the residents of St. Tammany are Caucasian, and only a small percentage (9.9%) of the residents are African Americans. The parish having the largest total population of the three compared parishes is St. Tammany Parish with a total of 191,268 residents. The largest percentage (29.9%) of St. Tammany residents is in the 25–44 year old category. Terrebonne Parish also recognizes its 25–44 year old category as being the majority of its residents at 29.8 %. Terrebonne Parish is also known for having the largest amount of water area and land area among the four discussed parishes. Terrebonne has 824.97 square miles of water area and 1254.93 square miles of land area [15].Tables 5, 6, 7, 8, 9 and 10 show the distribution of campylobacter, salmonella, and shigella in the various parishes for the years 2000 and 2001. The distribution of campylobacter for parishes having a rate of 0 per 100,000 for the year 2000 totalled 31 parishes and was 36 for the year 2001. The numbers of parishes reporting rates of 0.01 to 1.00 for campylobacter for these same years were 28 and 26 parishes respectively. In 2000, three parishes had rates between 1.01 and 2.00 and only two parishes reported rates above 2.00. During 2001, only two parishes had rates above 1.01.The distribution of salmonella for parishes having a rate of 0 per 100,000 for the years 2000 and 2001 were 13 and 4, respectively. For these same years, the number of parishes with rates ranging between 0.01 and 1.00 were 21 and 9 per 100,000, respectively. In the year 2000, 16 parishes reported rates ranging between 1.01 and 2.00, 8 parishes reported rates ranging between 2.01 and 3.00, and 5 parishes had rates greater than 3.01 per 100,000 people. In the year 2001, 34 parishes had rates ranging between 1.01 and 2.0, 13 parishes reported rates ranging between 2.01 and 3.00 and 4 parishes reported rates higher than 3.01.There were a total of 30 parishes that reported rates per 100,000 of zero for shigella during the years of 2000. In the year 2001, a total of 36 parishes also reported a rate of zero for Shigella. The number of parishes that reported rates ranging from 0.01 to 1.00 for the years 2000 and 2001 were 25 and 24 respectively. Only 9 parishes reported rates higher than 1.01 in 2000 while in 2001, 3 parishes reported this same rate. In 2001 only one parish had a rate that was greater than 2.00 per 100,000.Using the described statistical methods, no apparent significant correlation (p>0.05) between the amount of animal waste and the studied disease incidence rates among the parishes in the state of Louisiana was observed.The state of Louisiana is reported as having a total of 7,876,528 acres of farmland. Of this amount, the parishes of Red River, Washington, Terrebonne and St. Tammany have been reported as having 113,176 acres, 100,006 acres, 52,873 acres and 41,863 acres respectively [15].Campylobacteriosis is a Class C disease and must be reported to the state within five business days. Campylobacteriosis is estimated to affect about one percent of the population in the United States. Many of the cases are undiagnosed and unreported. Approximately six cases for each 100,000 persons in the U.S. population are reported to the Centers for Disease Control and Prevention (CDC), while in Louisiana the reported incidence is about five per 100,000. Most cases of campylobacteriosis are associated with the handling of raw poultry or eating raw or under cooked poultry meat. Water may be a source for sporadic cases or outbreaks [12]. An increased prevalence of campylobacteriosis is possible in humans inhabiting areas of intensive cattle production [16]. It is believed that the high rate of occurrence in infants is likely to be the result of cross contamination from parent to child. There is a reported increase of cases during the summer months [12]. The distribution of camyplobacter during the years 2000 and 2001 remained relatively the same. The majority of the parishes had rates less than 1.01 per 100,000. There were only two parishes that had rates greater than 2 per 100,000 in the year 2000 and none having this rate in the year 2001.E. coli O157:H7 is found in both dairy and beef herds in the majority of cattle farms across the United States [12]. Infection from this organism is considered as a Class B disease and must be reported to the state within one business day. It became reportable in Louisiana in 1996, with the number of cases ranging between five to twenty cases per year. The detection is higher among infants than among children and adults because infants with diarrhea are more likely to be brought to a medical facility to have stool examination. A higher number of cases are reported for Caucasian than for African Americans. A possible explanation for the low number of African Americans reporting the disease is possibly the result of a lack of access to medical care for this community. Because of this, more screening is done for Caucasian than for African Americans. Another possible explanation to the differences in the amount of cases reported between the Caucasian and African American community could be that the parishes reporting the highest rates of this illness also had a much lower minority population as compared to the majority population [12]. The Surveillance Summaries of the Morbidity and Mortality Weekly Reports noted that between the years of 1997 and 1998, there were four outbreaks caused by bacteria; three were attributed to E. coli O157:H7 and one to Shigella sonnei. One of these outbreaks occurred in the state of Illinois and involved three persons who drank from an untreated well located near a cattle pasture. Another outbreak involving 26 people was noted to have occurred in the state of Georgia at a water park. It is believed that a fecal accident in the children’s wading pool was the source [9].Salmonellosis is a Class B disease and must be reported to the state within one business day. The rates observed in Louisiana are the same as in the rest of the United States. The increase in the reported incidence rates after 1980 is possibly the result of improved reporting by all facilities. The seasonal pattern favoring the summer months is thought to be because salmonella experiences better growth at higher temperatures. The high rate of identification in the younger population may result from the prompt seeking of medical care when symptoms become evident among infants and young children. The high rate of identification may also be the result of more frequent ordering of stool examination from children when healthcare workers investigate diarrhea symptoms. These practices result in the over-sampling of the child population. Access to medical care and more screening among the Caucasian community may explain the differences in the reported rates by race. Although case fatality for salmonellosis is known to be extremely low at 0.2%, it appears to increase sharply with age. This may be due to the declining health and weaker immune systems often found among the older population [12]. The Infectious Disease Epidemiology Annual Report states that the rates of Salmonella infection in Terrebonne Parish are consistently high. Terrebonne Parish was among the parishes having the highest rate of salmonellosis for the years 2000 and 2001[12]. It is suggested that the medical facilities are good at culturing Salmonella, but the amount of water and land availability could also be a factor. The distribution of salmonella during the year 2000 and the year 2001 showed an increase during the year 2001 in the number of parishes in Louisiana that had rates greater than 1.01 per 100,000. The number of parishes having rates of 1.01 to 2 per 100,000 more than doubled from 16 in the year 2000 to 34 in the year 2001. The number of parishes having rates greater than 3 remained the same during both years, while those parishes with rates of 0 per 100,000 decreased from 13 parishes to four parishes.Shigellosis is a Class B disease and must be reported to the state within one business day. Although reported infrequently in the United States, outbreaks of shigellosis have been associated with food, drinking water and recreational activities in water. Only 0.3% of the cases are reported as occurring in outbreaks. Some shigella infections in the U.S. occur among young adult males as a result of sexual transmission among homosexual men. There is no peak among adult males in the state of Louisiana. The race distribution shows a high rate among Caucasian infants as compared to African American infants. This could be the result of Caucasian infants having better access to medical care. Shigella seems to occur throughout the year without any seasonal peaks [12]. The distribution of shigellosis during the years of 2000 and 2001 remained relatively the same. The majority of the parishes had rates less than 1.01 per 100,000. The year 2000 did not have any parishes with rates greater than 2.00. There was only one parish in the year 2001 that had a rate greater than 2.01 per 100,000.The highest rate among many of the studied diseases was not only among Caucasian, but in infants in the 0–4 year old age category and in children in the 5–9 year old age category. The largest segment of the population in Red River and Washington Parishes was found to be in the under 18 year old age group. This probably accounts for such large numbers of young residents being reported [12]. As mentioned earlier, a factor as to the possibility of why this segment of the population is represented in such high numbers is the possibility that this community also had more access to health care and more screening by health officials in which the disease would have been required to be reported This would be expected in a parish like St. Tammany in which 80.66 % of the residents were high school graduates or higher. All four studied parishes showed that greater than 50 % of its residents have a high school diploma or above. The parish having the lowest percent of high school graduates was Terrebonne Parish with 61.07 %. There is no correlation seen among the parishes when comparing the educational status of the residents to the percent of families living below the poverty level in 1999 [15]. In regards to the high number of infants and children being infected with the studied diseases, the risk from direct contact with fecal material at farms and petting zoos is also recognized as an important factor [17]. According to the Surveillance Summaries of the Morbidity and Mortality Weekly Reports between the years of 1997 and 1998, nine persons became ill from Shigella sonnei in Massachusetts. This outbreak was associated with a wading pool that included a sprinkler fountain and was used by many diaper-aged children.According to the 2000 census, Terrebonne Parish reported that 4.53 % of it’s business industry is considered as being a selected industry such as agriculture, forestry, fishing and hunting. The selected industry percent was highest for Terrebonne Parish than the other three studied parishes of Red River, St. Tammany and Washington. Washington Parish had the third highest percent of 2.1 when compared to the other three parishes of Red River, St. Tammany and Terrebonne. In terms of the water area per square mile, Washington Parish had the smallest number of 6.40 square miles as compared to the other three studied parishes. Terrebonne Parish is also noted as having the largest water area and land area in square miles [15]. Many of the water areas of these parishes include estuaries. Estuaries can become contaminated with fecal coliform bacterial pollution as a result of rainfall runoff from urbanized areas [18]. Water plays an important role in the transmission of campylobacteriosis [19].In 1997, the amounts per year of nitrogen and phosphorous in animal waste for Terrebonne Parish were found to be 390,000 and 100,000 pounds, respectively. Of the reported amount, 210,000 pounds of nitrogen per year were reported as being lost to the atmosphere [14]. Washington Parish ranked second in the state among the sixty-four parishes with 490,000 tons of cattle waste and ranked thirteenth with regards to poultry waste generated, with only 64 tons. Between the years of 1999 and 2001, Washington Parish had one of the highest reported rates for at least one of the bacterial diseases such as camphlobacteriosis, E. coli, salmonellosis and shigellosis [12]. In 1997, a total of 5,300,000 pounds per year of nitrogen in animal waste and a lower amount of 1,200,000 pounds per year of phosphorous in animal waste were reported for Washington Parish. Compared to the other parishes in the state, Washington Parish ranked fourth in the amount of phosphorus reported in animal waste and third in the amount of nitrogen reported in animal waste. Of the reported amount of nitrogen, 1,600,000 pounds per year was lost to the atmosphere [14]. There was no correlation found to exist between the amount of animal waste stored within the parishes and the incidence rates of the studied diseases. It must also be noted that the amounts of animal waste totals reported from the Environmental Defence and Get Active Software seem relatively high based upon the number of animals used in the calculation.Survey of literature information regarding diseases such as campylobacteriosis, E. coli O157:H7 infections, salmonellosis and shigellosis within the state of Louisiana indicated that these diseases have animal origins. It is possible that some of the cases were related to animal waste, but there is no clear indication that all cases have this origin. Although the four parishes surveyed had large amount of animal waste generated each year, statistics does not show a correlation between this and the studied diseases.State of Louisiana campylobacter average annual cases by seasonal distribution for years 1988–2001(LDHH 2001).E. coli O157:H7 average incidence rates by race and age for the years 1996–2001 in Louisiana (LDHH 2001).Average annual cases of E. coli O157:H7 by seasonal distribution for the years 1996–2001 in Louisiana (LDHH 2001).Annual percentages of total deaths by age caused by salmonella in Louisiana during the years of 1987–2001(LDHH 2001).Highest infection rates per number of cases for campylobacteriosis, salmonellosis, and shigellosis in Louisiana parishes (LDHH 2001).Highest number of reported cases of campylobacteriosis, salmonellosis, and shigellosis in Louisiana parishes in 1999 (LDHH 1999).Amount of cattle and poultry waste for Louisiana parishes (Environmental Defense and GetActive Software 2003).Education, socioeconomic status and demographics for Red River, St. Tammany, Terrebonne and Washington parishes in Louisiana for the year 2000.Distribution of campylobacter in various parishes for the year 2000.Number of cases per 100,000 peopleDistribution of campylobacter in various parishes for the year 2001.Number of cases per 100,000 peopleDistribution of salmonella in various parishes for the year 2000.Number of cases per 100,000 peopleDistribution of salmonella in various parishes for the year 2001.Number of cases per 100,000 peopleDistribution of shigella in various parishes for the year 2000.Number of cases per 100,000 peopleDistribution of shigella in various parishes for the year 2001.Number of cases per 100,000 peopleThis research was financially supported in part by a grant from the National Institutes of Health (Grant No. 1G12RR13459), through the RCMI-Center for Environmental Health at Jackson State University, in part by a grant from the U.S. Department of Education (Grant No. PO31B990006), through the Title III Graduate Education Program at Jackson State University and in part by Louisiana State University Agricultural Experiment Station, Hill Farm Research Station.
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+ The objective of this research was to compare the chemical/physical parameters and bacterial qualities of selected surface water streams in Louisiana, including a natural stream (control) and an animal waste related stream. Samples were collected and analyzed for fecal coliforms. Fecal coliforms isolated from these samples were identified to the species level. Chemical analysis was performed following standard test protocols (LaMotte 2002). An analysis of biological oxygen demand (BOD), chemical oxygen demand (COD), total organic carbon (TOC), total dissolved solids (TDS), conductivity, pH, temperature, ammonia nitrogen, nitrate nitrogen, iron, copper, phosphate, potassium, sulfate, turbidity, zinc and bacterial levels was performed following standard test protocols as presented in Standard Methods for the Examination of Water and Wastewater [9]. Results of the comparisons of the various surface water streams showed that phosphate levels, according to Mitchell and Stapp, were considered good for Lake Claiborne (control) and Bayou Dorcheat. The levels were found to be .001 mg/L and .007 mg/L respectively. Other streams associated with animal waste, had higher phosphate levels of 2.07 mg/L and 2.78 mg/L, respectively. Conductivity and total dissolved solids (TDS) levels were the lowest in Lake Claiborne and highest in the Hill Farm Research Station stream. It can be concluded from the data that some bacterial levels and various nutrient levels can be affected in water resources due to non-point source pollution. Many of these levels will remain unaffected.It is predicted that by the turn of the next century, water shortages will become more widespread. Predictions from the Water Resource Council allude to increases in the consumption of fresh water by almost 27%. The problem is even worse if groundwater is considered separately from other water sources. The Global 2000 study concluded that pollution, gross national products (GNP), and resource projections all imply rapidly increasing demands for fresh water [1].With the demand for fresh water growing, agriculture is under increasing pressure to ensure that its practices do not contribute to the decline of water quality both nationally and in the state of Louisiana. According to the U.S. EPA Louisiana 1996 305(b) report, fecal coliform bacteria are the most common pollutant in rivers and streams. It has been noted that of the waterways in Louisiana that were surveyed, 37% of the river miles, 31% of lakes, and 23% of estuarine water in Louisiana had some level of contamination [2]. Possible sources of elevated coliform counts include sewage discharges from municipal treatment plants and septic tanks, storm water overflows, and runoff from pastures and range lands.Many Louisiana waterways continue to have elevated levels of unhealthy bacteria due to human and natural contaminants. Studies conducted during the summer of 2001 by the Department of Health and Hospitals on the Tangipahoa River showed a strong correlation between high water caused by rain and increased levels of bacteria [3].Research conducted in 1999 on the Ross Barnett reservoir located in central Mississippi concluded that there exists a potential public health concern with respect to microbial contamination of the water. In most cases in this study, it was shown that the bacterial counts exceeded both the federal and state guidelines for minimizing the health risks associated with water-contact activities [4].Bursting and overflowing manure lagoons have spawned environmental disasters around the country, sending animal waste flowing into rivers, groundwater and coastal wetlands. In 1995, an 8-acre hog waste lagoon in North Carolina burst, spilling 25 million gallons of animal waste into the New River. The spill killed as many as 10 million fish and closed 364,000 acres of coastal wetlands to shell fishing [5].According to the water quality standards set by the EPA, E.coli is the most reliable of fecal bacterial contamination of surface waters in the U.S. An extensive epidemiological study demonstrated that E. coli concentrations are the best predictors of swimming-associated gastrointestinal illness.The EPA recommended recreational water quality standard for E. coli is based on two criteria: 1) a geometric mean of 126 organisms/100ml based on several samples collected during dry weather conditions or 2) 235 organisms/100ml for a single water sample [6]. The geometric mean is calculated by the equation: geometric mean of y=nth root of y1 * y2 * y3…yn. If either criterion is exceeded, the site is not in compliance with water quality standards and not recommended for swimming. The current EPA water quality standard for E. coli corresponds to approximately 8 gastrointestinal illnesses per 1000 swimmers [7].This study was designed to compare the chemical/physical parameters and bacterial qualities of selected surface water streams in Louisiana, including a natural stream (control) and an animal waste related stream.A portable pH/EC/TDS/Temperature meter was purchased from A. Daigger and Company, Inc. (Vernon Hills, IL). The SMART 2 Colorimeter, 25 mm test vials and reagents for ammonia nitrogen, copper, Phosphate, potassium, sulfate and zinc tests were also purchased from A. Daigger and Company, Inc. (Vernon Hills, IL). Other materials used such as COD Standard Range Mercury Free Tubes and a COD reactor, 110v were purchased from A. Daigger and Company, Inc. (Vernon Hills, IL) as well. A COD heater block was purchased from Bioscience, Inc. (Bethlehem, PA).A comparison of the data collected from Lake Claiborne to data collected from the Bayou Dorcheat, Hill Farm Research Station stream and Ray Pond was done. According to the Department of Environmental Quality, Lake Claiborne has the designated uses of primary contact recreation, secondary contact recreation, propagation of fish and wildlife and as a drinking water supply. It is because of these designated uses that Lake Claiborne was selected to serve as a control in the experiments. Bayou Dorcheat has been designated by DEQ as having uses such as primary contact recreation, secondary contact recreation, propagation of fish and wildlife, agriculture and as being outstanding natural resource water [8]. The Hill Farm Research Station stream was selected for this portion of the study because of its location and affiliation with cattle waste. The Hill Farm stream is located on LSU Ag Center property and is surrounded by cattle. The possibility of the influence of any other type of animal waste involvement or the influence of pollution from other areas is low. Ray Pond was selected for its association with broiler litter. The property surrounding Ray Pond is the site of a commercial broiler/egg company. The levels of nitrate nitrogen, ammonia nitrogen, phosphate, potassium, copper, zinc, pH, temperature, conductivity, turbidity, BOD, COD, TOC and the type and numbers of bacteria in these water resources were compared. The accuracy of all equipment was determined by using known standards. All test conducted were referenced from the SMART 2 Colorimeter manual [9] as well as the 20th edition of the Standard Methods for the Examination of Water and Wastewater [10].All liquid samples were collected with a volume of not less than 100ml. A space of at least 2.5 cm was left in the bottle to facilitate mixing by shaking. The containers used were according to the 20th edition of Standard Methods for the Examination of Water and Wastewater. Samples were collected in nonreactive glass or plastic bottles that had been cleansed and rinsed carefully, given a final rinse with distilled water, and sterilized [10]. Containers were lowered to a depth of not greater than 2ft below the surface to fill. The samples were placed immediately on ice in order to have a temperature of less than 10°C during a maximum transport time of 6h [10]. All sample collections were conducted during the summer months. Five duplicate samples from each site were collected once weekly for a period of ten weeks. Samples from Hill Farm Research Station were collected for a period of twelve months. Samples from Lake Claiborne (control) were taken from a site located at the bank of the spillway to avoid the proximity of any septic systems. This area of Lake Claiborne is static. Bayou Dorcheat samples was collected at a bank at least one mile away from community housing to avoid the possibility of septic systems in the area. The Bayou Dorcheat area has public sewage. The sample areas for Hill Farm Research Station stream and Ray Pond were located near the bank. The area of sample collection used for Hill Farm Research Station and Bayou Dorcheat had a minimal flow rate. The sample collection area of Ray Pond was static.Prior to placing the sample on ice, the temperature was taken by placing the probe of the portable pH/EC/TDS/Temperature meter into the collected sample.The pH was measured by using the conductivity/pH/TDS meter. The pH selection was chosen on the instrument and the probe was then inserted into the water sample. The appropriate pH reading was taken.Ten ml of water sample were vortexed and filtered onto a membrane filter using a sterile filtration unit. The approved technique used was from Clesceri et al. [10].After filtration, forceps were used to place the membrane filter on an MFC agar plate. The plate was then incubated in an incubator at 45 °C for 24 hours. The plates were then checked for bacteria colony growth.The API 20E System was used in conjunction with the API Profile Recognition System (bioMerieux, Inc., Hazelwood, MO) so that members of the family Enterobacteriaceae and other Gram-negative bacteria could be accurately identified.The API 20 E strip consists of 20 microtubes containing dehydrated substrates. These tests were inoculated with the bacterial sample suspension. Each sample was incubated for 18–24 hrs. at 35–37°C. This system is a standardized, miniaturized version of conventional procedures for the identification of Enterobacteriaceae and other Gram-negative bacteria.Nitrate nitrogen was measured by using Waterworks Test Strips (Thomas Scientific Swedesboro, NJ). The test strip was placed into the water sample. The color change on the strip was compared to the chart provided with the strips.The colorimeter test vial was filled to the 10 ml line with the sample. The test vial was then placed into the colorimeter and scanned as a blank. The test vial was then removed and 8 drops of ammonia nitrogen reagent #1 was added and mixed. The mixture was allowed to sit for 1 minute. One ml of ammonia nitrogen reagent # 2 was added. After waiting 5 minutes, the test vial was inserted again and a reading was taken.The test vial was filled with 10 ml of the sample water. The test vial was inserted into colorimeter and scanned as a blank. Two ml of VM Phosphate Reagent was added and mixed well. After waiting for a period of 5 minutes, a reading was obtained using the colorimeter.Ten ml of the sample water was added to the test vial and 5 drops of copper 1 reagent was added. The sample mixture was then mixed well. The solution turned yellow which indicated that copper is present. The test vial was then placed into the colorimeter and a reading was taken.A dilute zinc indicator solution was prepared by adding five ml of a zinc indicator solution to 17.8 ml of methyl alcohol. Afterwards, the indicator solution and methyl alcohol were mixed well. A colorimeter test vial was filled with the water sample to the 10 ml line. The test vial was then inserted into the colorimeter and scanned as a blank. One tenth of a gram of sodium ascorbate powder was added along with 0.5 g zinc buffer powder. The mixture was capped and shaken vigorously for 1 minute. Three drops of sodium cyanide, 10% was added to the mixture and mixed well. One ml of dilute zinc indicator solution was then added and mixed. Four drops of formaldehyde solution, 37% was added and was mixed by inverting 15 times. The test vial was then inserted into the colorimeter and scanned for a reading.The colorimeter tube was filled to the 10 ml line with the sample water and was then inserted into colorimeter and scanned as a blank. One tenth of a gram of the sulfate reagent was added. The test vial was then capped and shaken until powder dissolved.The mixture was allowed to sit for 5 minutes. After mixing the tube again, it was then inserted into the colorimeter and scanned for a reading.One hundreth of 5 g of Tetraphenylboron Powder was added to the test vial containing 10 ml of the sample water. The vial was then capped and shaken vigorously until all of the powder had dissolved. The mixture was allowed to stand undisturbed for 5 minutes. The same procedure was followed to make a blank. Prior to taking the measurement, the vials were mixed again to suspend any settled precipitate. The vials were then inserted into the colorimeter and scanned for a reading.The 25 mm test vial was with 10 ml of the sample water. Two drops of manganese sulfate solution was then added as well 2 drops of alkaline potassium iodide azide. The mixture was mixed by inverting several times. The resulting precipitate was allowed to settle. 8 drops of sulfuric acid 1:1 was added and was gently mixed until the precipitate has dissolved. The test vial was then inserted into the SMART 2 Colorimeter and a reading was taken.The COD heater block was preheated to 150 ± 2°C. Two ml of sample water was added and was mixed well. All of the previous steps were repeated in order to make a blank. Both vials were then placed in the preheated COD block heater and maintained at that temperature for 2 hours. The vials were allowed to cool for 20 minutes. After cooling, the test vials containing the blank and the sample were inserted in the colorimeter to take a reading.The conductivity and the TDS were both measured by using a portable Conductivity/pH/TDS Meter. The probe for the instrument was placed directly into the water sample. The type of measurement taken was then selected on the instrument panel. The correct measurement was taken by reading directly from the instrument panel.Ten ml of the water sample was placed into the colorimeter tube and scanned using the SMART 2 colorimeter for a measurement.The testing of the samples for BOD and TOC was conducted by Gulf States Environmental Laboratories, Inc. (Shreveport, LA). The protocol followed by Gulf States Environmental Laboratories, Inc. was in according to the 20th edition of Standard Methods for the Examination of Water and Wastewater [10].Descriptive statistics were applied to determine the mean values of all physical/chemical and bacterial parameters evaluated. Standard deviations were computed as measures of variance. Statistical analysis was performed using GraphPad InStat program version 3.00 for Windows 95, GraphPad Software, San Diego, California. The Dunn’s Multiple Comparisons test was applied to determine significant differences in mean values of each studied parameter among the sampling sites. The level of significance was considered at p ≤ 0.05.A comparative analysis involving the use of natural water sources that may or may not be associated with animal waste was conducted. Each site was selected based upon its designated use and/or association with animal waste. The sample sites chosen were Bayou Dorcheat, Hill Farm Research Station stream, Lake Claiborne, and Ray Pond. The mean values of chemical oxygen demand-COD (mg/L) for each sample site were found to be 44.07 ± 0.46, 21.13 ± 0.35, 23.73 ± 0.70, and 48.86 ± 0.35 mg/L, respectively. The mean values of biochemical oxygen demand -BOD (mg/L) were determined to be 6.63 ± 0.14, 5.56 ± 0.01, 3.96 ± 0.01, and 16.63 ± 0.22 mg/L, respectively in Bayou Dorcheat, Hill Farm Research Station stream, Lake Claiborne, and Ray Pond. Total organic carbon- TOC levels were found to be 11.51 ± 0.005, 5.02 ± 0.004, 7.29 ± 0.004, and 6.53 ± 0.004 mg/L in Bayou Dorcheat, Hill Farm Research Station stream, Lake Claiborne, and Ray Pond, respectively. Figures 1, 2, and 3 show the mean levels of COD, BOD, and TOC for each of the sample sites. It was determined that variations found among the BOD means for each sample site were highly significant (p<0.0001).The water flow rate for each of the sites was estimated at being minimal for Hill Research Station stream and Bayou Dorcheat. A static flow rate was noticed for Ray Pond and Lake Claiborne. Total dissolved solids TDS were evaluated at each of the sample sites. As shown in Figure 4, the lowest TDS was associated with Lake Claiborne. This number was noted to be 13.13 ± .01mg/L while the highest TDS was associated with Hill Farm Research Station at 49.07± .8.50 mg/L. The TDS levels for Bayou Dorcheat and Ray Pond 30.80 ± 2.20mg/L and was 21.27 ± 3.20mg/L, respectively. Comparisons of variations between the mean TDS levels for the sampling sites Hill Farm and Lake Claiborne, Hill Farm and Ray Pond and Bayou Dorcheat and Lake Claiborne show highly signicant differences (p<0.001). There were no significant differences in TDS concentrations (p>0.05) between the sampling sites of Bayou Dorcheat and Hill Farm, Bayou Dorcheat and Ray Pond, and Lake Claiborne and Ray Pond.Visual observations determined that Lake Claiborne was the clearest of the water sample sites. Further testing of the turbidity determined the levels to be similar to the visual observations in that Lake Claiborne also had the lowest turbidity level of 15.0 ± 2.1 NTUs. As seen in Figure 5, the other sampling sites had turbidity levels that were very similar to one another but were greater than that found in Lake Claiborne. Bayou Dorcheat had the highest turbidity level among all of the sampling sites at 40.70 ± 2.20 NTUs. Hill Farm had a level of 30.70 ± 5.30 NTUs and Ray Pond had a turbidity level of 35.47± 7.80 NTUs. Comparisons among mean levels of turbidity revealed highly significant (p<0.0001) variations among sampling sites.The pH was measured for each of the sampling sites. The pH values recorded were 6.46 ± 0.18 for Bayou Dorcheat, 6.40 ± 0.30 for Hill Farm, 6.70 ± 0.15 for Lake Claiborne and 7.90 ± 0.58 for Ray Pond. The pH values for the sampling sites are shown in Figure 6.Conductivity levels as well as the temperature were checked at all sampling sites. These levels are presented in Figures 7 and 8. Lake Claiborne had the lowest conductivity level (27 ± 1.07μS/cm) of all of the sampling sites. The highest conductivity level (92.70 ± 14.09μS/cm) was detected at the Hill Farm Research Station location. Bayou Dorcheat and Ray Pond were found to have conductivity levels of 61.60 ± 4.09μS/cm and 43.87 ± 12.17μS/cm, respectively.The temperature levels were similar for all the sampling sites. These temperatures were found to be 24.88 ±1.40°C, 20.97 ±2.50°C, 27.31 ± 0.62°C, and 28.22 ± 0.92°C, respectively for Bayou Dorcheat, Hill Farm, Lake Claiborne, and Ray Pond.The chemical parameters such as phosphate and sulfate were checked and it was noted that Lake Claiborne also had the lowest phosphate level of all the sample sites. The phosphate level was observed at 0.001 ± .003ppm for Lake Claiborne while Bayou Dorcheat had a measurement of .007 ± .004ppm. Hill Farm and Ray Pond had phosphate levels of 2.07±1.2ppm and 2.780 ± 0.770ppm, respectively. Figure 9 shows the mean values (n=15) of phosphate while figure 10 shows those of sulfate. The sulfate levels detected for Lake Claiborne were found to be the lowest value among all of the sampling sites of 5.67 ± 0.00ppm. Hill Farm was found to have the highest sulfate level at 6.20± 1.70 ppm. Bayou Dorcheat and Ray Pond had sulfate levels of 5.73 ± .70ppm and 5.93 ± 0.90ppm respectively.Samples from the Hill Farm stream were collected once per month for a period of one year. The results showed that only ten months of the sampled year had E. coli colonies found in the water samples. The Hill Farm stream had E. coli colony numbers ranging in numbers from 90 CFUs/100 ml to 1,800 CFUs/100 ml. Only two months showed colony numbers higher than 1,000 CFUs/100 ml for the Hill Farm stream.Comparisons of bacterial parameters between Lake Claiborne (control) and Hill Farm Research Station showed that the Hill Farm Research Station stream had higher counts of bacterial colonies present than that of Lake Claiborne. Analysis of ten weekly samples from Lake Claiborne only had two in which fecal coliform were present. An analysis over a ten month period on the Hill Farm Research Station stream revealed fecal coliform present during each sampling. Of the two positive analyses for Lake Claiborne, the bacterial counts were 100 CFUs/100 ml of sample and 200 CFUs/100 ml of sample. All colonies were identified as being E. coli. As mentioned earlier, the Hill Farm stream had E. coli colony numbers ranging in numbers from 90 CFUs/100 ml to 1,800 CFUs/100 ml. Six of the monthly data of the number of colonies found in the Hill Farm stream were higher than the colony counts of Lake Claiborne, three of the monthly data were equivalent and only one monthly data was lower than the colony numbers found in Lake Claiborne. The mean levels of bacteria in Lake Claiborne and Hill Farm Research station stream differ significantly (p < 0.0199). Figure 11 shows the mean numbers of the bacterial colonies found at each sampling site. The mean values of E. coli CFUs/100 ml for Lake Claiborne, Bayou Dorcheat, Hill Farm Research Station Stream and Ray Pond were 30.00 ± 67.50 CFUs/100 ml, 25.40 ± 22.39 CFUs/100 ml, 482.00 ± 555.43 CFUs/100 ml and 249.60 ± 181.64 CFUs/100 ml, respectively. Differences in bacterial levels among sampling sites were highly significant (p < 0.0001).According to the Department of Environmental Quality, Lake Claiborne has the designated uses of primary contact recreation, secondary contact recreation, propagation of fish and wildlife and as a drinking water supply. It is because of these designated uses that Lake Claiborne was selected to serve as a control in the experiments. Bayou Dorcheat has been designated by DEQ as having uses such as primary contact recreation, secondary contact recreation, propagation of fish and wildlife, agriculture and as being outstanding natural resource water [8]. The Hill Farm Research Station stream was selected for this portion of the study because of its affiliation with cattle waste and Ray Pond was selected for its association with broiler litter.All of the pH levels received for the natural water sources were found to be within a normal range and ranked as either good or excellent. The pH for natural water is usually between 6.5 and 8.5 although variations are known to occur. At extremely high or low pH values such as greater than 9.6 or less than 4.5, the water becomes unsuitable for most organisms [11]. The pH for Lake Claiborne was considered excellent, while Bayou Dorcheat, Hill Farm and Ray Pond had a pH that was considered good and within the normal range.The total dissolved solids -TDS level of all of the sampling sites were within the requirements of DEQ of not exceeding 200 mg/L. The conductivity levels were parallel to the TDS results in that both data showed Lake Claiborne as being the lowest and next in increasing order was Ray Pond and Bayou Dorcheat with Hill Farm being the highest. All of the temperatures received for each site were within the requirements of not exceeding 32°C [8]. A comparison of the data from the chemical oxygen demand- COD and the total organic carbon-TOC measurements showed that the Hill Farm Research Station stream had the lowest COD and TOC as compared to the other three sampling sites. Bayou Dorcheat was found to have the highest TOC while Ray Pond had the highest COD measurements. The biochemical oxygen demand -BOD was lowest for Lake Claiborne. At 3.96mg/L, this is ranked as good. Bayou Dorcheat and Hill Farm sampling sites had BOD measurements of 6.70mg/L and 5.56mg/L, respectively. Both of these readings are ranked as being fair. Ray Pond had a BOD measurement of 16.70mg/L and is ranked as being poor. It seems that the only one of the sampling sites having an acceptable level is Lake Claiborne. This lake is also the only sampling site having drinking water as one of its designated uses. BOD is a measure of the quantity of oxygen used by micro organisms in the aerobic oxidation of organic matter. The addition of nutrients to a water source can be a major force in having a high BOD. In waters of high BOD, a low diversity of aquatic organisms will replace the ecologically stable and complex relationships present in waters containing a high diversity of organisms [11].Turbidity levels for all sampling sites were rated as good except in the case of Bayou Dorcheat. The rating for Bayou Dorcheat in turbidity was fair. It is known that at high levels of turbidity, water loses its ability to support a diversity of aquatic organisms. Water becomes warmer as suspended particles absorb heat from sunlight, causing oxygen levels to fall [11].A comparison of the sampling site sulfate levels showed that none exceeded 15 mg/L [8]. The only sampling site utilized as a source of drinking water, Lake Claiborne, had the lowest sulfate value. The phosphate values showed that Bayou Dorcheat and Lake Claiborne were within the range of excellent. Hill Farm and Ray Pond which are both closely associated with cattle and poultry respectively, were within the range of good.Lake Claiborne was found to have fewer bacteria colonies than any of the other sampling sites. Over the course of ten months, only two were positive for coliform bacteria. The other samples were negative. The two positive months had colony numbers that were 100 CFUs/100ml and 200 CFUs/100ml of sample water collected. For surface water used as a primary contact recreation, the fecal coliform content shall not exceed a log mean of 200 CFUs/100ml [8]. Lake Claiborne did not exceed the bacteria criteria. Bayou Dorcheat reported a small number of coliform colonies for each of the ten months tested, but all of the positive months were also below the criteria established by DEQ. Hill Farm Research Station stream and Ray Pond both had a mean number of coliform colonies that exceeded the criteria. The mean numbers of coliform colonies were 482CFUs/100 ml and 249 CFUs/100 ml, respectively. The Hill Farm Research Station stream had E. coli numbers that exceeded the DEQ criteria during six of the ten months that were analyzed.It was proven by this experiment that natural water sources in close proximity of cattle and poultry farms will experience some negative effects on water quality over time. Amount of animal waste in regards to the volume of the water source can impact the degree of the negative impact. This was seen in the data collected from Hill Farm Research Station and Ray Pond. This negative effect is only obvious when the comparison was to Lake Claiborne which was the only natural water source analyzed that was used as a source of drinking water. This site served as the control within the experiments.It can be concluded from the data that some bacterial levels and various nutrient levels can be affected in water resources due to non-point source pollution. Many of these levels will remain unaffected.Mean values of chemical oxygen demand COD (mg/L) at each sample site (n=15).Mean values of biochemical oxygen demand BOD (mg/L) at each sample site (n=15).Mean values of total organic carbon TOC (mg/L) at each sample site (n=15).Mean values of total dissolved solids TDS (mg/L) at each of the sampling sites. Differences in mean levels at sampling sites are highly significant (p<0.001), n=15.Mean values of turbidity (NTU), (n=15) at each sampling site.Mean pH values (n=15) for each sampling site.Mean values of conductivity (μS/cm) at each sampling site. Differences in mean values among the sampling sites are highly significant (p<0.0001).Mean values (n=15) of temperatures (°C) at each sampling site.Mean (n=15) values of phosphate (ppm) levels at each sampling site.Mean (n=15) values of sulfate (ppm) levels at each sampling siteMean values of E. coli colonies/100 ml for each sampling site.This research was financially supported in part by a grant from the National Institutes of Health (Grant No. 1G12RR13459), through the RCMI-Center for Environmental Health at Jackson State University, in part by a grant from the U.S. Department of Education (Grant No. PO31B990006), through the Title III Graduate Education Program at Jackson State University and in part by Louisiana State University Agricultural Experiment Station, Hill Farm Research Station.
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+ The major concern for the halogenated compounds is their widespread distribution, in addition to occupational exposures. Several chlorinated alkanes and alkenes were found to induce toxic effects. In this study, we investigated the genotoxic potential of 1,1-dichloroethane in the bone marrow cells obtained from Swiss-Webster mice, using chromosomal aberrations (CA), mitotic index (MI), and micronuclei (MN) formation as toxicological endpoints. Five groups of three male mice each, weighing an average of 24 ± 2 g, were injected intraperitoneally, once with doses of 100, 200, 300, 400, 500 mg/kg body weight (BW) of 1,1-dichloroethane dissolved in ethanol. A control group was also made of three animals injected with ethanol (1%) without the chemical. All animals were sacrificed 24 hours after the treatment. Chromosome and micronuclei preparations were obtained from bone marrow cells following standard protocols. Chromatid and chromosome aberrations were investigated in 100 metaphase cells per animal and percent micronuclei frequencies were investigated in 1,000 metaphase cells per animal. 1,1-dichloroethane exposures significantly increased the number of chromosomal aberrations and the frequency of micronucleated cells in the bone marrow cells of Swiss-Webster mice. Percent chromosomal aberrations of 2.67 ± 0.577, 7.66 ± 2.89, 8.33 ± 2.08, 14.67 ± 2.51, 20.3 ± 3.21, 28 ± 3.61; mitotic index of 9.4%, 7.9%, 6.2%, 4.3%, 3.0%, 2.6% and micronuclei frequencies of 3.33 ± 0.7, 7.33 ± 0.9, 8.00 ± 1.0, 11.67 ± 1.2, 15.33 ± 0.7, 18.00 ± 1.7 were recorded for the control, 100, 200, 300, 400, and 500 mg/kg BW respectively; indicating a gradual increase in number of chromosomal aberrations and micronuclei formation, with increasing dose of 1,1,-dichloroethane. Our results indicate that 1,1-dichloroethane has a genotoxic potential as measured by the bone marrow CA and MN tests in Swiss-Webster mice.1,1-Dichloroethane (DCE) is a short–chain, chlorinated aliphatic hydrocarbon (i.e., halocarbon). It is used primarily as an intermediate in the synthesis of other halocarbons and high vaccum rubber. DCE is utilized to a limited extent as a degreaser, cleaning agent, and finish remover [1]. Environmental releases occur primarily by volatilization, but 1,1-dichloroethane can also be discharged into surface waters and soils. DCE is among the widely used chlorinated hydrocarbon and is currently assigned the classification of C (Possible Human Carcinogen) by the U.S. Environmental Protection Agency [2]. The major concerns for this group of compounds are their widespread distribution, in addition to occupational exposure [3]. Several chlorinated alkanes and alkenes have been found to induce genotoxic effects [4, 5], and also caused miscarriages in women [6]. Judging from the very limited information available it appears that the central nervous system (CNS), kidneys, and liver are the most likely target organs for DCE, causing depression, nausea, dizziness etc [7, 8]. The main objective of this study was to determine the genotoxic effect of 1,1-dichloroethane in the bone marrow cells of Swiss-Webster mice using chromosomal aberrations, mitotic index and micronuclei formation as the toxicological endpoints. These endpoints were selected based on the fact that the bone marrow assay for detecting chromosomal aberrations and micronuclei formation is very sensitive, faster, inexpensive and easier to run. It can also be used to evaluate genotoxic hazards [9].1,1-dichloroethane Cat No: 48512, Potassium chloride solution (0.075 M ) Lot No: 72K2447, Giemsa stain stock solution (0.4%) Lot No: 22k8416, May-Grunwald stain Cat No: MG-500 were purchased from Sigma-Aldrich (St. Louis, MO, USA). Methanol Cat No: A 452-4, Glacial acetic acid Cat No: A38-500, Super frost microscope slides Cat No:12-544-15 was purchased from Fischer-Scientific (Houston, TX, USA). Hanks Balanced Salt Solution Cat No: 14025-092 and Fetal Bovine Serum (FBS) Cat No: 16000-044 was obtained from GIBCO (Grand Island, NY, USA).Healthy adult male Swiss-Webster mice (6–8 weeks of age, with average bodyweight of 24 ± 2 g were used in this study. They were purchased from Simonsen Breeding Laboratories in Gilrey, California. The animals were randomly selected and housed in polycarbonate cages (three mice per cage) with steel wire tops and corn-cob bedding. They were maintained in a controlled atmosphere with a 12h: 12h dark/light cycle, a temperature of 22 ± 2° C and 50–70% humidity with free access to pelleted feed and fresh tap water. The animals were supplied with dry food pellets commercially available from PMI Feeds Inc (St. Louis, Missouri). They were allowed to acclimate for 2 weeks prior to the treatment.Groups of three mice each were treated with five different 1,1-dichloroethane dose levels 100, 200, 300, 400, 500 mg/kg BW. 1,1-dichloroethane was dissolved in ethanol (as required) and intraperitoneally administered to animals at the doses of 0, 100, 200, 300, 400, 500 mg/kg BW. Ethanol (1%) was used as the solvent control.Cytogenetic analysis was performed on bone marrow cells according to the recommendations of Preston et al (9) with slight modifications. Experimental animals were injected (i.p.) with colchicine (4mg/kg) 1.5 h prior to sacrifice. Animals were sacrificed by cervical dislocation 24 hours after the exposure; both femora from each animal were removed and freed of muscle. The muscle free femora bone was cut at the proximal end and a syringe needle with about 2 ml of mixture of fetal calf serum and Hanks balanced salt solution (3:1 ratio), was inserted at the distal end and bone marrow cells were flushed in to a centrifuge tube, and gently aspirated for uniform spread [10]. The cells were centrifuged at 800 rpm for 5min and the supernatant was discarded. The collected bone-marrow cells were incubated in KCL [0.075M] at 37° C for 25 min. The cells were then centrifuged at 2000-x g for 10 min, fixed in aceto-methanol (acetic acid: methanol, 1:3 v/v). Centrifugation and fixation were repeated five times at an interval of 20 min. The cells were re-suspended in a small volume of the fixative, dropped onto chilled slides, flame-dried and stained with freshly prepared 2% Giemsa stain for 3–5 min and were washed in distilled water to remove excess stain. 300 metaphase plates containing 40 ± 2 chromosomes were examined per animal to score different types of aberrations.The mitotic index was used to determine the rate of cell division. The slides prepared for the assessment of chromosomal aberrations were also used for calculating the mitotic index. Randomly selected views on the slides were monitored to determine the number of dividing cells (metaphase stage) and the total number of cells. At least 1000 cells were examined in each preparation. The mitotic index was calculated as the ratio of the number of dividing cells to the total number of cells, multiplied by 100.Mice were sacrificed by cervical dislocation 24h after the treatment. The frequency of micro-nucleated cells in femoral bone marrow was evaluated according to the procedure of Schmid [11], with slight modifications as reported by Agarwal and Chauhan [12]. The bone marrow was flushed out from both femora using 2 ml of Fetal Calf Serum and Hanks Balanced Salt Solution (3:1) and centrifuged at 2000-x g for 10 min. The supernatant was discarded. Evenly spread bone marrow smears were first stained in 5% May-Grumwald solution for 2.5 minutes, and then immersed in 20% Giemsa solution for 20 minutes. Subsequently the slides were placed in Sorensen’s buffer (pH 6.7 and pH 6.8) for 10 seconds, rinsed in distilled water, and then dried overnight. The following day, slides were sequentially immersed in acetone and xylene for 5 seconds. After the slides have dried, the frequency of cells with micronuclei was determined using a microscope. A total of 3000 cells were counted from each treatment group (1000 cells per animal).Data obtained in this study were analyzed using two-way ANOVA and Tukey test. All values were reported as Means ± SEMs. For all the experiments, the significance level was set at p≤ 0.05.1,1-Dichloroethane (DCE) induced a statistically significant increase in chromosome aberrations in metaphase bone marrow cells. Table. 1 shows the mean frequency of cells with aberrations (as percentages) calculated for each dose with 3 animals/dose. The most frequent types of aberrations were gaps, breaks and fragments. Chromatid-type aberrations were detected at high frequencies. Relatively higher frequencies of gaps were observed for all the doses tested. A quantitative assessment of the distribution of breaks and gaps revealed that the distal regions of the chromosomes were more vulnerable to the effects of 1,1-Dichloroethane. The frequency of chromosomal aberrations also increased with increasing doses of 1,1-dichloroethane (Figure 1). The mean percentages of the induced chromosomal aberrations were 2.67 ± 0.577 %, 7.66 ± 2.89 %, 8.33 ± 2.08 %, 14.67 ± 2.51%, 20.3 ± 3.21%, 28 ± 3.61 at 1,1-dichloroethane doses of 0, 100, 200, 300, 400 and 500 mg/kg BW respectively.The mitotic index was used to determine the rate of cell division. The slides prepared for the assessment of chromosomal aberrations were used for calculating the mitotic index. Mitotic Index depression was observed in the bone marrow cells of mice as compared with the control, it was found to be dose-dependent (Figure 2). Mean percentages of mitotic index were 9.4%, 7.9%, 6.2%, 4.3%, 3.0%, 2.6% for the doses 0, 100, 200, 300, 400 and 500 mg/Kg BW of 1,1-Dichloroethane respectively.The micronuclei frequencies in bone marrow cells after intra-peritoneal treatment with 1,1-dichloroethane are summarized in Table 2. 1,1-dichloroethane induced a dose-dependent increase in micronuclei frequency (Figure 3) and significant (p>0.05) differences from the control were observed. The mean frequency of micro-nucleated cells were 3.33 ± 07, 7.33 ± 0.9, 8.00 ± 1.0, 11.67 ± 1.2, 15.33 ± 0.7, 18.00 ± 1.7 at 1,1-Dichloroethane doses of 0, 100, 200, 300, 400, and 500 mg/kg BW respectively.Rodent animal bioassays are valuable tools for investigating the pharmacokinetics, mechanisms of action, and differential toxicity of various chemicals [3]. The data obtained from this study clearly shows that 1,1-DCE significantly increased the number of chromosomal aberrations and the formation of micronuclei as compared to the control. These results support earlier studies with halogenated compounds [13–15]. The increase in chromosome aberrations and micronuclei formation was in a dose-dependent manner. The different type of aberrations produced by the test chemical suggests their clastogenic potential. The increased chromosomal aberrations could be attributed to either an increase in induced DNA lesions or interference with their repair [16]. The chromosome gap, which represents only the loss of chromatin material [17], may be due to the damage to the protein part of the chromosome rather than the whole chromosome. The chromatid breaks represent the DNA double strand breaks that have not undergone G2 repair [18]. However, a number of factors can influence the time of appearance of chemically induced aberrations such as compound solubility, rate and distribution of biotransport, availability at the target site as influenced by time and cell permeability [19].Mitotic index depression was observed in the bone-marrow cells of mice as compared with the control. The decrease was in a dose-dependent manner. These results were in agreement with Rank & Nielsen [15] who observed a decrease in mitotic index in the Allium cepa root tips. Duma et al [20] found decreased mitotic index by the monocrotophos in rodents. Significant decrease in the mitotic index was observed with the Roundup herbicide [21].Chromosomal breaks or interference in the mitotic process results in lagging of the chromosomal material and during cell division, this leads to the formation of micronuclei [22]. Treatments with 1,1-DCE have shown a dose-dependent increase in the number of micronuclei. These results were in conjunction with the studies of Tafazoli & Kirsch-Volders [23] with chlorinated hydrocarbons in human lymphocytes; Jagetia & Jyothi [24] with Vinesine (desacetyl vinblastinamide sulphate) in mouse bone marrow cells; and Gutierrez et al [25] with 131I sodium iodide exposure in human. Several other studies have shown an increased in micronulei formation in treated cells [22, 26–29]. The induction of micronuclei by 1,1-DCE may be due to the ability of these compounds to affect DNA synthesis, which may lead to DNA strand breaks and as a result, breakage of chromosome and or loss of whole chromosome owing to the spindle failure.Our data indicate that 1,1-dichloroethane induces chromosome aberrations and the formation of micronuclei in the bone marrow cells of Swiss-Webster mice. The repression in mitotic index indicates the potential for 1,1-dichloroethane to induce growth arrest or to inhibit cell growth. These findings demonstrate that 1,1-dichloroethane has a strong clastogenic/genotoxic potential. The chromosome aberration assay and micronuclei test appear to be promising technique to assess the clastogenic/genotoxic potential of 1,1-dichloroethane and its compounds.Effect of 1,1-Dichloroethane on the frequency of chromosomal aberrations.Effect of 1,1-Dichloroethane on the percent Mitotic Index.Effect of 1,1-Dichloroethane on the percent Micronuclei induction.Frequency of chromosome aberrations in bone marrow cells of Swiss-Webster mice induced by 1,1-Dichloroethane.n: number of animals; SCA: Structural chromosomal aberrations; MI: mitotic index p< 0.05 compared to controlFrequency of micronucleated cells in mice bone marrow. Means followed by a common letter are not significantly different from each other at p≤ 0.05 (Tukey test).Micronuclei; DCE: 1,1-dichloroethane
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+ Exposure to particulate matter (PM2.5–10), including diesel exhaust particles (DEP) has been reported to induce lung injury and exacerbation of asthma and chronic obstructive pulmonary disease. Alveolar macrophages play a major role in the lung’s response to inhaled particles and therefore, are a primary target for PM2.5–10 effect. The molecular and cellular events underlying DEP-induced toxicity in the lung, however, remain unclear. To determine the effect of DEP on alveolar macrophages, RAW 264.7 cells were grown in RPMI 1640 with supplements until confluency. RAW 264.7 cultures were exposed to Hank’s buffered saline solution (vehicle), vehicle containing an NF-κB inhibitor, BAY11-7082 (25μM with 11/2 hr pre-incubation), or vehicle containing DEP (250μg/ml) in the presence or absence of BAY11-7082 (25μM with 11/2 hr pre-incubation) for 4 hr and TNF-α release was determined by enzyme-linked immunosorbent assay and confirmed by western blots. RAW 264.7 apoptotic response was determined by DNA fragmentation assays. U937 cells treated with campothecin (4 μg/ml × 3 hr), an apoptosis-inducing agent, were used as positive control. We report that exposure to the carbonaceous core of DEP induces significant release of TNF-α in a concentration-dependent fashion (31 ± 4 pg/ml, n = 4, p = 0.08; 162 ± 23 pg/ml, n = 4, p < 0.05; 313 ± 31 pg/ml, n = 4, p < 0.05 at 25, 100, and 250 μg/ml, respectively). DEP exposure, however, failed to induce any apoptotic response in RAW 264.7 cells. Moreover, inhibition of NF-κB binding activity has resulted in DEP-induced apoptotic response in alveolar macrophages, as demonstrated by the NF-κB inhibitor, BAY11-7082 studies. The results of the present study indicate that DEP induce the release of TNF-α in alveolar macrophages, a primary target for inhaled particles effect. DEP-induced TNF-α gene expression is regulated at the transcriptional level by NF-κB. Furthermore, DEP-induced increase in NF-κB-DNA binding activity appears to protect against apoptosis.Particulate matter (PM) is released into the ambient air from the combustion of fossil products by industrial and agricultural processes, and transportation. Diesel exhaust particles generated and emitted from diesel engines are a major component of atmospheric PM. The pulmonary epithelium and resident macrophages are primary targets of inhaled particulate matter (PM2.5–10). Increased morbidity and mortality from cardiopulmonary complications have been associated with exposure to PM2.5–10. Diesel exhaust particles are of the criteria air pollutants that are implicated in inducing lung disease and injury [1]. Many studies have implicated fine particles, including DEP in airway inflammation and hyper-responsiveness [2] and in exacerbation of asthma and chronic obstructive pulmonary disease (COPD) [3–5]. Benzo[a]pyrene, a major aromatic hydrocarbon constituent coupled with DEP was shown to induce the release of inflammatory cytokines in human airway epithelial cells [1]. Suppression of human alveolar macrophage phagocytic activity has been correlated with exposure to DEP [6]. Exposure to PM2.5–10 has been reported to induce significant release of the inflammatory cytokine IL-8, a potent neutrophil chemoattractant in human monocytes [7]. Moreover, urban air particulates, including DEP have been reported to cause apoptosis of human alveolar macrophages [8]. Macrophage programmed cell death may provide a mechanistic approach to understanding lung inflammation and injury attributed to exposure to ambient air particulates [9].In vitro studies have indicated that airway epithelial cells and macrophages can bind and ingest various types of PM [10]. Increased binding activity, in addition to released inflammatory cytokines will stimulate the alveolar macrophage (AM) to release increased amounts of TNFα that may induce apoptotic cell response in macrophages through the activation of DNA-binding nuclear factors, in particular, nuclear factor-κB (NF-κB). The signaling cascade of TNF-α-induced AM apoptosis may provide a mechanistic approach to the molecular mechanisms underlying PM-induced lung inflammation and injury. Therefore, the present study aimed at determining whether: 1) DEP activate alveolar macrophages and induce TNF-α gene expression, and 2) DEP induces an apoptotic response in alveolar macrophages.Takano et al. [11] have shown that DEP (250 μg/ml) increases NF-κB-DNA binding activity in the lung of mice associated with over-expression of macrophage-activating protein-1 (MP-1) and interleukin-1β (IL-1β) genes. Hiura et al. [12] have reported that DEP induces apoptosis in RAW 264.7 macrophages. There is increased evidence that apoptosis in macrophages may be regulated by the transcription factor, NF-κB. Furthermore, Koay et al. [13] have shown that macrophage counts are associated with TNF-α release in response to LPS in the lung and are essential for initiation of NF-κB-dependent immune response.Over-expression of pro-inflammatory genes in the lung is regulated at the transcription level [14, 15]. Many pro-inflammatory genes, for example, IL-8, IL-6, tumor necrosis factor-α (TNF-α), and granulocyte macrophage-colony stimulating factor (GM-CSF) have κB sites in their 5′-flanking regions [16]. Activation of several transcription factors, notably, NF-κB results in the expression of various proinflammatory genes, for example, TNF-α, IL-6, IL-8, and GM-CSF [17]. Increased activation of NF-κB has been demonstrated in airways and in sputum macrophages of asthmatics [18]. Glucocorticoids that inhibit NF-κB activation have been shown to reduce the survival of eosinophils, a characteristic in asthma [19].We report here that DEP induce the release of TNF-α in the alveolar macrophage cell line, RAW 264.7. Over-expression of TNF-α gene was found to be regulated through increased NF-κB-DNA binding activity. Furthermore, DEP-induced activation of NF-κB appears to protect against apoptosis in cultured RAW 264.7 cells.The murine alveolar macrophage cell line, RAW 264.7, RPMI 1640, antibiotics and supplements, and fetal bovine serum (FBS) were obtained from American Type Culture Collection (Rockville, MD). Hank’s buffered saline solutions (HBSS containing 30mM Hepes) were obtained from Clonetics (San Diego, CA). TNF-α antibodies and IL-1β were obtained from Santa Cruz Biotechnology (Santa Cruz, CA). Mouse recombinant sandwich immunosorbent assay (ELISA) kit specific for tumor necrosis factor-α (TNF-α) was purchased from Pierce-Endogen (Springfield, IL). Diesel-exhaust particulates (DEP, SRM 1975, treated for lipopolysaccharide, LPS and the organic fraction extracted) with a mean diameter of 0.3 μm were purchased from the National Institute of Standards and Technology (NIST, Rockville, MD).RAW 264.7 cells were cultured in RPMI 1640 with 10% FBS optimized for macrophage growth and supplemented with penicillin (100 units/ml) and streptomycin (100 μg/ml) (Gibco BRL) until confluency. RAW 264.7 cells were grown on plastic, 6-well format plates. Before exposure to DEP, the growth medium was aspirated and Hank’s Saline Solution buffered with Hepes (30 mM) (HBSS/Hepes; pH 7.4) was added.TNF-α is an important cytokine that plays a role in the activation of macrophages and induces macrophage apoptosis. To measure the effect of DEP exposure on TNF-α release in the alveolar macrophage, confluent monolayers of the alveolar macrophage cell line, RAW 264.7, were exposed to vehicle (HBSS/30 mM Hepes) alone, or DEP at 25, 100, or 250 μg/ml for 4 hr, and TNF-α release was determined using sandwich mouse TNF-α ELISA. Briefly, the plate was coated with a TNF-α-specific antibody by biotin. The sample is added where TNF-α will bind to its antibody. Following washing the residue of unbound protein, another specific TNF-α antibody is added to which a horse radish peroxidase (HRP) enzyme is attached by streptavidin. Methyl tert-butadiene is added as the substrate for HRP. Measurements were performed in duplicates and the assay is specific for TNF-α without any interference from other cytokines. TNF-α protein levels in the culture supernatants were calculated from corresponding absorbances measured at 450 nm using a Bio-Tek EL311 autoplate reader (Bio-Tek, Winooski, VT) and standard calibration curves. The ELISA kit has a 5 pg/ml lower limit of detection. Following ELISA determinations, western blots were performed to confirm TNF-α proteins.Denatured total proteins (30μg) of cytosolic and nuclear fractions were separated on 12% denaturing polyacrylamide gels and electrotransferred to Hybond-C nitrocellulose membranes. Following standard blocking and washing procedures, the membranes were incubated with polyclonal antibodies against TNF-α, and horseradish peroxidase-conjugated anti-goat IgG secondary antibodies. Detection of proteins was performed by enhanced chemiluminescence (ECL). Blotted membranes were exposed to Kodak hyperfilm to determine protein production.To determine whether exposure to DEP induces an apoptotic response in alveolar macrophages, 3–4 × 105 of RAW 264.7 cells were grown in plastic 6-well format plates for 3 days. Monolayers were treated with vehicle (HBSS/Hepes) alone (control), vehicle containing DEP at 250μg/ml (DEP), or IL-1β at 100ng/ml (IL-1β), for 4 hr in triplicate wells. Following treatment, floating and adherent cells were pooled together and lysed with 500 μl of lysis buffer (1% SDS, 10 mM Tris, pH 7.4) for 10 min. Lysed cells were transferred to 1.5ml Eppendorf tubes and centrifuged at 1,000 × g for 10min to separate low molecular weight DNA (oligonucleosome-sized fragments derived from apoptotic cells) from high molecular DNA (derived from viable cells). A 20μl aliquot of a 1:5 dilution of the supernatant containing oligonucleosomes was used to detect apoptosis using a DNA fragmentation kit (Roche Molecular Biochemicals) according to the manufacturer’s directions. DNA fragments were separated by electrophoresis on agarose gels. U937 cells treated with campothecin, an apoptosis-inducing agent were used as positive control.To determine the role of NF-κB in regulating the expression of TNF-α gene in alveolar macrophages, or the apoptotic response, monolayers of RAW 264.7 cells were pre-treated with BAY11-7082 (25 μg/ml × 1 1/2 hr), an NF-κB inhibitor. Following pre-treatment, RAW 264.7 cultures were treated with vehicle (HBSS/Hepes) alone, or vehicle containing DEP (250 μg/ml × 4 hr) and TNF-α release or DNA fragmentation were determined as discussed earlier.Experiments were replicated three times to ensure reproducibility. Comparisons of TNF-α protein levels between control and DEP-exposed groups were performed using one-way analysis of variance (ANOVA) (20). Data are expressed as mean ± standard error of the mean (SEM).To determine the effect of DEP exposure on TNF-α release, confluent RAW 264.7 monolayers were exposed to vehicle alone (HBSS/30mM Hepes) alone, or vehicle containing 25, 100, or 250μg/ml DEP for 4 hr, and TNF-α release was measured using mouse TNF-α ELISA. As shown in figure 1, DEP exposure induced significant release of TNF-α at the concentration levels 100 and 250 μg/ml (162 ± 23 pg/ml, n = 4, p < 0.05, and 313 ± 31 pg/ml, n = 4, p < 0.05, respectively). Exposure of RAW 264.7 cultures to a concentration level of 25 μg/ml for 4 hr did not result in any significant release of TNF-α (31 ± 4pg/ml, n = 4, p = 0.08) compared to control cultures (23 ± 4pg/ml, n = 4). The results in figure 1 show that DEP exposure induces TNF-α release in a dose-dependant fashion. In additional experiments, treatment of cultures of alveolar macrophages with IL-1β at a concentration of 100 ng/ml for 4 hr, resulted in a significant release of TNF-α (139 ± 31 pg/ml, n = 4, p < 0.05) compared to control cultures (23 ± 3 pg/ml, n =4). The effect of DEP exposure at a concentration of 100 μg/ml for 4 hr is similar to the effect induced by IL-1β, as demonstrated in figure 1.Figure 3 demonstrates the effect of DEP exposure on TNF-α protein production as determined by western blotting techniques.Many studies have demonstrated that TNF-α gene expression is regulated at the transcriptional level by NF-κB. Inhibition of NF-κB by BAY-11 (25 μM × 1 ½ hr pre-incubation) which inhibits the phosphorylation of IκB, resulted in total abrogation of TNF-α release (83 ± 5% inhibition, n = 4, p < 0.05), as demonstrated in figure 4. Treatment of RAW 264.7 cultures with BAY-11 (25 μM with 1 ½ hr pretreatment) also significantly inhibited IL-1β-induced TNF-α release (77 ± 4% inhibition, n = 4, p<0.05), similar to the inhibitory effect on DEP-induced TNF-α release (Fig. 4).Figure 5 demonstrates that treatment of cultures of RAW 264.7 cells with DEP at a concentration of 250 μg/ml for 4 hr failed to induce any apoptotic response. Apoptosis was clearly induced in U 937 cells treated with campothecin (3μM × 4 hr), an apoptosis-inducing agent, and served as positive controls (Fig. 5). However, inhibition of NF-κB binding activity by BAY-11 resulted in DEP-induced apoptotic response as demonstrated by DNA fragmentation studies (Fig 5).Alveolar macrophages are the lung cells responsible for ingestion and clearance of inhaled particles [21]. They play a key role in lung inflammation, asthma pathogenesis, and regulation of airway remodelling.This view has been supported by correlation between the severity of asthma and the level of macrophage activation [22]. Activation of macrophages has been associated with increased production of the pro-inflammatory cytokines, IL-1β, TNF-α, IL-6, IL-4, platelet activating factor, and leukotrienes [22, 23]. TNF-α is a potent mediator of inflammatory and immune responses. Increased production of TNF-α by activated macrophages has been associated with pulmonary inflammation [24, 25]. TNF-α has been shown to rapidly induce NF-κB activation in cells expressing TNF-α receptors, TNFR-1 and TNFR-2 [26].The results of the current study demonstrate that DEP exposure activates alveolar macrophages by inducing the release of the inflammatory cytokine TNF-α (Fig. 3). The release of TNF-α was found to be regulated at the transcriptional level by the nuclear factor NF-κB, as demonstrated by BAY11-70811, an inhibitor of IκB phosphorylation and its subsequent degradation (Fig. 4). Adamson et al. [27] have shown that urban particulate matter where DEP are a major component, instilled into rat lung induced significant release of TNF-α. The results of the current study are in agreement with the reported findings by Adamson and co-workers, and demonstrate that the RAW 264.7 in vitro model is useful in studying the effect of DEP in the lung.Takano et al. [11] have shown that DEP (250μg/ml) increases NF-κB-DNA binding activity in the lung of mice with significant increase in macrophage-activating protein-1 (MP-1) and IL-1β release. In the present study, inhibition of NF-κB binding activity completely abrogated TNF-α protein expression (Fig. 4), indicating that DEP-induced TNF-α expression is regulated by NF-κB. IL-1β and TNF-α are pro-inflammatory cytokines and are both regulated at the transcriptional level by NF-κB. Moreover, TNF-α and IL-1β induce rapid nuclear translocation of NF-κB by inducing rapid phosphorylation and degradation of the inhibitory protein, IκB [15, 28–30]. Our results also demonstrate that IL-1β-induced TNF-α expression is regulated by NF-κB, as demonstrated by the inhibitor BAY-11 studies (Fig. 4). NF-κB has been implicated in proapoptotic as well as antiapoptotic signaling pathways in the lung [31–34]. Recently, however, many studies have suggested that translocation of NF-κB subunits to the nucleus serves to protect against apoptosis and enhance resistance to stimuli-induced cytotoxicity [35–38].The apoptotic response in macrophages may be regulated by the transcription factor NF-κB. Koay et al. [13] have shown that macrophage counts are associated with TNF-α release in response to LPS in the lung, and are essential for initiation of NF-κB-dependent immune response. Blocking of NF-κB binding activity has been shown to increase inflammatory cell apoptosis [39, 40]. Alveolar macrophages are the primary cells involved in the clearance of inhaled particles, pathogens, and apoptotic cells in the lung. Recent studies have indicated that clearance of apoptotic neutrophils by macrophages may be a determining step in the regression of COPD [41, 42]. Hiura et al. [8] have reported that DEP induce apoptosis in RAW 264.7 macrophages. However, in the same studies by Hiura and co-workers, exposure to DEP after its organic constituents have been extracted failed to induce any apoptotic response in RAW 264.7 cells. DEP-bound polycyclic aromatic hydrocarbons (PAHs), halogenated aromatic hydrocarbons (HAHs), and quinones have been implicated in the induction of apoptosis in alveolar macrophages [8], rather than extracted DEP.In our studies, DEP were treated for LPS and the organic constituents were extracted. The failure of DEP to induce an apoptotic response in RAW 264.7 cultures in the present study may be attributed to the absence of active organic constituents bound to DEP surfaces, notably, PAHs. Alveolar macrophages are rich in enzymes active in the transformation of foreign substances, for example, cytochrome P4501A1 (CYP 1A1) that contribute to the generation of reactive oxygen species (ROS). These enzymes are induced by PAHs and other DEP-bound constituents. DEP-adsorbed organic compounds have been shown to generate ROS by activating CYP 1A1 and NADPH quinine oxidoreductase in human airway epithelial cells [43]. Inhibition of NF-κB activation by antioxidants in a Jurkat T cell line has suggested that reactive oxygen intermediates are involved in the signalling pathways of NF-κB activation [44–46]. Despite the numerous studies that have addressed the cytotoxic effect of DEP on the alveolar macrophage and their role in NF-κB activation, the mechanisms underlying DEP-induced NF-κB activation remain not well defined. Earlier studies by Janssen-Heininger et al. [47] have demonstrated different signaling pathways for ROS and TNF-α in the activation of NF-κB. Using dominant negative Ras constructs, the authors were able to demonstrate the involvement of Ras in ROS-induced NF-κB activation in human airway epithelial cells. In the same studies, TNF-α-induced NF-κB activation, however, was reported to be Ras-independent. Whereas, the signaling pathways underlying TNF-α-induced NF-κB activation are well defined, oxidant-induced signaling events leading to NF-κB activation remain not well characterized. In the present study, uptake of DEP by alveolar macrophages may have resulted in a respiratory burst associated with superoxide and H2O2 generation. Released H2O2 and superoxide may have activated the mitogen-activated protein kinases/extracellular-regulated kinase kinase kinase-1 (MEKK-1) downstream from Ras. MEKK-1 has been shown to activate IκB kinase (IKK) and the c-Jun N terminal kinases (JNK), both capable of activating NF-κB downstream [48]. The activation of NF-κB and the transcription and release of TNF-α may have created an autocrine loop, where binding of released TNF-α to its receptor activates TNF receptor-associated factor (TRAF). The activation of TRAF and the subsequent activation of NF-κB-inducing kinase (NIK) would augment NF-κB activation and prevent apoptosis.In summary, the results of our study show that exposure to DEP at physiologically relevant concentrations (25–250 μg/ml) significantly induce the release of TNF-α, a cytokine implicated in inflammatory and immune responses. This effect is induced by the carbonaceous core of DEP since DEP utilized in the present study were treated for LPS and the adsorbed organic constituents extracted. Moreover, our results indicate that TNF-α release and activation of NF-κB may protect against apoptosis in the alveolar macrophage cell line, RAW 264.7.Effect of Diesel Exhaust Particles (DEP) on Tumor Necrosis Factor-α (TNF-α) Production in Alveolar Macrophages. Confluent cultures of alveolar macrophages (RAW 264.7) were exposed to vehicle (HBSS/30 mM Hepese) alone, vehicle containing DEP at 25, 100, or 250μg/ml, or IL-1β at 0.1μg/ml for 4 hours. Following exposure, TNF-α release was measured in control and exposed cultures using mouse ELISA. Data (pg/ml) are presented as mean ± SEM (n). *Significant difference (p < 0.05) from control values.Tumor Necrosis Factor-α (TNF-α) Protein Analysis by Western Blotting. Confluent monolayers of RAW 264.7 cells were treated with vehicle (HBSS/Hepes) alone, vehicle containing DEP (250 μg/ml), or IL-1β (0.1μg/ml) for 4 hours. Following the various treatments cells were harvested and lysed by sonication in lysis buffer on ice. Denatured total proteins (20 μg) of cytosolic and nuclear fractions were separated on 12% denaturing polyacrylamide gels and electrotransferred to Hybond nitrocellulose membranes. Following standard blocking and washing procedures, the membranes were incubated with polyclonal antibodies against TNF-α, and horseradish peroxidase-conjugated anti-goat IgG secondary antibodies. Detection of proteins was performed by enhanced chemiluminescence (ECL). Panel shows bands corresponding to treatment groups: 1. Control Cultures, 2. DEP-Treated Cultures, and 3. IL-1β-Treated Cultures.Effect of Diesel Exhaust Particles (DEP) on Alveolar Macrophage Apoptosis. Confluent monolayers of RAW 264.7 cells were treated with vehicle (HBSS/Hepes) alone (control), vehicle containing DEP at 250 μg/ml (DEP), or IL-1β at 0.1 μg/ml for 4 hr. Following treatment, DNA was isolated and fragmentation was assessed by DNA ladder kit on agarose according to the manufacturer’s directions. U937 cells treated with campothecin, an apoptosis-inducing agent were used as positive control. Pane shows bands corresponding to treatment groups: molecular weight marker, 1; positive control, 2; control cultures, 3, 4, 5, in triplicate; DEP-treated cultures, 6, 7, 8, in triplicate; and IL-1β-treated cultures, 9, 10, 11, in triplicate.Effect of Nuclear Factor-κB (NF-κB) Inhibition on Tumor Necrosis Factor-α (TNF-α) Production in Alveolar Macrophages. Monolayers of RAW 264.7 cells were pre-treated with BAY11-7082 (25μg/ml × 1 1/2 hr), an NF-κB inhibitor. Following pre-treatment, RAW 264.7 cultures were treated with vehicle (HBSS/Hepes) alone, or vehicle containing DEP (250 μg/ml × 4 hr) or IL-1β (0.1μg/ml × 4 hr), and TNF-α release in cellular supernatants were determined by enzyme-linked immunosorbent assays (ELISA). The data (pg/ml) are presented as mean ± SEM (n). *Significant inhibition (p < 0.05) from DEP-, or IL-1β-treated cultures.Effect of Nuclear Factor-κB (NF-κB) Inhibition on Diesel Exhaust Particles (DEP)-induced apoptosis in Alveolar Macrophages. Monolayers of RAW 264.7 cells were pre-treated with BAY11-7082 (25 μg/ml × 1 1/2 hr), an NF-κB inhibitor. Following pre-treatment, RAW 264.7 cultures were treated with vehicle (HBSS/Hepes) alone, or vehicle containing DEP (250 μg/ml × 4 hr). Following treatment, DNA was isolated and fragmentation was assessed by DNA ladder kit on agarose according to the manufacturer’s directions. U937 cells treated with campothecin (3 μM × 4 hr), an apoptosis-inducing agent were used as positive control (positive control). Lanes are as follows: molecular weight marker, 1; positive control, 2; DEP-exposed cultures pre-treated with BAY11-7082, 3; DEP-exposed cultures, 4; control cultures, 5; and control cultures pre-treated with BAY11-7082, 6.Funding for this study has been provided by a grant from the American Lung Association of Mississippi, and grant G12RR12359 from the NIH.
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+ DMBA, 7,12-dimethylbenz[a]anthracene, is a widely studied polycyclic aromatic hydrocarbon that has long been recognized as a probable human carcinogen. It has been found that DMBA is phototoxic in bacteria as well as in animal or human cells and photomutagenic in Salmonella typhimurium strain TA102. This article tempts to explain the photochemistry and photomutagenicity mechanism. Light irradiation converts DMBA into several photoproducts including benz[a]anthracene-7,12-dione, 7-hydroxy-12-keto-7-methylbenz[a]anthracene, 7,12-epidioxy-7,12-dihydro-DMBA, 7-hydroxymethyl-12-methylbenz[a]anthracene and 12-hydroxymethyl-7-methylbenz[a]anthracene. Structures of these photoproducts have been identified by either comparison with authentic samples or by NMR/MS. At least four other photoproducts need to be assigned. Photo-irradiation of DMBA in the presence of calf thymus DNA was similarly conducted and light-induced DMBA-DNA adducts were analyzed by 32P-postlabeling/TLC, which indicates that multiple DNA adducts were formed. This indicates that formation of DNA adducts might be the source of photomutagenicity of DMBA. Metabolites obtained from the metabolism of DMBA by rat liver microsomes were reacted with calf thymus DNA and the resulting DNA adducts were analyzed by 32P-postlabeling/TLC under identical conditions. Comparison of the DNA adduct profiles indicates that the DNA adducts formed from photo-irradiation are different from the DNA adducts formed due to the reaction of DMBA metabolites with DNA. These results suggest that photo-irradiation of DMBA can lead to genotoxicity through activation pathways different from those by microsomal metabolism of DMBA.Polycyclic aromatic hydrocarbons (PAHs) are widespread genotoxic and tumorigenic environmental pollutants. It has long been known that PAHs require metabolic activation in order to exert their biological activities, including carcinogenicity [1–4]. Upon metabolism, PAHs are either metabolized into biologically active metabolites, including diol epoxides and free radical intermediates, which bind to cellular DNA forming covalent DNA adducts responsible for mutagenicity and carcinogenicity [1–4]. However, recent studies have demonstrated that activation of PAHs can also be achieved by photo-irradiation [5–7]. Studies have shown that upon photo-irradiation, some PAHs are more toxic to microorganisms, plants, and other organisms than PAHs themselves without light irradiation [7, 8–11]. Since skin is the largest organ in human and when concomitantly exposed to environmental chemicals and sunlight, these chemicals may be activated by photo-irradiation and exert adverse health effects [12, 13].We have been interested in the photochemistry and phototoxicity of PAHs, including DNA single strand cleavage [14–18], photoreaction [18, 19], DNA damage [21, 22], and DNA covalent adduct formation [23]. 7,12-Dimethylbenz[a]anthracene (DMBA), is a widely studied PAH that has long been recognized as a probable human carcinogen [1–4]. This study reports that light irradiation of DMBA results in the formation of photodecomposition products including benz[a]anthracene-7,12-dione, 7,12-epidioxy-7,12-dihydro-DMBA, 7-hydroxymethyl-12-methylbenz[a]anthracene (7-HOCH2-12-MBA) and 12-hydroxymethyl-7-methylbenz[a]anthracene (12-HOCH2-7-MBA). Photo-irradiation of DMBA, 7-HOCH2-12-MBA and 12-HOCH2-7-MBA in the presence of calf thymus DNA followed by 32P-postlabeling/TLC analysis indicated that multiple DNA adducts are formed, and these adducts are different from the DNA adducts formed from reaction of DMBA metabolites with DNA. These results suggest that photo-irradiation of DMBA can lead to genotoxicity through activation pathways different from those by microsomal metabolism of DMBA.DMBA, lead tetraacetate, 2,3-dichloro-5,6-dicyanobenzoquinone (DDQ), glucose-6-phosphate dehydrogenase (type XII, Sigma), NADP+, and glucose-6-phosphate were purchased from Sigma Chemical Co. (St. Louis, MO). Cloned T4 polynucleotide kinase (PNK) was obtained from U.S. Biochemical Corp. (Cleveland, OH). [γ-32P]Adenosine 5′-triphosphate ([γ-32P]ATP) (sp. act. >7,000 Ci/mmol) was purchased from ICN Biomedicals, Inc. (Costa Mesa, CA). All other reagents were obtained through commercial sources and were the highest quality available. All solvents used were HPLC grade.According to the method published by Boyland and Sims [24], 7-HOCH2-12-MBA and 12-HOCH2-7-MBA were prepared by reaction of DMBA with lead tetraacetate, separation of the resulting 7-acetoxymethyl-12-MBA and 12-acetoxymethyl-7-MBA by column chromatography, and hydrolysis to the hydroxyl DMBA. 7-Formyl-12-methyllbenz[a]anthracene (7-CHO-12-MBA) and 12-Formyl-7-methyllbenz[a]anthracene (12-CHO-7-MBA) were synthesized by oxidation of 7-HOCH2-12-MBA and 12- HOCH2-7-MBA, respectively with DDQ [25]. 7,12-Epidioxy-7,12-dihydro-DMBA was prepared according to Wood et al. [26].3-Methylcholanthene-induced female Sprague-Dawley rat liver microsomes were purchased from In Vitro Technologies (Baltimore, MA). Protein concentrations were determined using a protein assay based on the Bradford method using a Bio-Rad protein detection kit (Bio-Rad Laboratories, Hercules, CA).The UVA light box was custom made with a 4-lamp unit using UVA lamps (National Biologics). The irradiance of light was determined using an Optronics OL754 Spectroradiometer (Optronics Laboratories, Orlando, FL), and the light dose was routinely measured using a Solar Light PMA-2110 UVA detector (Solar Light Inc., Philadelphia, PA). The maximum emission of the UVA is between 340 – 355 nm. The light intensities at wavelengths below 320 nm (UVB light) and above 400 nm (visible light) are about two orders of magnitude lower than the maximum at 340–355 nm.A solution (2–3 mL) of 0.4 mM DMBA dissolved in 90% ethanol was placed in a UV-transparent cuvette and photo-irradiated under UVA light to receive a light dose of 2.6 J/cm2/min for a period of 40, 90, and 360 min, respectively. The reaction mixture was then concentrated to about 200 μL under reduced pressure. Reversed-phase HPLC separation of the resulting photodecomposition products was accomplished using a Prodigy 5 μ ODS column (4.6 × 250 mm, Phenomenex, Torrance, CA) eluted isocratically with 90% methanol in water (v/v) at 1 mL/min. For isolation of photodecomposition products in larger amounts, a Prodigy 5 μ ODS column (10 × 250 mm, Phenomenex, Torrance, CA) eluted isocratically with 90% methanol in water (v/v) at 5 mL/min was used.A solution (0.4 mM dissolved in 2 mL of 90% ethanol) placed in a UV-transparent cuvette was added with 20 mg calf thymus DNA in 1 mL tetrahydrofuran. The resulting solution was photo-irradiated under UVA light with a total light dose of 14 J/cm2. After incubation, DNA was isolated and the DNA adducts were analyzed by 32P-postlabeling/TLC.Following a published procedure [27], approximately 10 μg of DNA was 32P-postlabeled using nuclease P1 enrichment. Adducts were separated by thin layer chromatography performed on 0.1 mm Machery Nagel 300 polyethylene imine cellulose plates (Alltech, Deerfield, IL) using the following solvent directions, D1: 0.9 M sodium phosphate, pH 6.8; D2: 3.6 M lithium formate, 8.5 M urea, pH 3.5; D3: 1.2 M lithium chloride, 0.5 M Tris HCl, 8 M urea, pH 8.0. A final wash was conducted in D3 with solvent used in D1. Areas of radioactivity were measured with a Storm 860 phosphor imaging system (Molecular Dynamics, Sunnyvale, CA).In vitro metabolism was carried out by incubation of DMBA solution (0.8 mM dissolved in 200 μl of acetone) with shaking at 37° for 30 min in a 10 ml reaction mixture containing 0.5 mM of Tris-HCl (pH 7.5), 30 μM of MgCl2, 1 unit of glucose-6-phosphate dehydrogenase (type XII, Sigma), 1 mg NADP+, 6 mg of glucose-6-phosphate, 10 mg of microsomal protein, and 10 mg of purified calf thymus DNA. After incubation, the reaction was cooled with ice-water, then sequentially extracted with 5.0 mL phenol, 5.0 mL phenol/chloroform/isoamyl alcohol (v/v/v, 25/24/1), and 5.0 mL chloroform/isoamyl alcohol (v/v, 24/1). The DNA in the aqueous phase was precipitated by adding 1 mL 5 M sodium chloride followed by equal volume of cold ethanol and washed three times with 70% ethanol. After redissolving in 3 mL distilled water, the DNA concentration and purity were determined spectrophotometrically, and DNA adducts were analyzed by 32P-postlabeling/TLC analysis with the method described above.A Waters HPLC system consisting of a Model 600 controller, a Model 996 photodiode array detector, and pump was used for separation and purification of DMBA photodecomposition products. Direct exposure probe (DEP) mass spectrometry (MS) was performed on a ThermoFinnigan TSQ 700 triple quadrupole mass spectrometer operated in the electron ionization (EI) mode.Photo-irradiation of DMBA in ethanol/water (v/v, 90/10) by UVA light at a light dose of 2.6 J/cm2/min for a period of 40, 90, and 360 min, respectively was conducted and the reaction mixture was separated by reversed phase HPLC (Figure 1). Based on comparison of the HPLC retention time, UV-absorption spectrum, and mass spectrum with those of DMBA, the material contained in the chromatographic peak eluting at 19.0 min was identified as the recovered DMBA. As shown in Figure 1A, the amount of DMBA decreased and the amounts of photodecomposition products increased rapidly. For collection of sufficient amount of the photodecomposition products for structural identification, the products formed after 360 min of photo-irradiation were separated by repeated preparative HPLC (Figure 2). Based on mass (Figure 3A) and NMR (Figure 4A) spectral analysis, the material in the chromatographic peak eluting at 5.3 min (P5 in Figure 1C) was tentatively identified as 7-hydroxy-12-keto-7-methylbenz[a]anthracene (7-OH-12-keto-7-MBA). The chromatographic peak eluting at 6.6 min (P8) was identified as 7,12-epidioxy-7,12-dihydro-DMBA. This is based on the comparison of its UV-visible absorption spectrum, HPLC retention time, mass spectrum (Figure 3B), and NMR spectrum (Figure 4B) with those of the authentic sample (data not shown) [26]. The material in chromatographic peak eluting at 5.8 min (P6) (Figure 2) had a mass spectrum with a molecular ion M+ at m/z 272 (data not shown), suggesting this is an oxygenated DMBA. This compound has the mass spectrum, UV-visible absorption spectrum (Figure 5A insert) and HPLC retention time (Figure 5A) identical to those of the synthetic standard for 7-HOCH2-12-MBA (Figure 5B). Thus, it confirms that this photodecomposition product is 7-HOCH2-12-MBA. The material in chromatographic peak (P7) eluting at 6.2 min was similarly identified as 12-HOCH2-7-MBA using a synthetic standard. Based on comparison of HPLC retention time and UV-visible absorption spectrum (Figure 5C and insert) with those of the authentic BA-7,12-dione (Figure 5D and insert), the chromatographic peak (P9) eluting at 9.4 min was identified as BA-7,12-dione.The formation and decomposition of the five identified photodecomposition products, P5-P9, were studied. As shown in Figure 6, while DMBA completely decomposed at about 260 min under light irradiation, the photodecomposition products P5 and P8 reached the highest yield at about 400 min of irradiation time. Compound P9 kept increasing, suggesting that the other decomposition products gradually converted into BA-7,12-dione (P9). Compounds P6 and P7 also increased during the whole course of irradiation.Photo-irradiation of DMBA, 7-HOCH2-12-MBA, 12-HOCH2-7-MBA, 7-CHO-12-MBA, and 12-CHO-7-MBA in the presence of calf thymus DNA was carried out, the resulting DNA was isolated, and the DNA adducts were analyzed by 32P-postlabeling/TLC. Although both 7-CHO-12-MBA and 12-CHO-7-MBA were not formed as photodecomposition products in our study, these compounds were formed as reported by Wood et al. (26). Therefore, for facilitating in mechanistic understanding, the DNA adduct formation from these two compounds was also pursued.As shown in Figure 7, the 3′,5′-bisphosphate deoxyribonucleosides of DMBA (Figure 7A) and the four oxidized derivatives (Figure 7C–7F) were separated by thin-layer chromatography (TLC) into multiple spots. Analysis of these resulting TLC spots indicated that the spot profiles are nearly identical.Figure 8 shows the autoradiogram of 32P-postlabeled nuclease P1-treated calf thymus DNA from (A) photo-irradiation of the DNA in the presence of DMBA by UVA light (14 J/cm2) and (B) the metabolite mixture from metabolism of DMBA by rat liver microsomes in the presence of calf thymus DNA.For comparison of DNA adduct profile, the metabolites obtained from metabolism of DMBA by rat liver microsome in the presence of calf thymus DNA were also analyzed by 32P-postlabeling/TLC under identical conditions. Comparison of the DNA adduct profiles indicates that the DNA adducts formed from photo-irradiation of DMBA are different from those from reaction of DMBA metabolites (Figure 8).In this study, photo-irradiation of DMBA under UVA light resulted in the formation of multiple photodecomposition products, of which four products were identified, including benz[a]anthracene-7,12-dione, 7,12-epidioxy-7,12-dihydro-DMBA, 7-HOCH2-12-MBA, and 12-HOCH2-7-MBA, and one product tentatively assigned as 7-hydroxy-12-keto-7-MBA (Figure 1). Although Wood et al. [26] reported that 7-CHO-12-MBA and 12-CHO-7-MBA were produced from photo-oxidation of DMBA, these two compounds were not formed under our experimental conditions. This discrepancy illustrates that the formation of photo-oxidation products from DMBA is highly dependent on experimental conditions, particularly the light wavelength.The results of 32P-postlabeling/TLC analysis indicate that photo-irradiation of DMBA and the four oxidized derivatives (7-HOCH2-12-MBA, 12-HOCH2-7-MBA, 7-CHO-12-MBA, and 12-CHO-7-MBA) by UVA light in the presence of calf thymus DNA all generated multiple DNA adducts and the DNA adduct profiles are nearly identical.These results highly suggest that the photo-induced DNA adduct formation from DMBA is mediated through these oxidized derivatives by two distinct pathways: (i) oxidation of DMBA at the 7-methyl group to 7-HOCH2-12-MBA, then 7-CHO-12-MBA, then to the reactive species; and (ii) oxidation of DMBA at the 12-methyl group to 12-HOCH2-7-MBA, then 12-CHO-7-MBA, then to the reactive species. Also the kinetic study of photo-irradiation of DMBA, as shown in Figure 6, indicates that BA-7,12-dione is the final and stable product. Consequently, we conclude that the reactive photodecomposition product that can bind to DNA and form DNA adducts is from 7-CHO-12-MBA or 12-CHO-7-MBA, and is further oxidized to the inert BA-7,12-dione. Comparison of the DNA adduct profiles indicates that the DNA adducts formed from photo-irradiation of DMBA and from metabolism of DMBA by 3-methylcholanthrene-induced rat liver microsomes are different. These results suggest that photo-irradiation of DMBA can lead to genotoxicity through activation pathways different from those by microsomal metabolism of DMBA.Thus, our study indicates that photo-irradiation of DMBA generates genotoxic photo-oxidation products that can lead to DNA adduct formation. Besides, Boyland et al. reported that 7-HOCH2-12-MBA is able to cause adrenal apoplexy and mammary cancer in rats [27]. 7,12-Epidioxy-7,12-dihydro-DMBA has been shown to be toxic to chicken fibroblast cells [28]. Consequently, these results suggest that photo-irradiation of PAHs can generate genotoxic products and can be highly harmful to human health. This warrants further investigation.HPLC profiles of photoproducts of DMBA after irradiation with UVA light (2.6 J/cm2/min) in ethanol for (A) 40 min; (B) 90 min; (C) 360 min.Identification of some of the photoproducts of DMBA in ethanol after 360 min of irradiation using UVA light (2.6 J/cm2/min).Mass spectrum profiles of purified P5 (A) and P8 (B) of DMBA photoproducts.1H-NMR spectra of purified P5 (A) and P8 (B) of DMBA photoproducts.HPLC and UV spectrum profiles of 7-hydroxymethyl-12-methylbenz[a]anthracene standard (A), purified P6 from DMBA photoproducts (B); benz[a]anthracen-7,12-dione standard (C), and purified P9 from DMBA photoproducts (D).Time course of photodecomposition of DMBA and formation and photodecomposition of the identified DMBA photodecomposition products.Autoradiogram of 32P-postlabeled nuclease P1-treated DNA from photo-irradiation in the presence of (A) DMBA, (B) blank, (C) 7-HOCH2-12-MBA, (D) 12-HOCH2-7-MBA, (E) 7-CHO-12-MBA, and (F) 12-CHO-7-MBA in THF and calf thymus DNA by UVA light at a total dose of 14 J/cm2.Autoradiogram of 32P-postlabeled nuclease P1-treated calf thymus DNA from (A) photo-irradiation of the DNA in the presence of DMBA by UVA light (14 J/cm2) and (B) the metabolite mixture from metabolism of DMBA by rat liver microsomes in the presence of calf thymus DNA.This research was in part supported by the US Army Research Office DAAD 1901-1-0733 to JSU and the National Institutes of Health: NIH-SCORE S06 GM08047. We thank the NIH-RCMI for core Molecular and Cellular Biology and Analytical Chemistry facilities at JSU.
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+ Although the role of systemic proinflammatory cytokines, IL-1β and TNF-α, and their up-regulation of adhesion molecules, ICAM-1, VCAM-1 and E-Selectin, in the pathogenesis of cerebral malaria (CM) is well established, the role of local cytokine release remain unclear. Immunohistochemistry (IHC) was used to compare the expression of ICAM-1, VCAM-1, E-Selectin, IL-1β, TNF-α and TGF- β at light microscopic level in cerebral, cerebellar and brainstem postmortem cryostat sections from 10 CM, 5 severe malarial anemia (SMA), 1 purulent bacterial meningitis (PBM), 2 non-central nervous system infections (NCNSI) and 3 non-infections (NI) deaths in Ghanaian children. Fatal malaria and Salmonella sepsis showed significantly higher vascular expression of all 3 adhesion molecules, with highly significant co-localization with sequestration in the malaria cases. However, there was negligible difference between CM and SMA. TGF-β showed intravascular and perivascular distribution in all cases, but expression was most intense in the PBM case and CM group. TNF-α and IL-1β showed prominent brain parenchymal staining, in addition to intravascular and perivascular staining, in only the PBM case and CM group. The maximal expression of all 6 antigens studied was in the cerebellar sections of the malaria cases. Endothelial activation is a feature of fatal malaria and Salmonella sepsis, with adhesion molecule expression being highly correlated with sequestration. IL-1β and TNF-α are upregulated in only cases with neurodegenerative lesions, whilst TGF-β is present in all cases. Both cytokines and adhesion molecules were maximally upregulated in the cerebellar sections of the malaria cases.Despite technological advances and global economic development, malaria is still the parasitic disease responsible for the greatest number of deaths worldwide. Malaria parasites infects between 300 to 500 million people, causing up to 2 million deaths globally per year (mostly children in sub-Saharan Africa) from complications of primarily cerebral malaria (CM) and severe malarial anaemia (SMA) [1]. CM, characterised by seizures and loss of consciousness, is an important complication of Plasmodium falciparum infection with mortality rate of 15–20% [2]. Despite this high mortality rate, the pathogenic mechanisms of CM have not been well elucidated. Little is known about the blood brain barrier (BBB) interactions in CM that result in the neurological disorder, and it is unclear how the intraerythrocytic parasite, which sequesters in the cerebral microvasculature but rarely enters the brain parenchyma, influence parenchymal function to induce coma and death. The unavailability of infected human specimens and suitable animal models has hindered a thorough understanding of the pathogenesis.There is compelling evidence that the inflammatory response at the BBB in CM involve increased systemic production of proinflammatory cytokines, especially tumour necrosis factor (TNF)-α and interleukin (IL)-1β [2–6], and up-regulation of cerebral endothelial adhesion molecules, especially intercellular adhesion molecule (ICAM)-1, vascular cell adhesion molecules (VCAM)-1 and endothelial (E)-selectin, that facilitate sticking of parasitized erythrocytes (PEs) to blood vessels [7–10]. There is immunohistochemical (IHC) evidence for endothelial cell (EC) activation during CM in humans [8, 10–12], and animal models of CM [13]. Endothelial activation occurred not only in fatal CM, but also in the other fatal malaria syndromes, indicating that systemic endothelial activation is a feature of fatal malaria [8, 10, 11]. A role for tyrosine phosphorylation has been demonstrated in EC activation for two adhesion molecules (VCAM-1 and E-Selectin) induced by TNF-α [14].In addition to systemic proinflammatory cytokine production, local cytokine release could contribute to organ-specific pathophysiology, especially in the brain in the case of CM. Recent studies have shown the expression of proinflammatory cytokine protein, including TNF-α, in postmortem brain tissue in human CM [5, 15, 16]. TNF-α is over-expressed in cerebral microglia, astrocytes, monocytes and vascular endothelium in mice with CM, relative to controls [15–20]. However, the lack of a classical inflammatory response to the presence of PEs in the brain microvasculature indicates anti-inflammatory cytokine involvement, and recent reports have suggested an anti-inflammatory and neuroprotective role for transforming growth factor (TGF)-β in the host defense mechanism against neuronal loss in neurodegenerative diseases [21], including CM.TGF-β has been found in both haemorrhagic white-matter lesions of human CM [5], and in white-matter lesions of human immunodeficiency virus (HIV)-1 encephalitis brain samples [22]. This association of TGF-β with central nervous system (CNS) neurodegenerative lesions suggests an anti-inflammatory and neuroprotective role for TGF-β in the host defense mechanism against neuronal cell loss [21]. In mice infected with lethal or non-lethal strains of malaria parasites, a strong and sustained TGF-β response, beginning on the 5th to 6th post-infection day when the peak parasite replication has been reached, was associated with abrogation of mortality and resolution of infection [23]. Furthermore, neutralization of TGF-β leads to 100% mortality in BALB/c mice infected with normally non-lethal P. chabaudi A/J [23].In-vitro studies have shown TNF-α mRNA in inflammatory infiltrates within the meninges of experimental rabbit pneumococcal meningitis [24], and upregulation of neuronal TNF-α expression in response to bacterial lipopolysaccharide [25]. IL-1β is not expressed in normal human brain, but induced and expressed intraparenchymally in human CM brain and in meningeal infiltrating leukocytes of human meningoencephalitides cases [5]. In in-vitro studies, IL-1β is neurotoxic and rapidly induced in response to neuronal cell death, and therefore is suggested to play a causal role in ischaemic cell death and neurodegeneration in the brain [26]. Tongren and colleagues [27] showed that the proinflammatory cytokines (IL-1β and TNF-α) and TH-1 cytokine [interferon (IFN)-γ] had the highest level of mRNA expression in the cerebellum during late P. coatneyi infection in rhesus monkeys, agreeing with histopathologic observations of the preferential sequestration of PE in the cerebellum in rhesus monkey [27–29], and in human CM [30]. However, recent evidence from mice indicates that it may be overproduction of lymphotoxin-α (LT-α) rather than TNF-α that leads to CM, since mice deficient in TNF-α were found to be just as susceptible to CM as controls whereas LT-α deficient mice were resistant to CM pathology, dying from hyperparasitaemia and severe anaemia instead [31].Although, the role of cytokines and adhesion molecules has been extensively studied in human malaria, the role of local cytokine release in the brain in human CM is unclear. Children living in sub-Saharan Africa bear the brunt of malaria mortality, yet there is a dearth of relevant postmortem studies in African children, and most of the few human CM postmortem immunohistochemical (IHC) studies are in non-African adults. Therefore, the role of local cytokine release and associated up-regulation of adhesion molecules in human CM has not been adequately confirmed and substantiated, especially in sub-Saharan African children.The present study was aimed at ascertaining the role of cytokines (TNF-α, IL-1β and TGF-β) and adhesion molecules (ICAM-1, VCAM-1 and E-Selectin) in fatal CM, and used IHC techniques to examine and compare the distribution of these 6 antigens at light microscopical level in postmortem human cerebral, cerebellar and brainstem cryostat sections of Ghanaian children dying from CM, severe malarial anaemia (SMA), purulent bacterial meningitis (PBM), non-central nervous system infections (NCNSI) and non-infections (NI) cases.During the peak malaria season, from July to September 2001, all clinically certified deaths in children admitted to the Emergency Unit at the Department of Child Health, Korle-Bu Teaching Hospital, Accra, Ghana, with detailed clinical and laboratory records and in whom duly signed written informed consent had been obtained from parents or guardians after the death of their child were included in the study. Twenty-one (21) parents or guardians agreed to and signed the consent form for the participation of their children and to donate tiny brain tissue samples at autopsy for this study. Volunteered cadavers were immediately moved from the Emergency Unit to the morgue for cold storage at 4°C and a full autopsy with removal of brain tissue samples done within 12 hours of death.In each volunteered case, brain tissue blocks of about 0.3–0.5cm3 were surgically removed from 3 regions of the brain, namely cerebrum, cerebellum and brainstem. Brain smear cytology was done for each case at autopsy and stained with Giemsa for the demonstration of parasitized erythrocytes and/or malaria pigment in the cerebral microvasculature as previously described [32]. The full autopsy gross findings, autopsy brain smear cytologic/microscopic findings, in addition to the detailed clinical and diagnostic laboratory records were used to classify the 21 cases into five groups, namely:Cerebral malaria [CM].Malaria complicated by severe anaemia/severe malarial anaemia [SMA].Purulent bacterial meningitis [PBM] (i.e. central nervous system infection other than cerebral malaria)Non-central nervous system infection [NCNSI] (i.e. infection in an anatomic organ-system other than the central nervous system), and (5) Non-infection deaths [NI] (i.e. no focus of infection found clinically or at autopsy).A CM death was defined as clinically fulfilling WHO definition of severe malaria [33] with Blantyre coma score ≤2, and gross autopsy findings of slaty-gray discolouration and/or white matter petechial haemorrhages of brain, and/or brain smear cytologic findings of parasitised erythrocytes and/or malaria pigment in the cerebral microvasculature. SMA deaths fulfilled WHO definition of severe malaria [33] with hemoglobin ≤ 5 g/dl, and absence of gross autopsy findings of slaty-gray discolouration and white matter petechial haemorrhages, but presence of moderate to severe pallor of all internal organs, and absence of brain smear cytologic findings of parasitised erythrocytes and malaria pigment in the cerebral microvasculature. Deaths with clinically negative P. falciparum peripheral parasitaemia, absence of gross autopsy findings of slatey gray discolouration of brain, liver and spleen, and white matter petechial haemorrhages, and absence of brain smear cytologic findings of parasitised erythrocytes and malaria pigment in the cerebral microvasculature, were included as non-malartia controls.Based on above criteria, the 21 volunteered cases were made up of 10 CM, 5 SMA, 1 PBM, 2 NCNSI, and 3 NI deaths. The study was approved by both the Ethical and Protocol Review Committee of the University of Ghana Medical School, and the Institutional Review Board (IRB) of the Noguchi Memorial Institute for Medical Research, Accra, Ghana.All the sixty-three (63) brain tissue samples obtained at autopsy within 12 hours of death were immediately snap-frozen in liquid nitrogen, and stored at −80°C until used. Fourteen (14) cryostat sections of 5–7 μm thickness were cut from each of the 63 brain tissue samples, mounted unto SuperFrost® Plus (Menzel-Glazer, Germany) light microscope slides and 2 of each stained for the localization of the six antigens studied and two for negative controls. Two independent experiments, using 6 sections for the 6 antigens and 1 negative control section in each, were performed by identical methods for each of the 63 samples to ensure reproducibility.Cryostat sections were air-dried overnight at room temperature, fixed in 100% acetone for 10 minutes, and air-dried for 30 minutes at room temperature. Slides were immunostained immediately using a standard indirect alkaline phosphatase method. Briefly, sections were incubated with a 1 : 20 dilution of normal rabbit serum (DAKO, Denmark) in Tris-buffered saline (TBS), pH 7.4 for 30 minutes in a humidified chamber to reduce non-specific protein binding. Primary antibodies against ICAM-1 (mouse anti-human, monoclonal, Immunotec, UK; dilution 1 : 500), VCAM-1 (mouse anti-human, monoclonal, DAKO, UK; dilution 1 : 200), E-Selectin (mouse anti-human, monoclonal, DAKO, UK; dilution 1 : 200), TNF-α (mouse anti-human, monoclonal, Serotec, UK; dilution 1 : 250), IL-1β (mouse anti-human, monoclonal, Serotec, UK; dilution 1 : 250) and TGF-β (mouse anti-human, monoclonal, Serotec, UK; dilution 1 : 250) were applied for 60 minutes at room temperature. Antibodies were used as dilutions in TBS, pH 7.4.Sections were then washed with TBS pH 7.4 for 15 minutes (3 times, each for 5 minutes). Secondary incubation and staining were performed using indirect immunostaining microscopic technique with Universal DAKO Alkaline Phosphatase Anti-Alkaline Phosphatase (APAAP) Kit™ System 40 for Monoclonal Mouse Antibody (DAKO Corporation, USA). After incubating with freshly prepared chromogenic substrate for 8–10 minutes, the sections were then washed in tap water, lightly counterstained with Meyer’s Haematoxylin for 3 minutes, washed again, and mounted with glycergel (DAKO, Denmark). Negative control slides were prepared for each of the 63 samples by substituting the primary antibody with normal non-immune mouse IgG (DAKO, Denmark). Appropriate concentrations of primary antibodies were determined using optimization on malaria and non-malaria tissues.Slides were examined and graded blindly by two independent authors [HA & EKW], who were not involved in the immunostaining of the sections and had no access to the clinical, laboratory and autopsy details of volunteers until after the results had been collated. Ten fields per section, immunostained with ICAM-1, VCAM-1 and E-Selectin, were examined under a magnification of X400 and the degree of staining and the number of vessels showing staining were assessed using a five-tiered semi-quantitative scale: −, negative (no endothelial cell staining); +/−, positive endothelial cell staining on < 25% of vessels; +, positive endothelial cell staining on > 25% and < 50% of vessels; ++, positive endothelial cell staining on > 50% and < 75% of vessels; and +++, positive endothelial cell staining on > 75% of vessels.Similarly, ten fields per section, immunostained for TNF-α, TGF-β and IL-1β, were examined under a magnification of X100 and the location and intensity of staining were assessed using a five-tiered semi-quantitative scale: −, no red reaction product/colour; +/−, faint/mild staining in some fields; +, moderate staining in some fields; ++, moderate consistent staining in all fields; and +++, deep/strong consistent staining in all fields.Furthermore, ten fields per section, immunostained with ICAM-1, VCAM-1 and E-Selectin, were examined under a magnification of X400 for each of the 90 brain tissue sections, 2 each from the 3 regions of the brain of the 15 malaria cases (10 CM & 5 SMA), and the following four categories of vessels were counted, namely the number of vessels showing both receptor expression and sequestration (R+S+), the number of vessels showing receptor expression but no sequestration (R+S−), the number of vessels showing no receptor expression but sequestration (R−S+), and the number of vessels showing neither receptor expression nor sequestration (R−S−).For each of the six antigens studied, the percentage of immunostained cases and the intensity of staining were compared between the five groups of cases and the three regions of the brain from which samples were obtained for this study. The co-localization of sequestration and receptor expression was analyzed using a chi-square (X2) test on the 2 × 2 contingency tables constructed for each of the three adhesion molecules by comparing the expected and observed association in cerebral vessels for the malaria cases (both CM and SMA groups). The level of significance was five percent and all analyses were performed using Epi Info 2002 statistical software (CDC).The clinical and diagnostic details of the 21 studied cases are summarized in Table 1. Ten died of CM, five of SMA, one of PBM, two of NCNSI and three of NI. The two NCNSI cases were severe bronchopneumonia (SBP), and typhoid perforation and septicaemia (TP&S), whilst the three NI cases were nephrotic syndrome (NPS), bleeding duodenal ulcer (BDU) and haemolytic sickle cell (SC) crisis (HSC). There were 10 males and 11 females. The mean age of the CM deaths of 69.1 months (SD = 32.1) was significantly higher than that of the SMA deaths of 20.2 months (SD = 16.25) [X2 = 7.33, df = 1, P value = 0.0068]. Eighty percent (12 out of 15) of the malaria deaths occurred within 24 hours of admission. The mean peak parasitemia of the CM deaths of 45,107/μL (SD = 27,091) was significantly lower than that of the SMA deaths of 198,453 /μL (SD = 146,913) [X2 = 14.68, df = 1, P value = 0.0001].Immunostaining for the 3 adhesion molecules showed only vascular labeling (Figure 1A–F). There was no one particular receptor whose expression on the endothelial cell surface was consistently related to the presence of sequestered PEs in the cerebral microvessels of the malaria cases. Generally, if a vessel showed positive staining for a marker of endothelial activation, this was present throughout its length, rather than being related to the presence of PEs in one segment or over one endothelial cell. Within vessels from the same patient and same region of the brain of the fatal malaria cases, there was heterogeneity between parasite sequestration and the presence of receptor staining. Hence, not uncommonly, there were vessels engorged with PEs without any evidence of adhesion molecule expression and vice versa.ICAM-1 showed positive staining of the endothelia of < 25% of the brain microvessels (average score of +/−) in all the non-malaria deaths (Figure 1A), whilst VCAM-1 showed positive staining of the endothelia of < 25% of the brain microvessels (average score of +/−) in some (highest being in the cerebellar sections of 4 out of 5) of the non-malaria deaths (Figure 1B), except the case of typhoid perforation and septicaemia. E-Selectin staining was consistently and uniformly negative in all the cerebral microvessels (score of −) in all the non-malaria controls, with the exception of the case of typhoid perforation and septicaemia. ICAM-1 (Figure 1C), VCAM-1 and E-Selectin showed positive staining in 25 to 50% of the microvessels (average score of +) in all the brain sections of the case of typhoid perforation and septicaemia.The expression of these three sequestration receptors on the cerebral endothelium of the malaria cases showed an increase in both the proportion of cases showing expression and the intensity of staining. ICAM-1 showed positive staining in nearly all (> 75%) vessels (average score of +++) in all malaria cases (Figure 1D). VCAM-1 showed positive staining in most (50 to 75%) of the microvessels (average score of ++) in all the malaria cases, whilst E-selectin showed positive staining in 25 to 50% of the microvessels (average score of +) in all the malaria cases. There was no significant differences in endothelial staining between the 10 cases of fatal CM and the 5 cases of fatal SMA, since both groups had a similar increased intensity of ICAM-1, VCAM-1 and E-selectin expression of +++, ++ and + respectively. However, the increased adhesion molecule expression was more often not associated with sequestered PEs in the SMA cases, than for the CM cases. Thus, though heterogeneity between sequestration and receptor expression was observed in all the malaria cases, it was more extreme in the SMA cases (Figures 1D & 1E). The staining patterns of all the 3 adhesion molecules in brain sections from the case of typhoid perforation and septicaemia differed significantly from that of the other 5 non-malaria cases and the 15 malaria cases. The intensity of staining of all 3 adhesion molecules were significantly higher than that of the other 5 non-malaria cases, but not as intense as that of the 15 fatal malaria cases. Generally, the maximal expression (in terms of increased percentage of cases showing expression and intensity of staining) of all the 3 adhesion molecules in the malaria cases was evident in the cerebellar sections (Figures 1D & 1F).We found that the presence of sequestered PEs was highly significantly associated with the expression of ICAM-1 (P = 3.1 X 10−16), VCAM-1 (P = 1.2 X 10−12) and E-selection (P = 6.1 X 10−16) [degrees of freedom (df) = 1] in the fatal malaria cerebral vessels (Table 2). The relative risk of cerebral vessels expressing ICAM-1, VCAM-1 and E-Selectin showing sequestered PEs was 1.73, 1.53 and 1.67 respectively.Positive immunostaining for TGF-β showed intravascular and perivascular distribution (Figures 2A & 2B), whilst there was intravascular, perivascular and prominent brain parenchymal staining for IL-1β (Figures 2C–E) and TNF-α (Figures 2F). TGF-β was detected in intravascular and perivascular distribution in brain tissue from all 5 groups studied, but expression was more intense in PBM and CM groups (Figures 2A & 2B). TGF-β showed moderate intravascular and perivascular immunostaining in some fields (average score of +) of all the NCNSI sections (Figure 2A), whilst it’s staining was strong in intravascular and perivascular locations in all fields (average score of +++) of all the CM and PBM sections (Figure 2B).IL-1β staining was limited to only some (none of the brainstem sections) of the PBM brain sections (Figure 2C) and all of the CM brain sections (Figure 2D) in a predominantly intravascular and perivascular pattern, but in none of the sections of the other three groups (Figure 2E). Additionally, it was more intense in CM group than the PBM case (Figures 2C & D). IL-1β showed moderate intravascular and perivascular immunostaining in some fields (average score of +) of some of the PBM sections (none of the PBM brainstem sections showed any IL-1β staining) [Figure 2C]. IL-1β showed strong intravascular and perivascular immunostaining in all fields (average score of +++) of the CM sections (Figure 2D), whilst it showed no staining (score of −) in none of the fields of all the NCNSI sections (Figure 2E).TNF-α was expressed in all the PBM and CM (Figure 2F) brain sections in intravascular, perivascular and intraparenchymal pattern, but none of the sections of the other three groups. The intensity of staining was, generally, more intense in the CM group compared to the PBM case. TNF-α showed moderate intravascular, perivascular and parenchymal immunostaining in some fields (average score of +) of the PBM sections. TNF-α showed strong intravascular, perivascular and parenchymal staining in all fields (average score of +++) of all the CM sections (Figure 2F), whilst it showed no immunostaining in none of the fields (score of −) of all the NCNSI sections. Generally, the expression of all the three cytokines studied was highest in the cerebellar sections of the studied cases.The pathology of fatal P. falcipaium malaria has been extensively investigated, but many areas of controversy and inadequate knowledge still remain. Several hypotheses have been developed to explain the pathogenesis of CM. The release of Plasmodium GPI toxin, production of pro-inflammatory cytokines (both systematically and locally), up-regulation of cerebral endothelial adhesion molecule expression and associated sequestration of PEs and their downstream consequences, such as mechanical blockage, ischaemia, acidosis, haemorrhage, and nitric oxide production have been implicated in the pathogenesis [3]. Most studies, however, have focused on plasma and CSF levels of cytokines in clinical studies or have used animal models in tissue studies. There is a dearth of direct evidence for local cytokines release in human CM brains [5, 15, 16], mainly because of difficulties in obtaining human post-mortem tissue from malaria cases.Malaria-induced brain inflammation is known to be mediated partly by complex cellular and immunomodulator interactions involving co-regulators such as cytokines and adhesion molecules, resulting in the sequestration of Plasmodium-infected erythrocytes. However, the role of sequestered platelets and leucocytes, chemokines and chemokine receptors in malaria brain immunopathogenesis still remain unclear. Apart from the sequestration of Plasmodium-infected erythrocytes, recent studies [34–37] have revealed significant accumulation of platelets and leukocytes in the distal microvasculature of the brains of human cases of CM, suggesting a role for platelet and leukocyte sequestration in human CM pathology. Sarfo et al (2004) recently reported up-regulated expression of RANTES and its receptors (CCR3 and CCR5) in the cerebellar and cerebral regions of post-mortem human CM brains [38]. Additionally, others [39, 40] have reported increased migration of CCR5+ leukocytes into the brain in experimental murine CM models. These studies support the hypothesis that leukocyte recruitment by chemokines may play a role in the pathogenesis of human CM.In this study, we identified and localized the induced expressions of ICAM-1, VCAM-1, E-Selectin, TNF-α, TGF-β and IL-1β in 3 different regions of the brain during human CM, non-CM and non-malaria deaths to ensure a more extensive and exhaustic comparison in ascertaining the role of local production of cytokines and adhesion molecule expression in the brain in human CM. IHC analysis revealed differential expression patterns of the 6 antigens studied in the 3 brain regions and 5 groups of diseases. Fatal malaria (both CM and SMA) and Salmonella septicaemia were associated with induction of endothelial activation markers, with significantly higher levels of ICAM-1, VCAM-1 and E-Selectin expression on vessels in the brain compared to non-malaria controls. All the non-malaria (NM) controls, except the case of salmonella septicaemia, showed a low-level of ICAM-1 and VCAM-1 expression, but no E-Selectin expression consistently. ICAM-1 was most widely expressed and intense in the malaria cases, and hence may mediate the bulk of PE sequestration.Furthermore, we observed no significant differences between the endothelial receptor immunostaining in the CM and SMA cases, as previously reported by others [8, 10, 11], and indicate that systemic endothelial activation is a feature of fatal malaria and systemic sepsis. There was highly significant co-localization of sequestration with the expression of ICAM-1, VCAM-1 and E-Selectin in cerebral vessels of the malaria cases, as previously observed [10], and further supports a role for these receptors in sequestration in-vivo. The observation of vessels engorged with sequestered PEs without any evidence of adhesion molecule expression, and vice versa, was not a rarity. The heterogeneity observed in the distribution of sequestered PEs and receptor expression in the fatal malaria sections may suggest that other factors like sequestered platelets and leucocytes, in addition to the sequestered PEs, may play a role in the pathogenesis of CM. The heterogeneity between PE sequestration and adhesion molecule expression in our study may indeed reflect the degree of sequestration of chemokine-releasing leucocytes and the consequent endothelial cell activation at the site.TGF-β was detected in an intravascular and perivascular distribution, but not intraparenchymal, in the brain sections from all the 5 groups studied, but expression was most intense in the meningitis and CM groups, thus the cases with neurodegenerative lesions. As previously suggested [5], serum leakage may be the most probable principal source of the low-level expression in the 3 disease groups with no neurodegenerative lesions (SMA, NCNSI & NI), whilst the more intense expression in the 2 disease groups with neurodegenerative lesions (CM & PBM) may be the result of additional production by reactive glial responding to local tissue damage. This finding is similar to that of a previous report [5], and further supports the suggested anti-inflammatory and neuroprotective role of TGF-β in host defense mechanism against neuronal cell loss [21], since recently TGF-β has been associated with neurodegenerative lesions [5, 21, 22].TNF-α and IL-1β were detected within the brain parenchyma in only the PBM and CM groups, suggesting neuronal and/or glial up-regulation of TNF-α and IL-1β expression in response to local bacterial and malarial antigens, respectively. The brain parenchymal expression of TNF-α in the CM brain sections of the Ghanaian children we studied, collaborates a similar recent report in CM deaths in Malawian children [5]. Others have shown the up-regulation of neuronal TNF-α expression [25] and TNF-α expression in infiltrating meningeal leukocytes [24] in experimental animal models of meningitis, but we observed brain parenchymal TNF-α expression in the human PBM case studied. IL-1β, though present in CM, was not as high in intensity as TNF-α expression. Our observed parenchymal IL-1β protein expression in both the PBM and CM brain sections contrasts a previous report [5] that showed no staining for IL-1β in brains without CM infection, but IL-1β was only expressed using immunofluorescence on infiltrating leukocytes in PBM cases and not in the brain parenchyma [5]. We propose that this local proinflammatory cytokine release may be neurotoxic and contribute to ischaemic cell death in the brain, as previously suggested [26].Significantly, the endothelial activation in the case of Salmonella septicaemia was not as intense as that in the fatal malaria sections, and unlike the CM sections was not associated with local proinflammatory cytokine release. We, therefore, propose that the local presence of malarial antigens by way of sequestered PEs in the fatal malaria sections contributes to the more intense endothelial activation compared to systemic sepsis, and that the quantitatively more sequestered PEs in the cases of CM than SMA may account for the observed local release of proinflammatory cytokines in CM but not in SMA. Similarly, the local presence of bacterial antigens in the PBM case may account for the observed local release of proinflammatory cytokines in the PBM case. The lack of parasite adherence ligands (PfEMP-1) in the absence of malaria infection explains why sequestered PEs were not found in the Salmonella sepsis brain sections, despite the up-regulation of adhesion molecules. In the meningitis case, the observed increase in the expression of these proinflammatory cytokine expressions was not associated with an increased receptor expression.Generally, the expression of all the 6 antigens studied was maximal in the CM cerebellar sections. Others have observed maximal cerebellar cytokine [27] and adhesion molecules [28] in the P. coatneyi infected Rhesus monkeys. Additionally, our observed maximal expression of cytokines and adhesion molecules in the cerebellum in human CM correlates well with histopathologic observations of maximal sequestration of PEs in this region of the brain in both human CM [30] and Rhesus monkey model of CM [27–29]. The maximal up-regulation of adhesion molecules and cytokines in human CM cerebellar sections, though the other brain regions in all probability should be equally exposed to the elevated circulating proinflammatory cytokines in plasma and CSF, suggests that the maximal sequestration of PEs in this region of the brain is the trigger event for the local IL-1β and TNF-α expression. Most probably, parasite-derived factors resulting from sequestered PEs may have induced the local proinflammatory cytokine release. Our observed maximal cytokine and adhesion molecule induction and the previously reported maximal PE sequestration [30] and maximal up-regulation of RANTES and its corresponding receptors [38] in human CM cerebellar sections, correlates well with the documented cognitive impairment in Kenyan and Senegalese children surviving CM [41], since the cerebellum controls co-ordinated movement and some forms of cognitive learning [42].The current study, though limited by the absence of the assessment of peripheral blood for the 6 selected immunomodulator markers (sICAM, sVCAM, sE-Selectin, TNFα, IL-1β and TGFβ), suggests that local proinflammatory cytokine release may be a major immune mediator during human CM pathogenesis, particularly in the cerebellum. The mechanisms by which the induced proinflammatory cytokines in the brain in human CM mediate immunopathology is unclear. In conclusion, this study provides evidence to suggest that the induction of IL-1β and TNF-α in the brain in human CM is involved in further up-regulation of adhesion molecules, and may exacerbate the observed immunopathology associated with the disease, particularly in the cerebellum. Further studies currently in progress will reveal the specific role of these molecules in the exacerbation of CM. To our knowledge, this is the first report indicating maximal expression of adhesion molecules and cytokines in human CM cerebellar sections.(original magnification, X400): Immunohistology for ICAM-1, VCAM-1 & E-selectin. (A–B) Sections of non-malaria cases: (A) Staining for ICAM-1: positive in < 25% of brain microvessel (average score +/−). (B) Staining for VCAM-1: positive in < 25% of brain microvessels (average score +/−). (C) Section of typhoid perforation and septicaemia case: Staining for ICAM-1: positive in 25 to 50% of microvessels (average score +). (D–F) Sections of malaria cases: (D) Staining for ICAM-1: positive in > 75% of microvessels (average score +++). (E) Staining for ICAM-1: positive in > 75% of microvessels (average score +++). Note the more extreme heterogeneity between sequestration and receptor expression in the SMA case in 1E than the CM case in 1D. (F) Staining for ICAM-1: Positive in > 75% of microvessels (average score +++). Note the more intense staining of the CM cerebellar section in 1F compared to the CM cerebral section in 1D.(original magnification, X100): Immunohistology for TGFβ, IL-1β & TNFα. (A–B) TGFβ staining (intravascular & perivascular). (A) NCNSI section: moderate in some fields (average score +). (A) CM/PBM section: strong in all fields (average score +++). (C–E) IL-1β staining (intravascular, perivascular & parenchymal). (C) PBM section: moderate in some fields (average score +). (D) CM section: strong in all fields (average score +++). (E) NCNSI section: no staining in all fields (average score −). TNFα staining (intravascular, perivascular & parenchymal). CM/PBM section: strong in all fields (average score +++).Clinical and Diagnostic Details of the 21 Studied Cases.Pos: Parasitized erythrocytes adhering to cerebral microvessels; Neg: No malaria parasites or pigment in cerebral microvessels.SBP: Severe Bronchopneumonia with no bacterial growth after 7 days of incubation of blood culture.HSC: Haemolytic Sickle Cell crisis;BDU: Bleeding Duodenal Ulcer;PBM: Purulent Bacterial Meningitis.TP&S: Typhoid Perforation and Septicaemia with Salmonella typhi isolated in blood culture after 48 hours of incubation.NPS: Nephrotic Syndrome.Quantitation of Cerebral Vessels and Co-localization of Sequestration with Expression of Receptors in the 15 Malaria Cases.R+S+: Number of vessels showing both receptor expression and sequestration; R+S−: Number of vessels showing receptor expression but no sequetration; R−S+: Number of vessels showing no receptor expression but sequestration; R−S−: Number of vessels showing neither receptor expression nor sequestration.
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+ Inorganic ions, coenzymes, amino acids, and saccharides could co-exist with toxic environmental chemicals, such as polycyclic aromatic hydrocarbons (PAHs), in the cell. The presence of these co-existing chemicals can modulate the toxicity of the PAHs. One of the genotoxic effects by PAHs is light-induced cleavage, or photocleavage, of DNA. The effect of inorganic ions I−, Na+, Ca2+, Mg2+, Fe3+, Mn2+, Cu2+, and Zn2+ and biological molecules riboflavin, histidine, mannitol, nicotinamide adenine dinucleotide (NAD), glutathione, and glutamic acid on the DNA photocleavage by pyrene, 1-hydroxypyrene (1-HP), and 1-aminopyrene (1-AP), is studied. The non-transition metal ions Na+, Ca2+, and Mg2+, usually have very little inhibitory effects, while the transition metal ions Fe3+, Cu2+, and Zn2+ enhance, Mn2+ inhibits the DNA photocleavage. The effect by biological molecules is complex, depending on the photochemical reaction mechanisms of the compounds tested (1-AP, 1-HP and pyrene) and on the chemical nature of the added biological molecules. Riboflavin, histidine, and mannitol enhance DNA photocleavage by all three compounds, except that mannitol has no effect on the photocleavage of DNA by pyrene. Glutathione inhibits the DNA photocleavage by 1-AP and 1-HP, but has no effect on that by pyrene. NAD enhances the DNA photocleavage by 1-AP, but has no effect on that by 1-HP and pyrene. Glutamic acid enhances the DNA photocleavage by 1-AP and pyrene, but inhibits that by 1-HP. These results show that the co-existing chemicals may have a profound effect on the toxicity of PAHs, or possibly on the toxicity of many other chemicals. Therefore, if one studies the toxic effects of PAHs or other toxic chemicals, the effect of the co-existing chemicals or ions needs to be considered.Polycyclic aromatic hydrocarbons (PAHs) are widespread environmental pollutants produced during forest fire, volcanic eruption, incomplete burning of fossil fuels, petroleum products, as well as during tobacco smoke, food processing, operation of machinery including automobiles, airplanes and ships [1,2]. Chemical carcinogenesis studies of PAHs started in 1915 when Yamagawa and Ichikawa observed that repeated application of coal tar on rabbits over extended periods induced skin carcinomas on the rabbit ears [3]. The carcinogenic potential for PAHs was recognized in 1930 when Kennaway and Hiegar synthesized dibenz[a,h]anthrancene and determined its carcinogenicity [4]. However, the fluorescence spectrum of dibenz[a,h]anthrancene did not correspond exactly to that of the carcinogenic components of coal tar. Further investigation of other carcinogenic components of coal tar by Cook et al led to the discovery of benzo[a]pyrene as one of the major carcinogenic components [5]. Since these pioneering studies, more than 30 PAHs and several hundred of their derivatives have been reported to exhibit some carcinogenic effects [6–8]. It is reported in the 8th Report on Carcinogens that exposure to PAHs has been linked to the development of skin and lung cancers (National Toxicology Program, 1998).Under physiological conditions, PAHs themselves are inert molecules and generally considered nontoxic toward biological systems. PAHs affect humans usually after being activated by metabolic enzymes or light [9,10]. Many PAHs can be metabolized into diol epoxides, diones and other reactive intermediates that are capable of reacting with cellular DNA to form PAH-DNA covalent adducts or cause other forms of cellular damages [2,11–13]. Another pathway for PAH activation is photo-activation [10]. PAHs are activated by absorption of light energy and excited to their upper energy states. The excited state energy can be lost by emitting light or heat, or transferred to molecular oxygen, solvent molecules, or biological molecules in the cell to generate reactive intermediates that can damage cellular constituents such as cell membrane, nucleic acids, or proteins. It has been observed that PAHs can cause DNA single strand cleavage and DNA-PAH adduct formation upon UVA light irradiation [10, 14–17]. Generally, PAHs are more toxic when exposed to simulated solar radiation than if it is kept in the dark and the increase in toxicity can exceed 100-fold [18, 19]. Thus, the light-activated PAHs can cause cellular damages and exert toxicity including carcinogenicity. Since co-existing chemicals are usually involved in the photochemical processes of the phototoxic compounds, it is understandable that the co-existing chemicals should have a profound effect on the photochemical reaction as well as the phototoxic effects of these compounds.Through various pathways PAHs are known to be able to enter into the cells and co-exist with many other biological chemicals in the cell. It is logical to assume that the co-existing chemicals should have a profound effect on light-induced DNA cleavage by PAHs, and therefore, on the toxicity of these compounds if the cells, such as skin cells, are subject to light irradiation. Thus, we wish to report the effect of biologically relevant inorganic ions and molecules on the light-induced DNA cleavage by pyrene and its polar derivatives, 1-aminopyrene (1-AP) and 1-hydroxypyrene (1-HP). Pyrene was chosen as a representative PAH and 1-AP and 1-HP as its polar derivatives. The biologically relevant ions, Na+, K+, Mg2+, Ca2+, Fe3+, Cu2+, Mn2+, Zn2+, and I− and molecules, riboflavin, nicotinamide adenine dinucleotide (NAD), histidine, mannitol, glutathione, and glutamic acid, are chosen to study their effect on the light-induced DNA cleavage.Pyrene, 1-AP, and 1-HP were purchased from Aldrich Chemical Company (Milwaukee, WI) and used without further purification. Stock solutions (1 mM) were prepared in methanol and stored in brown containers in the refrigerator to exclude light. It was diluted with other solvents necessary to make the working solution before use. ΦX 174 phage DNA (supercoiled RF-1 or sc-DNA) with a molecular weight of 3.6 × 106 Da and 5386 base pairs was purchased from Promega Corporation (Madison, WI) and stored at −20°C. Ethidium bromide (EB), bromophenol blue, xylencyanol, sodium chloride (NaCl), potassium iodide (KI), magnesium chloride (MgCl2), manganese chloride (MnCl2), zinc chloride (ZnCl2), cupric chloride (CuCl2), ferric chloride (FeCl3), riboflavin, histidine, mannitol, NAD, glutathione, and glutamic acid were purchased from Sigma-Aldrich (St Louis, MO). Agarose, calcium chloride (CaCl2), sodium phosphate, TRIS-base, boric acid, and EDTA were purchased from Fisher Scientific (Houston, TX). All solvents used were spectroscopic grade. The water used (18 MΩ) was deionized by a Barnstead Nanopure Infinity water de-ionization system (Dubuque, Iowa). A rapid agarose gel electrophoresis apparatus (C.B.S & Scientific Co.) was used for gel electrophoresis. The NucleoVison Gel-Documentation System (NucleoTech Inc., CA) was used for quantification of the DNA.Solutions (a total of 60 μl of 10 mM sodium phosphate at pH 7.1 with 10% methanol) containing ΦX 174 phage DNA (27 μM in base pairs), 0.6 μM 1-HP (or 6 μM 1-AP or 60 μM pyrene), and a given biological ion or molecule were filled into the wells of a 3 × 8 flat bottomed TitertekTM plate (ICN Biochemical). These PAH concentrations were chosen to cause about 30% of DNA photocleavage under these experimental conditions. The plate was tightly covered with glass and placed onto a Pyrex glass support/filter, which was placed on an O-ring secured on a ring stand. The Pyrex glass served as a light filter to efficiently cut off any light below 300 nm that could damage DNA. A 100 W UVA lamp (type B, UVP Inc., Upland, CA) was placed beneath the Pyrex glass and the light was applied through the bottom of the Titertek plate from a fixed distance of 6.0 cm. The UVA intensity of the light output was measure to be 170 J/cm2 per hour of irradiation (UVA detector, Model PMA 2100, Solar Light Co., Inc., Philadelphia, PA). A stream of cold air was blowing toward the bottom of the Pyrex glass during the irradiation period to eliminate any heat. After irradiation, 12 μl of a gel-loading dye solution (bromophenol blue and xylencyanol in 50% glycerol) was added into each well of the Titertek plate and mixed. Then 14 μl of the sample was loaded into the wells of the pre-prepared 1% agarose gel and subjected to electrophoresis at 100V for 1–2h. Following electrophoresis, the gel was stained with Ethidium bromide (2mg/L) and analyzed with the Gel-Documentation System software. In the gel, there were two clear bands with the supercoiled DNA (sc-DNA) being the band further away from the origin and the relaxed open-circular DNA (oc-DNA) being the closer to the origin. The amount of sc-DNA and oc-DNA were quantified by the total fluorescence intensity of the bands after subtracting a common background as described in a previous publication [16].DNA photocleavage by 1-AP, 1-HP, or pyrene was carried out in the presence of biologically relevant inorganic ions. The choice of chloride salts is to eliminate any effect from the anion, because Cl− seems to have minimal effects on the light-induced DNA cleavage. All experiments were carried out with 27μM DNA and the results are shown in Figure 1 for pyrene and Table 1 for all three compounds. Lanes 1, 2, 3 in Figure 1 are the negative controls: DNA alone without light irradiation (lane 1), DNA alone with light irradiation (lane 2), and DNA + 60 μM pyrene without irradiation (lane 3). They all have the same percent of oc-DNA indicating that exposure to light alone (lane 2) and pyrene alone (lane 3) does not cause any cleavage to the DNA. Lane 4 is the positive control with DNA and 60 μM pyrene with 1h of irradiation. It shows that about 26% of the sc-DNA is cleaved. The tests for each ion are with dark control (ion + pyrene + DNA without irradiation) and light control (ion + DNA with irradiation) in addition to the regular test (ion + pyrene + DNA with irradiation). The dark control experiments confirm that all ions except Fe3+ cause only minimal DNA cleavage without light (lanes 5, 7, 9 in Figure 1A and 5, 7, 9, 13 in Figure 1B). Fe3+ itself can cause the sc-DNA to become oc-DNA at a concentration of 0.5 mM (lane 11 in Figure 1B). Light control with ions (without pyrene) shows the same trend as that for the dark control for most ions (except Fe3+) that they do not cause any cleavage to sc-DNA with 1 h of irradiation (lanes 11, 12, 13 in Figure 1A and 15, 16, 17, 19 in Figure 1B). Among all ions tested, Ca2+, Mg2+, Na+, and Mn2+ have either minimal effects or cause a decrease of DNA photocleavage (Figure 1A and B), whereas Zn2+ and Cu2+ enhance the DNA photocleavage by pyrene. The effect by Fe3+ cannot be assessed because Fe3+ itself can cause DNA cleavage.Table 1 lists the percent of DNA photocleavage by pyrene, 1-AP, and 1-HP in the presence of various ions. The effect of most of the ions tested on 1-HP and 1-AP induced DNA photocleavage has the same trend as that for pyrene. Main group ions Ca2+, Mg2+and Na+ have either minimal effect or cause a small decrease. This indicates that the effect by these ions on the DNA photocleavage is negligible. Whereas transition mental ions, except for Mn2+ which inhibits DNA photocleavage, Zn2+ and Cu2+ can cause an enhancement for the DNA photocleavage by all three compounds. In the presence of Cu2+, sc-DNA is completely cleaved due to the exposure to pyrene or 1-HP and light (Table 1 and bar 10 in Figure 1B). It is only slightly enhanced, however, by the exposure to 1-AP and light (Table 1). Zinc ion enhances and manganese ion inhibits the DNA photocleavage by all three compounds. The mechanism of effect of each transition metal on DNA photocleavage still is not clear. But it has been reported that semicarbazide induces DNA damage in the presence of Cu2+ through the formation of hydrogen peroxide and semicarbazide-derived free radicals [20]. Thus, it is assumed that Cu2+ enhances the DNA photocleavage through Cu2+ or PAH-derived free radicals. Although effect by Fe3+ cannot be assessed, in the presence of UV radiation, Fe3+ species also can undergo a photoredox process giving rise to Fe2+ and OH• hydroxyl radical [21] according to:OH• hydroxyl radical attacks and cleaves the DNA phosphate deoxyribose backbone in a largely sequence-independent manner [22].However, the effect by I− is different for 1-AP from 1-HP and pyrene, where I− enhances the DNA photocleavage by 1-AP and inhibits the DNA photocleavage by 1-HP and pyrene. This inhibitory effect on 1-HP or pyrene induced DNA photocleavage is in agreement with that I− inhibits the DNA photocleavage by 5- and 7-methylbenz[a]anthracenes [15]. It is known that iodide ion is an excited singlet state quencher by enhancing the intersystem crossing rate from an excited singlet state to an excited triplet state [23]. Therefore, the fact that the presence of I− quenches the DNA photocleavage by either 1-HP or pyrene indicates that the singlet excited state of 1-HP or pyrene is involved in the DNA photocleavage process. Since the presence of KI enhances the DNA photocleavage by 1-AP, it must involve a different mechanism such as the involvement of the triplet state of 1-AP.DNA photocleavage induced by the combination of 1-AP, 1-HP, or pyrene and light were carried out in the presence of some biologically important molecules. The rationale is that PAHs in the cell are likely to co-exist with these molecules and these molecules may affect the ability of PAHs to cause DNA photocleavage. Except riboflavin, which by itself can cause DNA photocleavage, all the other chemicals, histidine (5mM), mannitol (0.5mM), NAD (5mM), glutathione (0.5mM), and glutamic acid (0.5 mM) do not cause DNA cleavage upon light irradiation under these experimental conditions. The effect of these chemicals on DNA photocleavage is different for 1-AP, 1-HP or pyrene (Figure 2, Table 2). With 1-AP, all chemicals except glutathione, which inhibits the DNA photocleavage completely, enhance DNA photocleavage. For 1-HP, riboflavin, histidine and mannitol enhance DNA photocleavage, while glutathione and glutamic acid inhibit and NAD has no effect on DNA photocleavage. For pyrene, riboflavin, histidine and glutamic acid enhance DNA photocleavage, while NAD, glutathione, and mannitol have no effect on DNA photocleavage.The varying effects seen here should be due to differences in chemical/photochemical reactions these three compounds initiate that lead to DNA photocleavage.Riboflavin is a known photo-sensitizer that can facilitate light-induced chemical reactions of other compounds. The photodynamic action of riboflavin is generally considered to involve the generation of reactive oxygen species [24]. Photo-excited riboflavin gives rise to oxidative DNA damage predominantly by a type-I photoreaction, i.e. by direct one-electron or hydrogen transfer [25]. ESR experiments suggested that photo-excited riboflavin reacts with dGMP to produce riboflavin anion radical and guanine cation radical, but not with other mononucleotides. The estimated ratio of 8-OH-dG yield to total guanine loss indicates that the photo-excited riboflavin induces 8-OH-dG formation specifically at guanine residues located 5′ to another guanine [26]. As a result, riboflavin can cause DNA photocleavage by itself on one hand, and it can enhance the ability of all the three compounds to cause DNA photocleavage, on the other.Glutathione is a known reducing agent that usually traps free radicals such as hydroxyl free radicals [27]. Therefore, glutathione’s role here seems to be as a free radical scavenger. It has been reported that both histidine and mannitol are singlet oxygen quenchers by reacting with singlet oxygen to produce oxidation products of histidine. Both should inhibit the DNA photocleavage. However, the presence of 5mM histidine and 0.5mM mannitol enhances the DNA photocleavage for all three compounds. This indicates that histidine and mannitol play roles other than quenching singlet oxygen. It has been confirmed that the presence of histidine greatly lengthens the degradation half-lives of various PAHs and inhibits their degradation [14]. Thus, the presence of histidine maintains a higher concentration of the PAHs that are capable of generating species other than singlet oxygen, such as superoxide or PAH free radicals, to cause DNA single-strand cleavage. The effect by mannitol is still not clear and needs further investigation.PAHs are ubiquitous environmental carcinogens and may co-exist with the biologically relevant ions and molecules in the cell. Since it is known that PAHs can cause light-induced DNA cleavage [15–17], they are photomutagenic toward Salmonella typhimurium bacteria strain TA102 [28], and genotoxic to human skin cells (unpublished results), the effect of co-existing chemicals on the light-induced DNA cleavage of PAHs is of interest to human health. It has shown in this paper that the effect by the co-existing molecules on the light-induced DNA cleavage is complex. It depends on several factors:Reactive species generated by photosensitization of a certain PAH compound leading to DNA photocleavage. These species include, but not limited to, singlet oxygen, superoxide, hydroxyl radical, PAH or DNA free radicals, and PAH or DNA adducts [10]. DNA cleavage can be caused by one or by a combination of several of these intermediates.Nature of the co-existing chemicals. These chemicals can act either as scavengers of any of the aforementioned reactive species or as promoters of any of the intermediates that generates the reactive species. Because of these modifying effects, the phototoxicity of PAHs in cells should be studied in consideration of the co-existing chemicals.Effect of biologically relevant inorganic ions on the light-induced DNA cleavage by pyrene (60 μM). Lanes 1, 2, 3 are negative controls and lane 4 is the positive control. The lower bands in the gel are that of the sc-DNA and the upper bands are that of the oc-DNA. The concentrations for various ions are: 50 mM for I−, Na+, Ca2+, Mg2+ and 0.5 mM for the transition metals Mn2+, Zn2+, Fe3+, and Cu2+. Experiments with all ions are under three conditions: (1) Dark control: Ion + DNA + pyrene without irradiation (Lanes 5, 7, 9 in A and 5, 7, 9, 11, 13 in B); (2) Effect of ion on DNA cleavage by pyrene: ion + DNA + pyrene with 1 h of irradiation (Lanes 6, 8, 10 in A or 6, 8, 10, 12, 14 in B); (3) Light control: ion + DNA + H2O with 1 h of irradiation: (Lanes 11, 12, 13 in A or 15, 16, 17, 18, 19 in B).Effect of biologically relevant molecules on light-induced DNA cleavage by pyrene. ΦX-174 plasmid DNA (27μM in base pairs) was mixed with 60μM pyrene and various biological chemicals and was irradiated for 1 h with a 100 W UVA lamp. Lane 1 is the dark control and lane 2 is the positive control with pyrene and DNA but without any added chemicals. Lanes 3, 5, 7, 9, 11, and 13 are the mixtures of DNA and pyrene irradiated for 1 h in the presence of riboflavin (5 mM), NAD (5mM), Histidine (5mM), Glutathione (0.5mM), Mannitol (0.5mM), and Glutamic acid (0.5mM), respectively. Lanes 4, 6, 8, 10, 12, 14 are the same as 3, 5, 7, 9, 11, and 13, respectively, but without pyrene.Effect of biologically relevant ions on the percent of DNA photocleavage by 1-hydroxypyrene (1-HP), 1-aminopyrene (1-AP), and pyrene.The error for each sample is about ±20%. Fe3+ can cause DNA cleavage without 1-HP, or pyrene, or light. The concentrations for various ions are: 50 mM for I−, Ca2+, Mg2+, Na+ and 0.5 mM for transition metals Mn2+, Zn2+, Fe3+, and Cu2+. The concentration for 1-AP, 1-HP, and pyrene are: 6, 0.6, and 60 μM, respectively.Numbers represent enhancement on the DNA photocleavage.Percent of DNA photocleavage by 1-AP, 1-HP, and pyrene in the presence of biologically relevant moleculesThe numbers are the percent of DNA cleavage caused by 1-AP, 1-HP, or pyrene upon light irradiation in the presence of one of the chemicals. The standard error is ±20%.Numbers represent an enhancement.This research was in part supported by the National Institutes of Health: NIH SCORE S06 GM08047 and the US Army Research Office DAAD 1901-1-0733 to JSU. We thank NIH-RCMI for core molecular and cellular biology and analytical facilities established at JSU.
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+ Chemical carcinogenesis studies are powerful tools to obtain information on potential mechanisms of chemical factors for malignancies. In this study Western blot analyses, using monoclonal antibodies specific for three different cytochrome P450 (CYP) isozymes (CYP1A1, CYP1A2 and CYP2B), were employed to examine the effect(s) of 3-methylcholanthrene and/or pristane (2,6,10,14-tetramethylpentadecane) on the basal and inducible levels of expression of CYP proteins within Copenhagen rat tissues. Pristane exposure led to tissue specific differences in the CYP isozymes expressed and elicited increased CYP protein expression over 3-methylcholanthrene induced levels in microsomes isolated from liver, Peyer’s Patches, and thymus. Within the context of the chemical carcinogenesis model employed in this study, these observations correlated with the induction of B-cell malignancies by low doses of 3-methylcholanthrene and of thymic lymphomas by a high 3-methylcholanthrene dose. The data suggest that pristane treatment affects CYP isozyme expression. This pristane-mediated effect clearly could be a contributing factor in the chemical carcinogenesis of the previously observed lymphoid malignancies, and a possible basis for the tumor enhancing effects of pristane.Pristane is a ubiquitous isoprenoid which has been shown to induce B-cell plasmacytoma, following intraperitoneal administration into genetically susceptible BALB/cAnPt mice [1–6]. Pristane treatment altered CYP1A activity and polyamine regulation [7], and has been shown to affect several biological responses including changes in membrane fluidity, chromatin conformation and histone protein expression [8]. In addition, pristane exhibited tumor enhancing properties in our Copenhagen rat model in which direct Peyer’s patches (PP) injection with a low dose of 3-methylcholanthrene (3-MC) led to only B lymphoid malignancies, a high 3-MC dose treatment induced thymic lymphomas but no B cell malignancies, and co-treatment with 3-MC and pristane lead to higher frequency and decreased latency of 3-MC induced lymphoid malignancies [9]. Also, pristane elicited marked effects on transactivation of transfected genes [10–13]. These activities are comparable to 12-O-tetradecanoylphorbol 13-acetate (TPA), a known tumor promoter [14]. However, based on the disparity of the results obtained with TPA and pristane in protein kinase C substrate phosphorylation and activation of the chloramphenicol acetyltransferase genes [10] it appears that TPA and pristane exhibit different tumor promotion mechanism(s) of action in chemical carcinogenesis. TPA induces CYP1A and simultaneously suppresses CYP2B [15], but the mechanism of CYP1A induction by TPA appears to differ from other TPA inducible functions, as well as from those elicited by 3-MC [11–13]. TPA intercalates into DNA and modifies histone phosphorylation, thus altering the super-structure of chromatin [14]. On the other hand, pristane is absorbed from the gut contents, disseminates through the body, and the majority of pristane is localized to the plasma membranes [16].Membrane perturbation by pristane possibly leads to alterations in regulation of many different cellular products including cytochrome P450 (P450) isoforms, which are also membrane associated. Using pristane and the carcinogen 3-MC [17], which induces P450 isozymes (products of the CYP genes), we examined the effect(s) of pristane on CYP isozyme expression within the suspected target organs. Not only are CYP isozymes specifically regulated [18], they are inducible [19–30] and catalyze the biotransformation of both exogenous and endogenous compounds to highly reactive intermediates which can cause toxicity and/or carcinogenicity [25–30]. We undertook these studies in light of the fact that pristane influences CYP1A activity in a 3-methylcholanthrene (3-MC) dose dependent manner [7], leading to speculation as to whether or not tissue specific expression of CYP protein correlated with CYP1A activity in the preferential induction of lymphoid malignancies.The following commercially available chemicals were employed in this study: 3-methylcholanthrene (3-MC) and β-mercaptoethanol from Eastman Kodak Company (Rochester, NY); 2,6,10,14-tetramethylpentadecane (pristane), sunflower seed oil, bovine serum albumin, sodium dodecyl sulfate (SDS), ammonium persulfate, N,N,N′,N′-tetramethylethylenediamine (TEMED), nitro blue tetrazolium (NBT) tablets, 5-bromo-4-chloro-3-indolyl phosphate (BCIP), trizma base, Coomassie blue R250, bromphenol blue, α-nitrocellulose membrane, acrylamide, bis-acrylamide, glycine, glycerol, goat serum from Sigma Chemical Company (St. Louis, MO); protein assay reagent from Biorad Laboratories (Richmond, CA); and sodium phenobarbital (PB) from J.T. Baker Chemical Company (Phillipsburg, NJ). All other inorganic salts and solvents used were of analytical grade. Alkaline phosphatase conjugated and horseradish peroxidases labelled secondary antibodies were obtained from Southern Biotechnology Assoc., Inc. (Birmingham, AL). Monoclonal antibodies (MAbs) specific for CYP1A1, CYP1A2, and CYP2B isozymes have been reported [31–32].Female Copenhagen rats (8–10 weeks old) were initially obtained from the National Cancer Institute, Frederick, MD., and were subsequentially bred and maintained under NIH guidelines. Animals were provided with a commercial diet (Harlan Teklan Laboratory, Madison, WI.), and water ad libitum. All rats were anesthetized prior to injections, surgery or euthanasic procedures. The Copenhagen rat strain was selected due to its low rate of spontaneous tumor development [33]. Where indicated, pristane was administered as a 1 ml, intraperitoneal (i.p.) injection either alone, 2 weeks prior, or 4 weeks prior to either animal sacrifice or 3-MC treatment (See Table 1). Age matched rats were separated into specified groups containing six animals per group. Group 1 (BASAL) rats were normal, untreated controls (H20 or SSO treated rats served as controls as well, and yielded similar results). Group 2 rats were treated with only pristane and were sacrificed either two weeks or four weeks after pristane priming. Group 3 rats were treated with only 3-MC (3, 12 or 24 hrs; directly injected into Peyer’s patches (PP) with 0.05μg, 5μg, or 500μg of 3-MC) as previously described [9]. The 3-MC, dissolved in sunflower seed oil (SSO), was injected into the five PP which were most distal on the small intestine. Within group 4, rats received 3-MC and pristane (2 wks) co-treatment.Data obtained from Western analyses of tissue samples (group 3 – 4 rats) employed 3-MC doses which were carcinogenic in our rat tumor model. However, these treatment protocols generated expression levels of CYP which were at the lower limits of detection in PP and thymus (THY). Therefore, in order to further examine the effects of pristane on CYP isozyme expression in PP and THY, protocols using very high doses of inducers were employed. To maximally induce CYP1A1 and CYP1A2 [19, 24], 3-MC (25mg/kg/day for 4 consecutive days; M VH) dissolved in SSO was administered i.p. within group 5 rats. Phenobarbital (PB), used as a control, (75mg/kg/day for 4 consecutive days) was dissolved in H2O and injected i.p. into group 6 rats, as an efficient means to induce CYP2B1 and 2B2 [30, 32]. It should be noted that animals were injected and sacrificed in the morning whenever possible in order to minimize possible variations from circadian rhythms. All animals were sacrificed one day after the last dose of pristane, 3-MC, PB or control vehicle treatment.Microsomal fractions were prepared from LIV, PP, and THY, of rats according to a previously described method [7, 34]. As indicated, tissues were initially rinsed in ice cold 1.15% KCl and homogenized in four volumes of a buffer containing 0.1 M KCl, 0.1 M Tris-acetate, and 1 mM EDTA (pH 7.4), using a Tekmar Tissumizer and sonified at 50% power for 4 consecutive 15 sec bursts. Homogenates were centrifuged at 10,000 × g for 20 min at 4°C. The supernatants were removed and centrifuged at 100,000 × g for 1 hr at 4°C. The microsomal pellets were then suspended in a buffer containing 0.1 M Na2H2P2O7 and 1 mM EDTA (pH 7.4), and frozen at −80°C until used for Western blot analyses.Rat microsomes from each of the treatment groups (group 1 – group 6) were prepared and analyzed by Western blots. Each microsomal preparation and each Western blot analysis was repeated in four different experiments in order to statistically compare and evaluate each band such that even marginal changes in band intensity represented effects of the inducer.For determinations of levels of CYP protein expression, samples were prepared in Laemmli buffer (Invitrogen) consisting of 125 mM Tris-HCl (pH 6.8), 10 mM DTT, 17.4% glycerol, 3% SDS, 0.025% bromophenol blue, and 10mM β-mercaptoethanol. Samples were loaded into wells at the indicated concentration of microsomal protein (a concentration of 20μg of microsomal protein/track, with the exceptions of liver microsomes from very high dose 3-MC and PB treated rats which were loaded at 0.2–2μg/track and thymic samples which were loaded at 200μg/track). Total protein concentrations were determined by the methods of Bradford and Lowry, using the Bio-Rad Protein Assay reagents. Based on chemical assays for protein (Biorad) and densitometric scans of Coomassie blue stained gels (see below), the total amount of protein loaded onto each gel lane was within 10% of the average load per well.SDS-gel electrophoresis was performed according to Laemmli using 16cm long gels with a 12% resolving gel with a 4% stacking gel for the resolution of microsomal proteins. For each experiment, there were four identical gels run overnight, of which three were electrophoretically transferred to nitrocellulose and used for Western blot analyses, and one gel was used for staining with Coomassie blue dye. The nitrocellulose filters were probed with MAbs specific for CYP1A1 or CYP1A2 (3-MC inducible isozymes), as well as CYP2B (PB inducible CYP isozymes), including CYP2B1 and CYP2B2 which share 96% sequence homology [32], as previously described, followed by alkaline phosphatase conjugated second antibody probing and development with NBT and BCIP. Molecular weight markers were included on each gel in order to identify the expected area of migration of the CYPs (CYP1A1, 55 or 56,000 kDa; CYP1A2, 52,000 kDa; and CYP2B1 and 2B2, 51,000 kDa and 52,000 kDa, respectively) under these conditions.Electrophoresed proteins were transferred to nitrocellulose filter paper by electroblotting overnight at 15 V in pre-chilled transfer buffer containing 25 mM Tris, pH 8.3, 192 mM glycine, and 20% methanol. Filters were blocked with non-fat dry milk or bovine serum albumin, dependent upon recommended conditions for each monoclonal antibody. Following rinsing, membranes were probed with primary antibody, washed, incubated with HRP-conjugated secondary antibody and developed by enhanced chemiluminenscence.Relative densities (OD/mm2) of Coomassie stained gels or developed immunoblots were quantified with a densitometer interfaced with a computer equipped with the ImageQuant Software package (Molecular Dynamics Inc., Sunnyvale, CA). The densitometer generated a general graphics file which was transferred to Microsoft Excel graphics programs for Windows (Microsoft Corp., Redmond, WA) and analyzed using volume integration parameters (1 pixel = .007744 mm2).Within hepatic samples from untreated rats (Group 1), no CYP1A1 was detected, although basal levels of CYP1A2 and CYP2B proteins (note two bands since our MAb probe specific for CYP2B detects both the CYP2B1 and the CYP2B2 isozymes) were apparent. However, pristane treatment (Group 2) affected the expression of the latter mentioned CYPs without affecting CYP1A1 (see Fig. 1). This effect was best demonstrated with 20μg and 2μg of microsomal protein for CYP1A2 and CYP2B, respectively. Based on the densitometric scans of these profiles pristane elicited 1.7 and 1.4 fold increases in CYP1A2 and 1.6 and 1.5 fold increases in CYP2B after 2 and 4 weeks post pristane treatment, respectively.Representative data obtained from LIV samples isolated from rats which were either untreated (Group 1), treated with a single i.p. injection of only pristane (Group 2) or treated with a very high dose of either 3-MC or PB with or without the pristane treatment are presented (Fig. 2). It is important to note that only 0.2μg of microsomal protein per lane was loaded onto the gel used to generate this immunoblot. This was 10× less than the lowest amount (2μg) depicted in Fig. 1 (i.e. basal levels of CYP1A2 protein were below detection limits within this particular gel since at least 20μg of microsomal protein was needed). However, following very high dose treatment with 3-MC (Group 5), not only were appreciable levels of CYP1A2 detected, but also CYP1A1; the increase in CYP1A2 represented a >41 fold increase over basal control levels whereas the amount of CYP1A1 increased at least 1000 fold (assuming >200μg of microsomal protein was necessary to detect basal levels; see Fig. 1, i.e. no bands with CYP1A1). As expected, neither CYP1A1 nor CYP1A2 were elevated after very high dose PB treatment (Group 6), whereas CYP2B was clearly induced (24 fold greater than basal, based upon densitometric scans). Again, pristane affected the responses, although in a somewhat different fashion. Specifically, pristane treatment (2wk or 4wk) led to decreased expression of CYP1A1 protein (1.4 and 1.2 fold, respectively), CYP1A2 protein (1.3 and 1.1 fold, respectively) and CYP2B protein (2.9 and 1.5 fold, respectively) after very high dose treatment with 3-MC, as well as decreases in CYP2B protein (1.3 and 1.2 fold, respectively) after very high dose treatment with PB.The effects of carcinogenic doses of 3-MC (within the animal model employed) on hepatic CYPs were also examined. A representative immunoblot depicting the effects of 3-MC, 3 and 12hr post injection into the PP of either unprimed (Group 3) or pristane primed (Group 4) rats are depicted in Fig. 3. Note that the amount of LIV microsomal protein was 100 fold more than that employed in Fig. 2; hence the basis for detection of CYP1A2. Again, CYP1A1 levels were below detectable levels. Also, as previously observed, pristane treatment led to increased expression over basal levels of CYP2B. However, increased expression over constitutive levels of CYP1A2 in pristane treated rats (without 3-MC exposure) was detected by employing this amount of protein. In addition, increased amounts over untreated control levels of CYP1A2 were observed after 3-MC exposure. Unexpectedly, less CYP1A2 appeared to be present in the LIV 3hr after PP injection with 500μg than with 0.05 or 5μg of 3-MC. Upon co-treatment with pristane, CYP1A2 levels (when compared to hepatic levels at the respective times from rats not treated withpristane): 1) decreased after 3hr and increased after 12hr at the 0.05μg 3-MC treatment, 2) decreased after 3 and 12hr at the 5μg 3-MC treatment, and 3) increased after 3hr, but decreased after 12hr at the 500μg 3-MC treatment. With respect to CYP2B, pristane co-treatment led to levels which were increased after 3 and 12hr at the 0.05μg 3-MC treatment, but decreased after 3 and 12hr at the 5μg and 500μg 3-MC treatments.PP were also examined with respect to the constitutive expression (Group 1) of the CYPs, as well as CYP protein levels after exposure to pristane (Group 2) and/or 3-MC treatment with very high dose (Group 5) or carcinogenic (Groups 3 and 4) protocols. Neither CYP1A1, CYP1A2, nor CYP2B was detected within the PP of untreated rats (Group 1), nor were any of these 3 isozymes of CYP detected following very high dose 3-MC (Group 5) or PB (Group 6) treatment. Furthermore, treatment with only pristane (Group 2) did not elicit detectable levels of the CYP1A isozymes. However, PP injection with 500μg of 3-MC (Group 3) resulted in increased levels of CYP1A2, but not CYP1A1, without affecting basal levels of CYP2B (see Fig. 4). This was the only dose of 3-MC which elicited appreciable levels of CYP1A2 within PP. Note that the maximum increase in CYP1A2 protein was 12hr after PP injection with 3-MC; a decrease was observed 24hr post exposure. Co-treatment with pristane (Group 4) elicited increased expression of CYP1A2 3hr post 3-MC injection of the PP, although no apparent differences in the untreated versus pristane treated rats injected with 500μg of 3-MC were subsequently observed in microsomal protein samples prepared either 12hr or 24hr post 3-MC treatment.With respect to the THY (see Fig. 5), the salient points were as follows: First, neither CYP1A1, CYP1A2 nor CYP2B was detected in THY from either normal rats (Group 1), rats treated with only pristane (Group 2) or rats treated with the carcinogenic doses of 3-MC (0.05, 5 or 500μg) with (Group 3) or without pristane co-treatment (Group 4). In fact, the only method to elicit detectable amounts of a CYP within THY was by very high dose 3-MC treatment (Group 5), in which case CYP1A2 was the only CYP detected; pristane treatment elicited an increased amount of the CYP1A2 protein. However, very high dose PB treatment (Group 6) did not elicit the expected induction of CYP2B1 or CYP2B2. Furthermore it should be noted that 200μg of microsomal protein was used for the thymic immunoblots (i.e. the maximum protein load for the SDS-PAGE employed), which illustrates differences in the tissue specific expression of the CYP isozymes examined in this study.Densitometric scans were employed in these studies for two purposes. One was to assure that equivalent amounts of protein were applied to each gel lane, and the other was to statistically validate any possible differences which were detected on subsequent immunoblots. One representative gel with samples from each different tissue is shown (Fig. 6), but for each group examined, four gels were run, with similar patterns emerging.Based on densitometric scans, variability in the amount of protein present in each gel lane was ≃10% of the mean value obtained from all respective samples on a particular gel. Regardless of treatment, the scans substantiate that the evaluations made in this study represent effects of the chemical treatment and not differences due to non-equivalent amounts of protein within each gel lane or gels. Figures 6 shows a representative gel and Figure 7 represents its subsequent immunoblot.These figures illustrate that there were gross differences in the amounts of microsomal material needed in order to conduct these analyses. These differences depended on the tissue, treatment protocol and cytochrome species which was examined. For example, in studies of rats treated with the 500μg carcinogenic dose of 3-MC (Groups 3 – 4), 20μg of microsomal protein per gel lane was used for PP and LIV samples and 200μg for the THY samples. With respect to the samples generated using very high doses for either 3-MC (Group 5) or PB (Group 6), only 0.2 – 0.5μg of protein was typically needed to detect the CYP within liver samples, as seen in Fig. 2. Another example supporting the need to load different amounts of protein in seen in samples analyzed following PP treatment with the carcinogenic doses of 3-MC (Group 3), only 500μg of 3-MC elicited detectable levels of CYP1A2 in the PP; in this case pristane co-treatment (Group 4) affected the initial (i.e. 3hr post 3-MC injection) response, as seen in Fig. 4. With respect to the THY, only very high dose 3-MC treatment (Group 5) was effective in inducing CYP and in this case only CYP1A2 was detected; pristane co-treatment led to an augmentation of the response (Fig. 5). Thus, although there may have been differences in the apparent amounts of particular proteins which were present within different tissues or even within a particular tissue isolated after various treatments, the total amount of protein loaded onto each gel lane was equivalent.Figure 8 represents densitometric compiled values obtained from PP samples following treatment with 500μg 3-MC (panel A) and THY following very high 3-MC dose treatment (panel B) which were adjusted to reflect the amount of microsomal protein used to generate the immunoblot signals. As observed, the levels of CYP1A2 detectable in PP and THY were far below levels detectible in LIV. The data from quantitative analyses of liver samples was tabulated and the averages from four Western blot scans for each condition studied are presented in Table I. The table also summarizes the results of Western analyses of the CYPs examined in this study. This data represent the results of the scans which addressed the effects of basal (Group 1) versus the very high dose (Group 5) or carcinogenic dose (Group 3) effects of 3-MC on CYP in the LIV. In theory any change in intensity of the bands was indicative of a change in the amount of CYP protein expressed as a result of treatment with the indicated compound(s). Therefore, the desitometric scans validated the idea that observed differences were due to different tissues or various treatments, not loading inconsistencies.The complexity of chemical carcinogenesis has led to the notion that various regulatory systems respond to exposure to a broad range of endogenous and xenobiotic compounds which can cause cancer. Inherent to this notion is the existence of cellular control mechanisms which adjust to changing biological and biochemical conditions within the cell. Along these lines, the manifestation of cancer appears to be a reflection of modifications to normal cellular events, which involve regulation. One such cellular kinase regulated system involves the CYP monooxygenases, which respond directly to external xenobiotics. With respect to cancer, several lines of evidence support the existence of a relationship between changes in the expression of CYP and carcinogenesis. For example, certain enzymes within the CYP enzyme multigene family have been found to influence the overall process of tumorigenesis and therefore are of interest in studies involving chemical carcinogenesis.One of the most prevalent classes of chemical carcinogens in the environment which may act as initiators in chemical carcinogenesis is the polycyclic aromatic hydrocarbons (PAH). The PAH employed in this study was 3-MC. Although it has been implicated in many cancers, the parent compound is considered to be chemically inert. However, the 3-MC inducible monooxygenases catalyze the bioactivation of 3-MC eliciting mutagenic, carcinogenic or teratogenic derivatives. The studies presented above examined the possible relationship between basal and/or inducible CYP protein levels and exposure to different doses of 3-MC or pristane under conditions which elicited a preferential induction of different types of lymphoid malignancies.In the context of target organs specificity, within the PP co-treatment with pristane and low dose 3-MC resulted in only B-cell malignancies and no thymic tumors. On the other hand, with a high dose only thymic tumors developed, and no B cell malignancies were seen [9]. Subsequent studies revealed that within PP CYP1A enzymatic activity observed at low 3-MC dose was no longer detectable following high 3-MC dose treatment [7]. These studies reveal that CYP1A2 protein was detectable within PP only following the 500μg 3-MC dose treatment and pristane co-treatment was associated with a higher level of CYP1A2, but only at 3 hr after injection of 3-MC. These findings are reproducible and confirm the notion that B-cell tumors occur within PP as a result of CYPIA-catalyzed metabolic activation of low doses of 3-MC, but the abundance of CYP1A2 protein present after high 3-MC dose treatment coupled with the fact that no EROD activity was detectable at this dose treatment [7] infers that CYP1A2 protein was present but inactivated following 500μg dose 3-MC treatment and that pristane enhances this toxicity and/or carcinogenicity. A possible explanation is that reactive species of 3-MC which can bind to DNA or other macromolecules within the cell poisoned the PP, a gut associated lymphoid organ comprised of predominately B lymphocytes, due to extensive DNA adduct formation in the PP. However, it can not be ruled out that the decreased CYP1A enzymatic activity observed following high dose 3-MC treatment was due to interactions between CYP1A2 and another CYP isozyme.Comparatively, within THY, the target organ for T cell malignancies, Western blot analyses revealed that appreciable levels of CYP1A2 protein were expressed only after a very high 3-MC dose treatment. Although THY is not the initial site of injection, there is evidence which supports the fact of dissemination of 3-MC to the THY [9]. Also, the microenvironment within the thymus is predominantly composed of T-cells, as opposed to B-cells within PP, such that reactive species generated from high dose 3-MC treatment may not inactivate the thymus tissue. In light of the observed 3-MC dose effects, the oxidation of 3-MC to reactive species by CYP1A, dissemination of 3-MC to the THY, and in turn, a propensity towards thymic lymphomas following high dose 3-MC treatment, the data presented strongly support the hypothesis that very high 3-MC dose treatment affects the induction of thymic lymphomas and are attributable to the levels of CYP1A protein expression.Target organ specificity is a key determinant in both tumor type and location and pharmacokinetics parameters such as absorption, distribution, dissemination or elimination may be affected by different routes of exposure [16]. Along these lines, the PP was the injection site in our model. It is possible that expression of the 3-MC inducible CYP isozymes within PP rendered this tissue more prone to carcinogenesis, whereas the THY was more protected. Within LIV, where most of the CYP isozymes are present, the levels of CYP1A and CYP2B create a balance between toxification and detoxification, thus allowing effective CYP1A induced clearance of 3-MC from LIV and increased susceptibility of target organs. Most associations with chemical carcinogenesis in rats have been CYP1A1 in LIV, but only CYP1A2 was present in detectable levels in our target organs due to our route of exposure. This imbalance between CYP isozymes rendered the PP and THY more suceptible to the deleterious effects of 3-MC and pristane and helps to explain the preferential induction of lymphoid malignancies within these target organs.In light of the possible mode(s) of action of pristane, membrane perturbation by pristane may alter the phospholipid environment. Along these lines, it has been well established that drug-metabolizing enzymes located in the cell membranes can be affected by compounds which influence membrane structure and fluidity [23]. In other words, the functions of CYP are intimately related and highly dependent upon membrane interactions/function. Previous studies have clearly established that pristane is membrane associated [16] and elicits marked biological effects on membrane fluidity, as well as chromatin conformation [8]. Thus, the premise that pristine perturbs the plasma membrane and leads to altered expression, and/or function of the CYPs, is plausible. This may explain in part how pristane influences a diverse group of pathological processes, including carcinogenesis.In light of the results presented above and those in a previous study [7], a myriad of effects associated with pristane may have important implications in chemical carcinogenesis. Pristane clearly elicits marked effects on CYP protein expression and enzymatic activity. As such, it may affect cellular susceptibility to carcinogenic metabolites of 3-MC. More importantly, pristane appears to be a factor in the induction of experimental lymphoid malignancies and speculatively may be an etiological dietary factor in human cancer. As such, these data taken in conjunction with other observations from our laboratory, as well as from other investigators, demonstrate the biological relevance of studies to address the mechanism(s) of action of pristane.The effect of pristane on basal levels (C) of CYP in LIV. Immunoblots of LIV samples from unprimed (−) or pristane primed (+ = 1 ml pristane i.p. 2 wk before animal sacrifice; ++ = 1 ml pristane i.p. 4 wk before animal sacrifice) rats which did not receive 3-MC were used to establish background levels for untreated LIV. Hepatic microsomes were applied at 2μg, 20μg or 200μg of microsomal protein/well to determine the optimal loading concentration of LIV microsomal proteins necessary for detection of CYPs using Western blot procedures. Four duplicate gels were run, three were probed with the indicated, specific MAb probes (anti-CYP1A1, CYP1A2 and CYP2B) and one was stained with Coomassie blue R-250 to establish loading consistencies.Immunoblot of rat LIV microsomes from untreated (C), very high dose 3-MC treated (3-MC was dissolved in SSO and administered i.p. at 25mg/kg/day for 4 consecutive days) or PB (PB was dissolved in H2O and administered i.p. at 75mg/kg/day for 4 consecutive days treated animal). All microsomal samples were applied to the gel at a final concentration of 0.2μg of microsomal protein/well to access the effects of 3-MC (M) or PB (P) very high dose treatment on the expression of CYP in unprimed (−) or pristane primed (+ = 1 ml pristane i.p. 2 wk before animal sacrifice; ++ = 1 ml pristane i.p. 4 wk before animal sacrifice) LIV samples. This Western blot analyses served as a positive control for very high dose treated LIV. Three of the 4 duplicate gels were antibody probed (CYP1A1, CYP1A2, or CYP2B as indicated) and one was protein-stained as described in the Methods and Materials section.The effects of carcinogenic doses (0.05, 5 or 500μg) of 3-MC and pristane (− = unprimed; + = pristane primed) on CYP levels in the LIV. Animals were treated with indicated doses of 3-MC for 3 or 12hr exposure times. All microsomal samples were applied to the gel at a final concentration of 20μg of microsomal protein/well. The procedure for Western blot analyses is described in the Methods and Materials section. Three of the 4 duplicate gels were probed with the indicated antibody probe and one was stained with Coomassie blue. Equivalent concentrations of untreated control samples (C = control liver) were included on the gels as indicated.The temporal effects of 500μg of 3-MC injected into the PP in unprimed (−) or pristane primed (+) rats on the expression of CYPs in PP. Microsomal samples were prepared from rat tissues following treatment with 500μg 3-MC dose for the 3, 12 and 24 hr exposure time with or without pristane priming. Samples from PP were applied to the gel at a final concentration of 20μg of microsomal protein/well. The procedure for Western blot analyses is described in the Methods and Materials section. Three of the 4 duplicate gels were antibody probed and one was protein-stained. Equivalent concentrations of untreated control samples (CL = control liver; CPP = control PP) were included on the gels as positive or negative controls.The effects of 500μg 3-MC injected into the PP, 3-MC or PB very high dose treatment on expression CYP in THY. This immunoblot was generated from thymic microsomal samples from unprimed or pristane primed (− = unprimed; + = pristane primed) rats which were either untreated (CT), carcinogenic dose 3-MC treated (3, 12 or 24 hr), very high dose PB (P) or 3-MC (M) treated. Only thymic microsomal samples were applied to the gel at a final concentration of 200μg of microsomal protein/well. The procedure for Western blot analyses is described in the Methods and Materials section. Three of the 4 duplicate gels were antibody probed and one was Coomassie blue stained.Representative Coomassie blue stained SDS-Page gel of rat Peyer’s Patches (PP), THY (T = thymus) and LIV (L = liver) microsomal proteins isolated following treatment with 3-MC (500μg dose; 12hr exposure) and with (+ = 2 wk pristine priming) or (−) without pristane. Samples were loaded at 20μg of protein/well for PP and L, or 200μg of protein/well for THY samples. Control samples (C = no 3-MC) were run in lanes next to the corresponding Tx (treatment with 500μg 3-MC directly into the PP) samples for comparison purposes as shown on the immunoblot in Fig. 7. The mobilities of molecular weight markers (MW) are indicated. Gels were electrophoresed according to the Laemmli method.Composite immunoblot of PP, THY (T) and LIV (L) microsomal proteins from unprimed (−) or pristane primed (+) rats which were either untreated (C) or treated with 3-MC (Tx). All tissues were isolated from rats 12 hours after injection of 500μg of 3-MC into the PP. The optimized concentrations of microsomal proteins which were loaded into each of the wells of the gel are noted and the immunoblot generated was probed with MAbs specific for CYP1A1, CYP1A2 or CYP2B as indicated.Bar graphs generated from studies of the effects of 500μg 3-MC injected into the PP on CYP expression in PP and the effects of very high dose 3-MC treatment on expression of CYP in THY. Densitometric values of bands from immunoblots derived from rat PP (Panel A) and thymic (Panel B) microsomes were graphically represented using SIGMAPLOT. Values were determined by volume integration and represent the sum of all pixel values minus background. Microsomes were prepared from rats following indicated treatment protocols and immunoblots were analyzed as described in the text. The data depict densitometer units versus pristane treatment (−= unprimed; + = 2 wk pristane primed) or treatment with 3-MC (PP: 3hr, 12hr or 24hr; THY: very high dose, i.p.). Control LIV values are included on each graph as a point of reference.Composite of Densitometric Values of Immunoblots from Rat Liver Microsomes.Values were determined by volume integration and represent the sum of all pixel values minus background 1 pixel = 0.007744 mm2; P = Pristane; Time = 3-MC exposure time; ND = No detectible band; M = 3-MC; PB = phenobarbital; VH = Very high 3-MC dose; − = No pristane; + = 2wk pristane; ++ = 4wk pristane.This work was supported in part by the NIH-EARDA grant, Grant No. 1G11HDO46519-01, from the National Institute of Child Health and Human Development; the Excellence in Partnerships for Community Outreach grant (Grant # 632301-1210) and through the National Institutes of Health Grant GM 44982.
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+ Sunlight is a known human carcinogen. Many cosmetics contain retinoid-based compounds, such as retinyl palmitate (RP), either to protect the skin or to stimulate skin responses that will correct skin damaged by sunlight. However, little is known about the photodecomposition of some retinoids and the toxicity of these retinoids and their sunlight-induced photodecomposition products on skin. Thus, studies are required to test whether topical application of retinoids enhances the phototoxicity and photocarcinogenicity of sunlight and UV light. Mechanistic studies are needed to provide insight into the disposition of retinoids in vitro and on the skin, and to test thoroughly whether genotoxic damage by UV-induced radicals may participate in any toxicity of topically applied retinoids in the presence of UV light. This paper reports the update information and our experimental results on photostability, photoreactions, and phototoxicity of the natural retinoids including retinol (ROH), retinal, retinoid acid (RA), retinyl acetate, and RP (Figure 1).Sunlight is a complete carcinogen and is responsible for the induction of squamous cell and basal cell carcinomas in humans (1). Sunlight radiation is a continuum of wavelengths divided into five regions: (i) infrared (above 800 nm); (ii) visible (between 400 to 800 nm); (iii) UVA (between 315 to 400 nm); (iv) UVB (between 280 to 315 nm); and (v) UVC (between 200 to 280 nm). Humans are not exposed to UVC light due to stratospheric absorption of UVC by ozone.Vitamin A (all-trans-retinol; retinol) (ROH) is an important regulator in epidermal cell growth, normal cell differentiation, and cell maintenance (2). It also affects vision, reproduction, morphogenesis, and pattern formation. Metabolism is required to convert retinol to biologically active metabolites, e.g., all-trans-retinoic acid (RA, tretinoin) and its cis-isomers (3). Retinyl palmitate (RP) is the principal storage form of retinol in humans and animals and can be enzymatically hydrolyzed back to retinol in vivo. Both RA and its 9-cis isomer exhibit very high biological activities, including inducing epidermal growth and differentiation.Because the ester forms of retinol are thermally more stable than retinol, both PR and retinyl acetate have been commonly used in cosmetic products. As such, the number of cosmetic products containing RP has increased rapidly in the last two decades. Data available from FDA’s Voluntary Cosmetics Registration Program, compiled in accordance with Title 21 Section 720.4 of the Code of Federal Regulations (4), indicate that 102 cosmetic formulations in 1981, 355 cosmetic formulations in 1992, and 667 formulations in 2000 contained RP. Retail product categories containing RP include moisturizing preparations, skin care preparations, night skin care preparations, lipsticks, suntan gels and preparations, makeup preparations, and bath soaps and detergents (5).Although RP is thermally more stable than retinol (6), both compounds are still thermally unstable. As shown in Figure 2, the experimental results from our laboratories indicated that 2% RP in an oil-in-water emulsion (cream) decomposed gradually at 4°C, with about 38% decomposed on day 7. Also, RP, as are retinoids in general, is chemically unstable and its chemical stability is highly dependent on environmental conditions such as solvent, temperature, and availability of oxygen (7). RP is easily thermally-isomerized to the 13-cis isomer. Thermal isomerization is favored in lipophilic solvents and emulsions containing high compositions of oils (7). Anhydroretinol is one of the major decomposition products of RP (8).It has been shown that RP is much less stable under photoirradiation conditions than is retinol (6). The results shown in Figure 3 confirm that RP (1% in oil-in-water cream) photodecomposed faster than ROH under differing doses of UVA radiation. Boehnlein et al. (9) reported that after RP was topically applied in acetone to human skin, about 18% of RP penetrated the skin in 30 hrs. In addition, approximately 44% of the absorbed RP was hydrolyzed to retinol by skin esterases. As a result, this suggests that the RP and retinol present during a subsequent irradiation would be subject to photodecomposition. Indeed, we have demonstrated that photoirradiation of RP in ethanol under solar simulated light at an irradiance equivalent to terrestrial sunlight resulted in the formation of multiple photodecomposition products (Figure 4). Furthermore, sunlight-induced photodegradation of retinyl esters proceeds much faster than that of retinol, and it has been suggested that cellular retinol binding protein (CRBP) protects retinol from photodegradation (9, 10).Photochemical reactions of retinoids proceed through several different routes, including photoisomerization, photopolymerization, photooxidation, and photodegradation (11–13). The types of photodecomposition products that are formed are highly dependent on experimental conditions including vehicle, retinoid concentration, dosage and wavelength of the light, photoirradiation time, and the presence of other agents or impurities that can interfere with the photochemical reactions.Upon photoirradiation, retinoids including ROH, RA, RP, and retinyl acetate isomerize into a mixture of trans- and cis-isomers (14). The solvent can affect both the extent of isomerization of the trans-retinoids and the relative amounts of the different cis-isomers formed (15). All-trans-RA and 13-cis-RA are known to undergo Z-E isomerization and oxidation when exposed to light and air (16). Photodegradation of RA by fluorescent lamps resulted in five isomerization products (17).During photoirradiation, the primary retinoid photodecomposition products can also isomerize to secondary products. For example, as shown in Figure 4, anhydroretinol (AR) formed from photoirradiation of RP is initially in a trans-form. But during photoirradiation, it isomerizes into a mixture of all-trans-AR and three cis-ARs (Figure 5). The ratio of these isomers is photoirradiation time dependent.Reddy and Rao (18) studied the photoirradiation of retinyl acetate, and showed that isomerization of retinyl acetate and its photodecomposition product (AR) proceeded through an ionic photodissociation mechanism (Figure 6). Upon photoirradiation, retinyl acetate absorbs light energy and results in charge redistribution and change in bond order. Transition into a highly polarized singlet excited state was followed by either isomerization to the cis-retinyl acetate isomers, or release of the acetate anion to form a carbocation. Isomerization of this carbocation resulted in the formation of the cis-AR isomers.Photoreaction of retinoids produces oxidized products, such as ROH 5, 6-epoxide, 4-keto-ROH, and photodecomposition (or photodegradation) products. Photoirradiation of retinol in ethanol by UVC light (254 nm) resulted in the formation of retinal, ROH 5,6-epoxide, 5,8-epoxyretinol, and 13,14-epoxyretinol (19). The photooxidation of retinal formed 5,8-endoperoxide (19). Photooxidation of RA in 90% ethanol resulted in the formation of a number of oxidized products (16). The exposure of RP in methanol to UVA yielded palmitic acid, AR, and 4,5-dihydro-5-methoxyanhydroretinol (20,21). RP irradiated by UVC light (254 nm) produced AR, palmitic acid, and 2-butenyl palmitate (19). AR was also formed from photoirradiation of RP in ethanol under UVA light (22). Photooxidation of retinyl acetate in benzene with a trace of water produced the non-retinoids dihydroacetinidiolide, 2-hydroxy-2,6,6-trimethylcyclohexanone, β-ionone, geronic acid and desoxyxanthoxin (23).It is known that UVA light can photoactivate endogenous and exogenous photosensitizers (19). As expected, photooxidation of retinol and RP in the presence or absence of a photosensitizer resulted in the formation of different products (19). Since skin contains various endogenous photosensitizers and exogenous photosensitizers from topical application, photoreactions of retinoids in the skin are much more complicated than photoreactions performed in simple solutions.Besides the ionic photodissociation mechanism described above, photoirradiation of retinoids through free radical mechanism and generation of reactive oxygen species (ROS) has also been reported. In general, Type I photosensitization reactions occur when light is absorbed by a chromophore, and this molecule enters a photoexcited singlet state. It may then undergo intersystem cross-over and form a transient excited triplet state that can interact with other molecules and produce radicals via hydrogen transfer. In the presence of oxygen, superoxide anions may be formed, subsequently generating hydroperoxide radicals or hydrogen peroxide. In the absence of oxygen, radical anions may be formed or photoadditions can occur (13, 24, 25). Thus, irradiation (>300 nm) of retinal and retinol in methanol produces free radicals, as does irradiation of RP in dimethylformamide (13). Our study has confirmed that photoirradiation of retinol in 70% ethanol in water generated free radicals as detected by electron spin resonance (ESR) spectroscopy (Figure 7).In addition to radicals and ions, ROS are produced following irradiation of retinoids. For example, single oxygen is generated when retinal is illuminated in the presence of oxygen (26, 27). ROS can initiate lipid peroxidation and produce lipid alkoxy radicals and several tumorigenic small aldehydes (28, 29).Because they are exposed to the environment, the skin and eyes are more vulnerable to phototoxic damage than other organ systems. Retinoids are naturally abundant in both of these tissues and participate in specific phototoxic mechanisms. Photoirradiation of retinoids may generate acute and chronic toxicity through the formation of photoreaction products that are toxic, or photoexcitation of retinoids forming the excited retinoid species that exert toxicity directly or indirectly (Figure 8). Retinoids absorb light in the UVA range (315–400 nm) and thus would be photoexcited by light containing UVA. For instance, RP has a maximum UV-visible absorption at 326 nm (3) and thus, may be able to absorb UVA light and act as a photosensitizer. Thus, photoactivation of RP could generate short-lived ROS that have been shown to damage DNA and proteins and lead to tumors.The depth to which solar radiation penetrates tissues, and thereby induces adverse biological effects, depends strongly on its wavelength. In addition, the spectral transmission/absorption characteristics of solar-exposed tissues (Figure 9) influence penetration of light. These and other factors result in the generation of action spectra for photoinjuries to skin and eye components. The action spectra for each photohazard category (Figure 10) were applied to the most sensitive components of the eye or skin to arrive at threshold limit values (TLV) for ocular and cutaneous exposures to monochromatic and broadband light sources. The TLV were established in 1996 by the American Conference of Governmental Industrial Hygenists (ACGIH) (30).Negligible amounts of UVC are present in solar radiation at sea level and are generally incapable of penetrating the outer cornified layers of the epidermis or the cornea of the eye. However, protective shielding is required to protect the eyes and skin of workers from intense sources of UVC in the workplace, such as electric arc welding. Failure to prevent ocular UVB or UVC exposure leads to photokeratitis of the cornea (31), also known as “welder’s burn” and “snow blindness.” UVA and UVB penetrate the anterior chamber of the eye where absorption by the lens may induce cataract formation.The anterior layers of the eye are transparent to visible wavelengths but are absorbed by retina and the underlying choroid. Ocular exposure to intense blue light is believed to play a role in age-related macular degeneration and, possibly, ocular melanoma. Infrared radiation is strongly absorbed by the cornea and excessive exposure can cause corneal burns. Solar UVB penetrates the epidermis of nonpigmented skin to cause erythema and skin cancers. The superficial capillary bed of the dermis can be reached by UVA, visible, and infrared radiation and may cause photodynamic effects via absorption by blood-borne chromophores. Cutaneous exposure to visible or infrared wavelengths that penetrate the full depth of normal skin can produce thermal burns.Phototoxic autofluorescent lipofusin granules, comprised of complex indigestible lipid/protein aggregates, accumulate intracellularly as a consequence of aging. The accumulation of lipofusin in retinal pigment epithelial (RPE) cells is associated with increased photooxidative damage and is believed to contribute to macular degeneration and blindness. The major fluorophore detected in organic extracts of lipofuscin from RPE cells is an unusual pyridinium bisretinoid called A2E (Figure 11) generated as a byproduct from phosphatidylethanolamine and excess all-trans-retinal that evades recycling by RPE cells. Because it cannot be metabolized further, A2E accumulates in RPE lysosomes where it impairs clearance of phospholipids derived from phagocytosed rod outer segments and sensitizes RPE cells for apoptosis. Irradiation with blue light of the lipofuscin organic extracts of RPE cells generates ROS and produces lipid peroxidation. Irradiation of A2E also releases ROS, but the quantum yield is insufficient to account for the amount of ROS produced by lipofuscin. Ocular lipofuscin granules also contain all-trans-retinal which is 73-fold more efficient than A2E at photochemically generating singlet oxygen and 3.2-fold more efficient at generating superoxide radicals.Photoirradiation of RP by UVA light generates ROS, induces lipid peroxidation and causes DNA single strand cleavage in supercoiled ΦΧ174 plasmid DNA (32, 33). The phototoxicity of RP in human skin Jurkat T-cells also leads to cytotoxicity and DNA damage (33). Irradiating A2E with 21.6 J/cm2 430 nm light in PBS generated several derivatives containing 1–7 epoxide groups per A2E molecule (34). Cultured human RPE cells treated with A2E epoxides exhibited DNA damage detected by using the single cell gel electrophoresis/COMET assay.Certain naturally occurring retinoids influence gene expression directly by acting as agonists (e.g., RA) or antagonists (e.g. AR) for retinoic acid or retinoid-x receptors that regulate the transcription of genes containing functional retinoic acid response elements (RAREs). The expression of ROS-responsive genes may also be affected by retinoids acting either as antioxidants or as pro-oxidants via photochemical generation of ROS.UV radiation also influences gene expression by several mechanisms (see reference 35 for review). The release of latent growth factors (EGF, TGF-β) or cytokines (TNF-α) by UV exposure can stimulate their cognate receptors in an autocrine or paracrine manner. UV exposure can also trigger the release of lipid signaling molecules such as prostaglandin E2, platelet activating factor, ceramide, and lysophosphatidylcholine. Ligand-independent stimulation of receptors (EGFR, KGFR, Fas, TNFR, and TGFβR) and activation of downstream signal transduction pathways (ras, PKC, PLCγ, Smad2/3) can be caused by UV radiation. The classic cellular “UV response” involves transcriptional activation of immediate early genes c-fos and c-jun.and activation of the transcription factors NF-kB and AP-1. MAP kinase family members Mek-1/2, ERK-1/2, p38 MAPK, and c-Jun N-terminal kinases (JNKs) are activated by UV radiation and transduce signals to activate various transcription factors including Ets-1. ATF-3, a member of the CREB/ATF family, is strongly induced in cultured fibroblasts by sublethal doses of UVA (36). UVB-induced DNA damage is detected by the ataxia-telangiectasia mutated (ATM) and ataxia-telangiectasia and RAD3-related (ATR) systems. DNA damage signals are transduced via the checkpoint kinases CHK-1 and CHK-2 to p53, activating its ability to regulate the promoters of genes relevant to cell cycle arrest and apoptosis.The global effects of UV or retinoids on gene expression have been surveyed using DNA microarray technology (37–39). Coexposure to retinoids and UV provides the potential for synergistic or antagonistic effects on gene expression. In our laboratory, we compared the expression of the insulin-like growth factor-1 gene in the skin of Skh-1 hairless mice exposed to suberythemal doses of simulated solar light in the presence or absence of a topical cream containing 13% RP over 13 weeks (Figure 12). Exposure to simulated solar light or topical RP induced IGF-1 expression, although the IGF-1 promoter lacks a canonical RARE. The mechanisms by which RP and SSL influence IGF-1 expression are under investigation.There have been no studies on the effect of topically applied retinol, retinal, RP, and retinyl acetate on the carcinogenicity of UV light. However, retinyl acetate was reported as a co-carcinogen in carcinogenesis studies with butylated hydroxyanisole (BHA) using male F344 rats (41). The effects of topically applied RA on photocarcinogenicity in mice have been investigated by several research groups (42–47). The results are quite varied, with retinoid application increasing, decreasing, or having no effect on photocarcinogenesis. The drastically different effects on tumor incidence can probably be ascribed to differences in study design, particularly the use of different light sources and doses of light radiation. However, the current knowledge of the effects of RA on photocarcinogenesis does not allow a mechanistic explanation for the different outcomes.Regarding human toxicity, the long-term consequences of using cosmetics containing RP are currently unknown. It has been demonstrated that photoirradiation of RP can result in forming toxic photodecomposition products, generate ROS, induce lipid peroxidation, and cause DNA damage. Also, topically applied RP produces many of the cutaneous changes associated with the use of drug products containing RA which in some instances can enhance photocarcinogenesis. Thus, a study of the photocarcinogenesis of RP, under conditions relevant to the use of RP in cosmetics, is timely and important. As a consequence, RP has recently been nominated by the U.S. FDA and selected by the National Toxicology Program (NTP) as a high priority compound for phototoxicity and photocarcinogenicity studies. The goal of these studies is to provide relevant information necessary for risk assessment of RP in cosmetic creams.Names and numbering of retinoids.Decomposition of 2% RP in oil-in-water cream at 4°C.Photodecomposition of RP and ROH.HPLC profile of photodecomposition of RP under sun simulated light (10mJ.CIE/cm2).HPLC separation of cis- and trans-isomeric AR. HPLC analysis was conducted on a Vydac C18 column (4.6 × 250 mm) eluted isocratically with water in methanol (v/v; 14/86) at 1 mL/min.Ionic photodissociation mechanism of isomerization of retinyl acetate under UVA light irradiation to cis-retinyl acetate and a mixture of cis- and trans-AR.ESR measurements of 0.35 mg/ml ROH in 70% ethanol in water containing 200 mM spin trap α-(4-Pyridyl-1-oxide)-N-tert-butylnitrone (POBN) after being photoirradiated with UVA light at wavelength 320 nm. Conventional ESR spectra were obtained with a Varian E-109 X-band spectrometer. ESR signals were recorded with 15 mW incident microwave and 100 kHz field modulation of 1.25 G. All measurements were performed at room temperature.Potential photoreaction pathways of retinoids leading to phototoxicity and tumor formation.Penetration of skin and eye by solar radiation. (A) Diagram of skin. (B) Diagram of eye.Schematic action spectra for hazardous photoexposures (31).A2E, a fluorescent retinoid derivative isolated from retinal pigment epithelial cellsSimulated solar light and retinyl palmitate increase mRNA levels for insulin-like growth factor-1. Groups of HR/Skh-1 hairless mice (n=13) were exposed to simulated solar light (SSL) and/or 13% retinyl palmitate (RP) three times per week for 13 weeks. Whole skin samples were collected in RNAlater (Ambion) 48 hours after the final treatments. Total RNA was extracted using RNeasy Mini kits (Qiagen). IGF-1 and 18S rRNA levels were determined using qRT-PCR. The relative ratio of expression (RRE) for each group was calculated by the method of Pfaffl (40). Log transformation was used to normalize the data prior to Two Way ANOVA (p < 0.05).We thank Drs. Frederick A. Beland and John Bucher for critical review of this manuscript. This research was supported in part by an Interagency Agreement #2143-0001 between the Food and Drug Administration/National Center for Toxicological Research (FDA/NCTR) and the National Institute for Environmental Health Sciences/National Toxicology Program (NIEHS/NTP). Through this agreement, this research was supported by an appointment (S.C.) to the Postgraduate Research Program at the NCTR administered by the Oak Ridge Institute for Science and Education through an interagency agreement between the U.S. Department of Energy and the FDA.
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+ In Eastern cultures, such as India, it is traditionally recommended that women but not men cover their heads while working in the scorching sun. The purpose of this pilot study was to determine whether there was any scientific basis for this cultural tradition. We examined the differential cytotoxic effects of ultraviolet A light (UVA) on an established T cell line treated with female and male sex hormones. CD4+ Jurkat T cells were plated in 96 well plates at 2 × 106 cells/ml and treated with 17β-estradiol (EST) or testosterone (TE). These cells were irradiated by UVA light with an irradiance of 170 J/cm2 for 15min at a distance of 6 cm from the surface of the 96-well plate. Controls included cells not treated with hormones or UVA. The effects of EST and TE were investigated between 1 and 20 ng/mL. Cytotoxicity by fluorescein-diacetate staining and COMET assay generating single strand DNA cleavage, tail length and tail moment measurements were examined. The effect of estrogen (5ng/mL) on apoptosis and its mediators was further studied using DNA laddering and western blotting for bcl-2 and p53. We found that EST alone, without UVA, enhanced Jurkat T cell survival. However, EST exhibited a dose-related cytotoxicity in the presence of UVA; up to 28% at 20 ng/ml. TE did not alter UVA-induced cytotoxicity. Since TE did not alter cell viability in the presence of UVA further damaging studies were not performed. COMET assay demonstrated the harmful effects of EST in the presence of UVA while EST without UVA had no significant effect on the nuclear damage. Apoptosis was not present as indicated by the absence of DNA laddering on agarose gel electrophoresis at 5ng/ml EST or TE ± UVA. Western blot showed that estrogen down regulated bcl-2 independently of UVA radiation while p53 was down regulated in the presence of UVA treatment. EST and TE have differential effects on UVA-induced cytotoxicity in Jurkat T-lymphocyte which suggested that women may be more susceptible to the harmful effects of solar irradiation than men.In Eastern cultures women but not men are advised to cover their heads while working in the scorching sun. To understand the potential underlying immune mechanism we chose T cells as the target cell for investigation. We assume that the difference between male and female is primarily due to the difference in sex hormone levels and the solar impact on the skin is mainly due to ultraviolet light A (UVA) light [1]. Thus, we formulated the hypothesis that sex hormones under the UVA radiation alter T cell responsiveness. To accomplish our goals we studied the effect of estrogen and testosterone on established Jurkat CD4+ T lymphocytes. First, the cytotoxicity of testosterone and 17β-estradiol ± UVA was determined. Then, the genotoxicity of UVA radiation on 17β-estradiol treated T cells was investigated along with the regulation of p53/bcl-2 pathway.The tanning industry is rapidly growing in the United States. Currently, more than 1 million Americans use commercial tanning facilities every day [2]. The biggest categories of users are adolescents and young adults, especially women [1, 3]. Lamps currently used for recreational tanning emit UVA primarily or exclusively [2]. UV light is known to induce skin cancers by causing DNA gene mutations and inducing immunosuppression [4]. Thus, it is important to study the direct effects of UVA radiation on immune T cells.Sex hormones exert powerful effects in the susceptibility and progression of numerous human and experimental autoimmune diseases. This has been attributed to direct immunological effects of sex hormones that impact a clear gender dimorphism on the immune system. Globally, estrogens activate T-cell dependent humoral response thus contributing in the disease process while testosterone suppress T-cell immune responses and virtually always result in the suppression of disease expression [5]. Solar radiation has been proposed to play a detrimental role in the pathogenesis of cancer especially involving the skin and cervical cancer while a protective role in prostate cancer and colorectal cancer. Complex three way interactions between sex hormones, immune cells and UVA of the solar radiation appear to be involved in gender dimorphism of the immune system and immune system-related diseases such as cancer and autoimmunity.The aim of this study was to determine whether treatment with female and male sex steroids altered the cytotoxic effects of UVA on Jurkat CD4+ T lymphocytes and to determine related mechanism of action.All reagents were purchased from Gibco (Grand Island, NY) unless otherwise stated. Hormones were obtained from Sigma (St Louis, MO). Fetal bovine serum (FBS) was obtained from Hyclone Laboratories (Logan, UT). The human Jurkat T cells were purchased from the American Type Culture Collection (Rockville, MD). The cells were cultured in a humidified atmosphere with 5% CO2 at 37ºC. The standard growth medium was prepared using RPMI 1640, 10% FBS and 1% antibiotic (100U/mL penicillin, and 100 μg/mL streptomycin,) and 2 mM L-glutamine.The T cells were placed in a 96-well plate with 100 μl in each well. Two nine well sets were used in each 96-well plate for each hormone concentration. One set was covered with aluminum foil as a dark control, and the other was irradiated with UVA light using a type B 100 W UVA lamp from UVP Inc. (Upland, CA). The lamp emitted a UVA light band near 365 nm with an irradiance of 170 J/cm2 per/h at a distance of 6 cm from the surface of the 96-well plate. The cells were irradiated for 15 minutes. Among the nine wells of each replicate set, six well were used for cell viability assays and the remaining three for the Comet assay. For cell viability assay the fluorescein diacetate was added directly to the wells and allowed to incubate for 35 minutes and read using a Fluroskan II microplate reader (Lab Systems, Helsinki, Finland) with an excitation wavelength of 485 nm, and an emission wavelength of 538 nm.Cytotoxicity assay was carried out as previously described in our laboratory [6]. Briefly, cells were counted and resuspended at 20,000 cells/100 μl (100 μl/well) in media. Aliquots of 100 μl of cell suspension were placed in wells of microtiter plates, and 100 μl of different concentrations of 17β-estradiol or testosterone (resulting in a final concentration of 0 or 1, 2, 5, 10 and 20ng/mL) were added to the respective wells used to treat the cells. The plates were incubated for 30 minutes at 37ºC. After incubation, cells were centrifuged and washed twice with PBS. The PBS was removed and discarded and aliquots of 100 μL of fluorescein diacetate (10ng/mL) added. After 35 min incubation, the plates were read as described above.Cells were counted (10,000 cells/well) and re-suspended in media with 10% FBS. Aliquots of 100 μL of the cell suspension were placed in 96 well plates, treated with 100μl aliquot of either, media, 17β-estradiol or testosterone at 10ng/ml and incubated in a 5% CO2 at 37ºC for 72hrs. After incubation, the cells were centrifuged, washed with PBS, and re-suspended in 100 μL PBS. In a 2 mL tube, 20 μL of the cell suspension and 200 μL of melted agarose were mixed and 75 μL pipetted onto a pre-warmed slide. The slides were placed at 4ºC for 15 min and then placed in chilled lysis buffer for 45 min. Slides were washed twice for 5 min with Tris-Borate-EDTA (TBE) and electrophoresed in a horizontal gel apparatus at 25 V for 10 min. Slides were placed in 70% ethanol for 10 min, removed, tapped, and placed in an alkaline solution (99 mL H2O, 100 μL of 0.1 mM Na2EDTA and 1 M NaOH) for 45 min. Slides were air dried for 2.5 hrs, stained with SYBR Green and allowed to set for 4 hrs at room temperature. The slides were viewed with an Olympus fluorescence microscope and analyzed using LAI’s Comet Assay Analysis System software (Loates Associates, Inc. Westminster, MD).Apoptosis was determined by electrophoresis of nucleosomal fragments using a standard procedure for precipitating cytosolic nucleic acid [7]. Briefly, 1 × 106 cells were pelleted (1200 × g, 5min) and lysed for 15 min (250μl, 0.4% Triton-X, 20mM Tris, 0.4mM Na2EDTA) at 4ºC. Nuclei were then pelleted (13,000 × g, 5 min, 4ºC) and the supernatant was transferred to a clean microfuge tube. Nucleosomal fragments were precipitated overnight with an equal volume of isopropanol after adjusting to 0.5 M NaCl. The pellet which represents precipitated cytosolic DNA, was washed twice in 70% ethanol, dried briefly, and re-suspended in 40 μL TE (10 mM Tris-HCl, 1 mM Na2EDTA) with 1 mg/ml DNase-free RNase. Results were identical to those in which total DNA was prepared, but this modification facilitates resuspension of the DNA fragments which are transported from the nuclear location to the cytosol. A total of 15 μl was electrophoresed on a 1.8% agarose gel and stained with ethidium bromide for visualization. A representative experiment is shown in Figure 7. The picture of the gel in Figure 7 is the consequence of one of three experiments.The procedure utilized was as outlined in the manuscript by Jenkins et al. [7]. Briefly, cells were pelleted, washed with PBS. The PBS was discarded and cells were lysed in 100 μl of protein lysis buffer containing protease inhibitors. The lysed cellular protein solution was placed on ice. SDS-PAGE (12.5%) gels were loaded with sample and electorphoresed at 125 V. After electrophoresis, the gel was transferred to a nitocellulose membrane (Sigma, St. Louis, MO) overnight at 4ºC. After transfer, the membrane was placed in blocking solution (5% milk) on a shaker for 60 min at RT, washed with PBS-0.05% Tween (PBST). This was incubated with primary antibody (p53 or bcl-2 [mouse anti-human monoclonal antibodies]) diluted in 1% milk/PBST solution on a shaker 1 hr at RT, and washed 3 times for 5 min in PBST. The membrane was incubated with secondary antibody conjugate horse radish peroxidase (goat anti-mouse IgG monoclonal) for 60 min at RT on a shaker and washed 30 min in PBST. The chemiluminescent substrate was added for 1 min prior to the membrane being exposed to X-ray film.Cytotoxicity and comet (n=70) assay measurements were reported as means ± standard deviation (SD). Statistical analysis was performed by one-way ANOVA for multiple samples or by Student’s-t-testing with matched pairing if appropriate. For statistical analysis F-statistic ANOVA was applied to determine if there were significant differences in genotoxicity with regard to hormone and UVA exposure. Differences were considered significant at P values ≤0.05.The viabilities were tested in the concentration range 0 to 20ng/mL. Figure 1 and Figure 2 depicts the viability of Jurkat T cells with estrogen and testosterone ± UVA light. Figure 1 showed that estrogen alone without UVA enhanced T cell survival. Estrogen in the absence of UVA increases cell viability with 5ng/ml, 10ng/mL and 20ng/ml exhibiting 126%, 140% and 128%. However, in the presence of UVA light and estrogen, the survival decreases to 78% at 20ng/ml. The means of survival for estrogen treated T lymphocytes ± UVA light are statistically significant (2ng/ml and 5ng/ml, p<0.05 and 10ng/ml and 20ng/ml, p <0.01).Figure 2 shows that in the absence of UVA the cell viability varies from 92–97% while in the presence of UVA light there was a transient decrease in viability to 85% at 2 ng/mL but was steady at 5 to 20 ng/ml from 88 to 108%. The differences in means of survival with testosterone ± UVA did not reach statistical significance. Testosterone (20 ng/ml) + UVA partially improve cell viability.Figure 3 shows comet assay picture of cells ± UVA with estrogen at 5ng/mL. The nuclear DNA of untreated cells was round while estrogen + UVA were severely dispersed and fragmented. The percentages of DNA cleavage was approximately 8.5, 4.1 and 7% fragmentation (Figure 4), and the lengths of comet tail varied 37, 34 and 27% (Figure 5) while tail moment varied from 9.1, 2.7 and 5.8% (Figure 6) at 5, 10 and 20 ng/ml estrogen concentrations respectively. There are several ways to measure the severity of DNA fragmentation. In our study the variables studied were the percent of DNA fragmentation (percent of DNA in the Comet tail versus total DNA), tail moment and the length of the comet tail. Tail Length is the distance between the head and the last DNA fragment and Tail Moment is the product of % DNA and tail length mathematically written as (%DNA × Tail Length). The higher the percent of DNA fragments, the more severe is the damage. Similarly, the longer the comet tail, the smaller is the DNA fragment, the more severe is the damage.DNA laddering was determined as shown in Figure 7. This shows that UVA exposure of estrogen treated cells does not alter DNA laddering of T cells.Western blot (Figure 8) shows that bcl-2 was significantly and equally down regulated by estrogen and that ± UVA radiation had no effect. p53 is absent from the cells that were not treated with estrogen but was present when estrogen was added. UVA treatment decreased p53 expression.The sun emits a wide variety of electromagnetic radiation, including infrared, visible, ultraviolet A (UVA; 320 to 400 nm), ultraviolet B (UVB; 290 to 320 nm), and ultraviolet C (UVC; 10 to 290 nm) [3]. The only ultraviolet radiation wavelengths that reach the Earth’s surface are UVA and UVB. UVA rays pass deeper into the skin. However, its predominance in the solar energy reaching the Earth’s surface (tenfold to one hundredfold more than UVB) permits UVA to play a far more important role in contributing to the harmful effects of sun exposure than previously suspected. Thus, the protection from UVA has profound implications on public health worldwide.Studies have suggested that UV exposure can negatively affect the body’s immune system by interfering with the production of disease-fighting T cells and therefore lower the body’s natural defences against infection [3]. UV exposure modifies local and systemic immune responses by activating the T cell suppressor pathway. It has been suggested that prolonged exposure to sunlight may induce systemic or local immune alterations, which may facilitate the development of non-Hodgkin’s lymphoma [8]. The effects of prolonged sunlight exposure on peripheral blood cells showed an increase in cells expressing the interleukin-2 receptor (IL-2R) and, more specifically, an increase in the T cells expressing IL-2R and HLA-DR antigens [9]. These findings suggest that prolonged intense exposure to sunlight may be associated with immunostimulation, rather than immunosuppression [9]. Thus, it is important to study directly the effect of UVA radiation on immune T cells.The sun’s UVA rays are also linked with skin cancer and with the wrinkling that comes from sun exposure. UVA radiation primarily mediates singlet-oxygen damage triggering immediate pre-programmed cell apoptosis (T < 20 min) by immediately opening the cyclosporine A-sensitive (“S” site) mitochondrial megapore, while super oxide anions initiate another cyclosporine A-insensitive (“P” site) final apoptotic pathway [10]. This would imply that longer than 20 minute exposure of UVA radiation leads to total apoptosis which can be detected by DNA laddering.Sun exposure in some cases can be helpful. UVB rays are the sun’s ultraviolet rays linked with tanning and burning. Exposure to sunlight reduces the risk of prostate cancer [11]. Prognosis of breast, colon, ovary, and prostate as well as non-Hodgkin lymphoma may be related to synthesis of vitamin D(3) which is dependent on the degree of UVB exposure [12–15].The dilemma that sun exposure on one side is dangerous and on the other hand helpful can be explained by understanding the difference between UVA and UVB radiation. UVA radiation is harmful while UVB radiation which is linked with Vitamin D synthesis is beneficial [15, 16].Despite the harmful levels of solar UV radiation, mechanisms have evolved to protect cells and to repair damaged molecules. The cell component most vulnerable to injury is nuclear DNA. A number of different DNA repair mechanisms have been established [17]; the best known being photo reactivation, excision repair [18], post replication repair and SOS repair [17]. It seems that 15 minute UVA exposure of T cells treated with estrogen is not sufficient to give a permanent damage as there was no apoptosis in the DNA laddering assay.Effect of UVA light on Jurkat T cells in combination with azulene [19] and hydroxypyrene [20] has been previously studied. No study has yet appeared in the literature which studies the effect of UVA irradiation on estrogen- and testosterone-treated Jurkat T cells.A variety of evidence exists that hormones are thought to exert modulatory effects on immune function [7]. It is well documented that sex hormones testosterone and 17-β-estradiol play an important role in autoimmunity, pregnancy, menopause and prostate cancer [5, 7]. Direct hormone-specific effects on Jurkat CD4 + T lymphocytes have been studied previously by Jenkins et al. and McMurray et al [7, 21]. 17β-estradiol has been shown to inhibit Jurkat T cell proliferation, stimulate accumulation of cells in S and G2/M phases of the cell cycle, and induce apoptosis over 72 h in a dose-dependent manner [7]. Bcl-2 is a marker for uncontrolled cell growth [22, 23]. Bcl-2 protein and mRNA were also reduced in estrogen treated Jurkat T lymphocytes [7]. Additionally, bcl-2 protein levels were suppressed in association with estrogen-induced apoptosis. In our hands also bcl-2 protein was suppressed in the presence of estrogen and UVA radiation exerted no influence on its expression.Testosterone maintains lean body mass, bone density, skin elasticity and libido. It also modulates many physiological processes, including cytokines and immune T cells [24], in addition to being a precursor for the formation of estrogen [25]. Estrogen (17β-estradiol) modulates the course of both the menstrual cycle and menopause, so imbalances are directly linked to symptoms such as weight gain, headaches, premenstrual syndrome, mood swings, and abdominal cramps. Estrogen deficiency at menopause increases a woman’s risk of bone loss and osteoporosis. With the onset of menopause, however, decreased synthesis of 17β-estradiol is accompanied by an increased incidence of cardiovascular disorders and accelerated progression of renal diseases [26].Our data shows that testosterone >5 ng/ml improves lymphocyte viability while UVA exposure does not induce significant cytotoxicty. On the other hand, estrogen improves viability in a dose-response fashion but in the presence of UVA radiation it is cytotoxic. This implies that females are more prone to lymphocyte damage when exposed to sun specifically by UVA radiation as they have estrogen in the blood stream and lymphocytes in the skin.Several pathways have been described for apoptosis in different systems, but they all converge with activation of the protease caspase 3 and subsequently of endonuclease activity that results in fragmentation of nuclear DNA [27–32]. Experimentally, this DNA fragmentation has been demonstrated in model systems as a “ladder” following electrophoretic separation of DNA extracts prepared from the cell population. This method lacks sensitivity because the proportion of apoptotic cells in the population must be large to be detected. A method has been developed to detect DNA damage in single cells called the “comet” or single cell gel electrophoresis assay that can detect effects in subpopulations of cells [33]. Since UVA radiation in the presence of estrogen affected cell viability, we investigated the cytotoxic phenomenon by studying DNA damage using the comet assay. The comet assay revealed that DNA cleavage, tail moment and tail length were all increased in the presence of UVA radiation at a concentration >5 ng/ml of estrogen. Thus, our data shows that for a 15 minute exposure of UVA + estrogen, at the individual cell level, is damaging but not a sufficient number of cells are damaged at the population level to show DNA laddering.DNA damage caused by UV radiation, which if left un-repaired, results in molecular alterations in the skin and blood which eventually lead to skin mutations found in skin cancer. Skin biopsies, as well as blood samples when taken after a single UV exposure were tested for p-53. Specifically, the study analyzed the amount of p53 protein in the skin and blood. Within 24 hours of the first UV exposure, p53 protein was present in all layers of the epidermis [34]. The increased presence of p53 protein in the skin signifies that the body is responding to the cell damage due to UV exposure. p53 is a protein which allows cells to slow down their reproduction process so that damage from UV radiation can be repaired [35–36]. Though the body is trying to repair the damage, there is the risk of a mistake in the repair process, which increases as the number of altered cells multiply. If there is a ‘miss’ in the cell repair process, subsequent replication of the altered cell may yield a clone of abnormal cells which may eventually appear as a skin cancer. The p53 tumor suppressor plays a key role in protection from the effects of different physiological stresses (DNA damage, hypoxia, etc.), and loss of its activity has dire consequences, such as cancer [36, 37]. In our study we found that estrogen up-regulates the expression of p53, but that p53 is down- regulated in the presence of UVA. Thus, the absence of p53 exhibits a decrease in defence mechanisms of T cells in the presence of estrogen when exposed to 15 minutes of UVA radiation. The mechanism of action of UVA radiation is depicted in Figure 9.Most autoimmune diseases have been associated with the female gender, e.g. systemic lupus erythematosus (SLE). Female sex hormones, primarily estrogen and progesterone, have been implicated in this predisposition of the female gender to autoimmunity. Compared to the male sex steroid testosterone, these female hormones have been shown to have different effects on lymphocytes with respect to activation, proliferation, apoptosis and cytokine production [7, 21]. These differences may be responsible for the differences in the prevalence of autoimmunity and the immune hyper- responsiveness seen in women [7, 21, 38]. Interestingly, the autoimmune disease SLE commonly exhibits photosensitivity and skin rashes. Clinically uninvolved skin characteristically exhibits histologic evidence of immune reactivity. Photosensitivity is included as one of the eleven clinical diagnostic criteria for lupus (38). Furthermore, SLE patients may exhibit a flare of their systemic autoimmune disease after sun exposure (38). Enhanced DNA cleavage and reduced viability of lymphocytes with UVA, presumably an anti-immune effect, suggests estrogen might not play a role in cutaneous lupus, suggesting other female hormones, such as progesterone, may be the culprit. This is in agreement with recent reports suggesting that progesterone rather than estrogen are responsible for the immune hyper responsiveness in SLE [38]. Thus, the solar flares seen in females with SLE are not due to estrogen and therefore not related to the findings in this study.In 2004, more than one million cases of non-melanoma skin cancers are expected to be diagnosed in the United States [39]. An obvious question is whether men or women are at higher risk in view of the findings in this study. Unlike the case with cutaneous SLE, we are unable to interpret our findings with respect to the gender predisposition to skin cancer for the following reasons: basal cell and squamous cell carcinoma (since they are highly curable) are not traditionally included in overall cancer statistics, and no information is available regarding the proportions of women or men who intentionally tan and develop skin cancer. With respect to the issue of skin cancer in India, and the potential preventive effects of covering up in the scorching sun, it has not been possible to obtain appropriate statistics from India regarding the incidence of cancer in women and men.In conclusion, we have shown in our study that estrogen adversely affects T cells under the influence of UVA. Secondly, the UVA induced cytotoxicity caused damage at the individual cell level but the damage is probably repaired by the cellular processes as shown by the laddering assay. Western blot analysis showed that estrogen down regulated p-53 in the presence of UVA treatment. These studies suggest that the mere presence of estrogen may modulate the activity of T cells. It is possible that prolonged duration of UVA light may be very harmful to females and may be protective to males. This study provides a scientific basis for the cultural tradition in India in which females not males cover their heads with usually a cloth while in the scorching sun.Cell viability of Jurkat T cells at different concentrations of estrogen ± UVA.Cell viability of Jurkat T cells at different concentrations of testosterone ± UVA.Comparison of Comet assay of Jurkat T cells untreated as controls and cells treated with estrogen at 5 ng/ml ± UVA.Comet assay of Jurkat T cells treated with estrogen ± UVA where Y-axis is the percentage of DNA Cleavage and x-axis is the different concentrations of estrogen used in comparison with controls which had no estrogen.Comet assay of Jurkat T cells treated with estrogen ± UVA where Y-axis is the Tail Length (μm) and x-axis is the different concentrations of estrogen used in comparison with controls which had no estrogen.Comet assay of T cells treated with estrogen ± UVA where Y-axis is the Moment and x-axis is the different concentrations of estrogen used in comparison with controls which had no estrogen.DNA laddering of cells treated with estrogen ± UVA.Expression of bcl-2 and p53 by western blot of cells treated with estrogen in the presence and absence of ultraviolet AThis figure represents the mechanism of action of estrogen. Under the influence of UVA radiation estrogen causes increased T cytotoxicity which results in increased DNA damage and decreased p53.
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+ Exposure to lead has been well recognized in a number of work environments, but little is known about lead exposure associated with machining brass keys containing lead. The brass that is widely used for key manufacturing usually contains 1.5% – 2.5 % of lead. Six (6) licensed locksmiths and 6 case-matched controls successfully completed the pilot study to assess the prevalence of increased body lead burden of professional locksmiths. We measured both Blood Lead (atomic absorption spectrometry), bone-lead (KXRF) and had each subject complete a health and lead exposure risk questionnaire. One locksmith had not cut keys during the past two years, therefore this subject and case-matched control was excluded from the blood lead analysis only. The average blood-lead concentration (±SEM) for the 5 paired subjects was 3.1 (± 0.4) μg/dL and 2.2 (± 0.3) μg /dL for controls. Bone measurements, including all 6 paired subjects, showed tibia lead concentration (±SEM) for locksmiths and controls was 27.8 (± 2.3) μg /g and 13.7 (± 3.3) μg /g, respectively; average calcaneus lead concentration for locksmiths and controls was 31.9 (± 3.7) μg /g and 22.6 (± 4.1) μg /g, respectively: The t-test shows a significantly higher tibia lead (p<0.05) and blood lead (p<0.05) for locksmiths than for their matched controls, but no significant difference for calcaneus lead (p>0.10). Given that the mean tibia bone lead concentration was 13.1μg/g higher in locksmiths than in their matched controls, this average difference in the two groups would translate to an OR of increased hypertension in locksmiths of between 1.1 and 2.3, based on the published literature. Even with the very small number of subjects participating in this pilot study, we were able to demonstrate that locksmiths had significantly higher current exposure to lead (blood lead concentration) and significantly higher past exposure to lead (tibia lead concentration) than their age, sex and ethnically matched controls. Additional research is needed to fully identify the prevalence and associated risk factors for occupational exposure of lead in this previously understudied profession.Adverse health effects of lead exposure in the workplace are well-known for both workers and their families [1, 2]. More than 140 at-risk occupations for lead exposure have been identified [3] and the National Institute of Occupational Safety and Health (NIOSH) estimates that more than 3 million workers in the United States are potentially exposed to lead in the workplace [4]. Among these are workers of brass foundry and brass product manufacturing industries. A major route of exposure to lead and other toxic substances in these workers is inhalation of fumes and dust particles [5]. Workers exposed to lead fumes and dust particles may inadvertently contaminate their homes and expose family members with lead dust particles transported on their clothes, skin, hair, tools and in their vehicles [6, 7].Brass that is widely used for key manufacturing usually contains 1.5% – 2.5 % of lead [8]. Even the brass alloy manufactured with especially reduced lead content (so-called “reduced lead brass”) of American Society for Testing and Materials (ASTM) specification (B-121) has a nominal 2% lead. Another brass product, the so-called “free cutting brass”, (ASTM B-16/B-219) has a nominal 3.25% lead [9, 10].Concerns about this exposure prompted the Attorney General for the State of California to sue 13 key manufacturers and distributors for allegedly failing to warn that their products expose consumers to the toxic chemical lead in violation of Proposition 65 [11]. Proposition 65 – otherwise known as the California Safe Drinking Water and Toxic Enforcement Act of 1986 – was passed by the voters to protect the public from exposure to toxic substances known to cause cancer or be harmful to reproductive health. State conducted studies found the average exposure level for the lead residue on the hands of consumers handling brass keys was 19 times the Proposition 65 “No Significant Risk Level” of 0.5 μg per day limit [11].Although the lawsuit focused on protecting the end user of this product, the consumers, nothing was said about the locksmiths who actually spend more time than anybody else handling, grinding, and polishing lead-containing keys. Some keys are coated with a thin layer of metal such as nickel or chromium to minimize exposure to the underlying lead toxin, but this provides no protection for the locksmiths who must cut through the coating to the underlying metal. As a result of the grinding and polishing, lead-containing dust is dispersed throughout the workplace environment. The fine fraction of airborne dust may be inhaled directly, or during reintrainment after settling. Lead deposited on worksite surfaces may be ingested as a result of hand to mouth activity or contamination of open food or beverage containers. Lead deposited on work clothes may be translocated to the home environment, resulting in secondary, or “take-home” exposure to family members, particularly young children [6, 7, 12, 13].The facts on which this lawsuit was based provide critical insight into the potential occupational hazard for exposure to lead for locksmiths. Because they handle and grind keys on a daily basis, they sustain thousands of exposures during their entire active careers. The chronic low level exposure to lead spanning a 10 to 40 year career may produce significant cumulative total body burden of this well-known toxin. This is the first study to show a risk for lead exposure from yet another occupational source not previously examined.We recruited six locksmiths and six case-matched controls from volunteers residing in Los Angeles, California. Locksmiths selected for the study were at least 18 years of age, worked in his/her profession for at least one year and posses a valid California license as a locksmith or recent retiree. The six volunteer control subjects were matched to each locksmith by age, sex and ethnicity.Excluded from the study were unlicensed individuals who duplicate keys, individuals with known histories of exposure to lead from other sources (i.e. painters, battery workers, metal foundry workers, firearm instructors, subjects with retained bullets in their bodies, etc.), pregnant women in the first two trimesters, children less than 18 years of age and those unable to give informed consent. Subjects with paraplegia, quadriplegia, movement disorder or individuals without at least one natural leg free of metal plates, screws or other objects were excluded because such conditions would interfere with bone lead determinations. All data collection was completed in a single visit on or after the date the subject consented to participate.The Institutional Review Board of the KDMC approved the protocol and a written informed consent was given by each study participant.Blood lead was measured by atomic absorption spectrometry, using Zeeman background correction and graphite furnace. All blood lead measurements were performed by Westcliff Medical Laboratories, Inc. of Garden Grove, California. Westcliff fulfills the requirements for CLIA 1988 laboratory testing and participates in proficiency testing with the American Association of Bioanalysts and the College of American Pathologists. Abnormal test results are confirmed by repeating the test in duplicate, to ensure its reproducibility and accuracyIn vivo bone lead concentration (bone lead) was measured in study subjects by a K-shell X-ray fluorescence (KXRF) measurement system [14]. Bone lead measurements were made at the mediolateral and proximodistal midpoint of the anterior right tibia diaphysis and at the lateral surface of the right calcaneus. Because of measurement error, estimates of bone lead are sometimes negative, especially when true bone lead approaches zero. Bone lead slowly accumulates during the life of a person [15], with residence times for calcaneus and tibia bone lead estimated at 11 to 29 years (95 percent CI) and 16 to 98 years (95 percent CI), respectively [16]. Measurement of bone lead provides an estimate of cumulative past exposure of the subject. Estimates of bone lead were calculated from KXRF spectra using the recently corrected formulation [17, 18].During the bone lead measurements, a trained project staff member administered a structured questionnaire to capture demographic and medical information and to determine risk factors (occupational, avocational, life style, and habits) for lead exposure [19].The data were obtained from 6 professional locksmiths and 6 control volunteers matched to each locksmith by age, sex and ethnicity. The data are shown in the Table 1. Bone lead concentrations (tibia lead and calcaneus lead) are given in micrograms per gram in bone minerals, blood lead is given in micrograms per deciliter.The t-test for all 6 subject pairs showed significantly higher tibia (p<0.05, t=3.62, df(degrees of freedom)=10) and no difference for calcaneus (p>0.10, t=1.78, df=10). We excluded from the blood t-test subject number 4 and his case matched-control because the locksmith reported that he had not practiced his trade during within the past two years. The results of his blood lead did not reflect recent exposure to lead, a finding that corresponds with his work history. The t-test for the remaining 5 subject pairs showed significantly higher blood lead (p<0.05, t=3.14, df=8) for locksmiths than for their matched controls.Even with the very small number of subjects in this pilot study, we were able to demonstrate that locksmiths had significantly higher current exposure to lead (blood lead concentration) and significantly higher past exposure to lead (tibia lead concentration) than their age, sex and ethnically matched controls.Following the establishment of general industry lead standards in 1978, monitored data on permissible exposure limits have shown significant decreases in lead exposure for these facilities. Such decreases have not occurred in the construction industry since the enactment of construction industry standards in 1993 [4, 20]. Because permissible limits are based on monitored blood and air specimen standards that were established a quarter century ago (see Table 2), they have not taken advantage of current data that shows adverse health consequences also occurs at lower exposure limits [21–23].Until the State of California lawsuit, lead exposure from handling brass keys and in locksmiths cutting them has been essentially overlooked. This is not surprising since most of the exposure may be “low level”, resulting in blood lead concentrations that generally do not produce overt signs or symptoms during the active course of employment. There is only one case study in the English literature where the potential for lead exposure from grinding keys has even been evaluated. Ferguson reported that industrial hygiene monitored air samples were well within permissible Occupational Safety and Health Administration (OSHA) limits for three hardware store employees who cut an average of 5 keys per day [24].In cases where overt lead exposure does exist, it is often unrecognized or misdiagnosed due the nonspecific nature of lead toxicity, particularly in settings where lead exposure is not traditionally expected. The first report of serious lead poisoning occurring in the California plastic industry was uncovered when a worker requested his personal physician to perform a lead test on him after seeing “Lead” printed on bags of powder that he routinely handled at work. The worker with a blood lead level of 164μg/dl was suffering from abdominal pain, fatigue and constipation. Several of his co-workers also had elevated blood lead levels and two were found with blood leads 108 and 114μg/dl [25].Adding lead to brass, typically about 2% lead, makes it easier to cut the keys [11]. Although the dose appears low, inhaled lead particles that penetrate to the unciliated airways (100% for particles less than <1μm and up to 50 % of particles that are <50μm) are completely absorbed; particles cleared from the ciliated airways and those settling in the area above the trachea are typically ingested with 10% to 15 % being absorbed in adults and up to 50% in children [5, 26, 27]. Even though most of the lead exposure among locksmiths may be “subclinical”, recent research has nonetheless indicated that chronic low level lead exposure to adults poses significant risks for hypertension, adverse reproductive outcome, and possibly renal dysfunction [21–23, 28, 29].Bone lead as a biomarker suggests the impact of low level lead exposure may not occur until a certain cumulative exposure has taken place [30]. In the locksmith industry, where biomonitoring for past lead exposure by blood lead or industrial hygiene data is practically nonexistent, KXRF offers the optimal tool for assessing the magnitude (and clinical impact) of long term cumulative exposure. KXRF measurement of lead in bone identifies a tissue compartment that may serve as a potential source of increased lead exposure during pathological or physiological states associated with high bone lead turnover, such as osteoporosis, hyperthyroidism, bone fractures, metabolic bone disease, cancer, nutritional deficiency, etc.Our particular concern with this presumably moderate environmental exposure to lead in the work place is an endogenous factor that may contribute to substantial health risk in locksmiths during their lifespan. Constant exposure to lead, even in small amounts, leads to accumulation of lead in bones [31–35]. Lead can be actively mobilized from bone to blood during periods of normal homeostasis and periods of demineralization, such as long-term bed rest, osteoporosis, pregnancy and etc [36–38]. The ability of the skeleton to act as an endogenous source of lead to the blood has been shown in studies of the kinetics of lead after cessation of the occupational exposure [39–41]. Bone lead concentration is significantly related to health risk, especially in elevating blood pressure and increasing risk of hypertension. In these studies odds ratios (OR) for increased incidence of hypertension range from 1.1 to 1.9 for every 10μg/g increase in bone lead concentration, depending on the age, sex, and physiological status of the subjects studied [28, 29, 42, 43]. Applied to the present results, these studies suggest that the locksmiths in our study have odds ratios for increased risk of hypertension between 1.1 and 2.3 compared to the matched controls.Lead dust and particles are known to be a hazard to the family of workers who transported these particles in their clothing from the workplace to their homes. Children and women of childbearing age are especially at-risk from the adverse affects of lead exposure. Skeletal body stores from chronic lead exposure are mobilized during pregnancy at an accelerated rate and contribute significantly to blood lead levels during this period [44]. Since lead readily crosses the placenta and the developing blood brain barrier, lead exposure of the fetus and young child during pregnancy has potential implications for significant neurobehavioral disorders [44]. Exposure to high doses of lead during pregnancy has also been associated with an increased frequency of miscarriages and stillbirths among women working in the lead trades [45]. There is increasing evidence indicating that lead not only affects the viability of the fetus, but development as well. Developmental consequences of prenatal exposure to low levels of lead include reduced birth weight and premature births [46]. Lead is an animal teratogen, though most studies in human have failed to show a relationship between lead levels and congenital malformations [46]. Though regulatory limits on blood lead levels suggest there are levels of lead exposure without significant health risks, an increasing body of evidence indicates that there may be no lower limits of safe exposure to lead.Data from the Department of Consumer affairs for the State of California show over 1300 licensed locksmiths located in Los Angeles County alone. There are thousands key makers that are not licensed locksmiths but exposed to lead from this occupational source, such as workers in hardware stores, home improvement stores, discount outlets and other merchants providing key duplication services.This pilot study has identified an occupationally-related exposure to lead from yet another source that has not been previously examined. Additional research is needed to identify appropriate prevention measures and establish the relationship between chronic low level occupational lead exposure and body lead burden and corresponding health effects.Bone and Blood Lead Measurement for Locksmiths and Case Matched Controls Charles R. Drew University of Medicine and Sciences, Los Angeles, California, December 2003 to January 2004.Data from subject #4 and case-matched control were not included in blood lead statistical analysis – locksmith had not practiced his trade for over two years. All 6 subject pairs were included in bone lead data analysis.W = White; AA = African American; H = HispanicOSHA Lead Standards for Air and BloodSTAUDINGER, K. C.; ROTH, V. S.: Occupational. Lead Poisoning. American Family Physician, 1998. Adapted with permission from Occupational exposure to lead: final standard. U.S. Department of Labor, Occupational Safety and Health Administration. Federal Regist 1978; no. 29 CFR 1910.1025.Supported in part by the the Research Centers for Minority Institutions Grant No. RR03026 and the National Center for Research Resources Grant No. RR11145.C36000 Free Cutting Brass www.setonresourcecenter.com/dtSearch/dtisapi6.dll?cmd=getdoc&DocId=197898.
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+ This investigation involved the synthesis of metal complexes to test the hypothesis that structural changes and metal coordination in pyridine thiosemicarbazones affect cell growth and cell proliferation in vitro. Thiosemicarbazones are well known to possess antitumor, antiviral, antibacterial, antimalarial, and other activities. Extensive research has been carried out on aliphatic, aromatic, heterocyclic and other types of thiosemicarbazones and their metal complexes. Due to the pronounced reactivity exhibited by metal complexes of heterocyclic thiosemicarbazones, synthesis and structural characterization of di-2-pyridylketone 4N-phenyl thiosemicarbazone and diphenyl tin (Sn) and platinum (Pt) complexes were undertaken. Shewanella oneidensis MR-1, a metal ion-reducing bacterium, was used as a model organism to explore the biological activity under aerobic conditions. A comparision of the cytotoxic potential of selected ligand and metal-complex thiosemicarbazones on cell growth in wild type MR-1 and mutant DSP-010 Shewanella oneidensis strains at various concentrations (0, 5, 10, 15, 20 or 25 ppm) was performed. The wild type (MR-1) grown in the presence of increasing concentrations of Sn- thiosemicarbazone complexes was comparatively more sensitive (mean cell number = 4.8 × 108 ± 4.3 × 107 SD) than the DSP-010, a spontaneous rifampicillin derivative of the parent strain (mean cell number = 5.6 × 108 ± 6.4 × 107 SD) under comparable aerobic conditions (p=0.0004). No differences were observed in the sensitivity of the wild and mutant types when exposed to various concentrations of diphenyl Pt- thiosemicarbazone complex (p= 0.425) or the thiosemicarbazone ligand (p=0.313). Growth of MR-1 in the presence of diphenyl Sn- thiosemicarbazone was significantly different among treatment groups (p=0.012). MR-1 cell numbers were significantly higher at 5ppm than at 10 to 20ppm (p = 0.05). The mean number of DSP-010 variant strain cells also differed among diphenyl Sn- thiosemicarbazone complex treated groups (p=0.051). In general, there was an increasing trend in the number of cells from about 5.0 × 108 cells (methanol control group) to about 6.0 × 108 cells (25ppm). The number of cells in methanol control group was significantly lower than cell numbers at 20ppm and 25ppm (p = 0.05), and numbers at 5ppm treatment were lower than at 20 and 25ppm (p = 0.05). Furthermore, a marginally significant difference in the number of MR-1 cells was observed among diphenyl Pt- thiosemicarbazone complex treatment groups (p = 0.077), and an increasing trend in the number of cells was noted from ~5.0 × 108 cells (methanol control group) to ~5.8 × 108 cells (20ppm). In contrast, the DSP-010 variant strain showed no significant differences in cell numbers when treated with various concentrations of diphenyl Pt- thiosemicarbazone complex (p = 0.251). Differences in response to Sn- metal complex between MR-1 and DSP-010 growing cultures indicate that biological activity to thiosemicarbazone metal complexes may be strain specific.This investigation involves the synthesis of metal complexes to test the hypothesis that structural changes and metal coordination in pyridine thiosemicarbazones affects growth and cell proliferation in vitro. Shewanella oneidensis MR-1[1], a metal ion-reducing bacterium, and DSP-010, a spontaneous rifampicillin derivative of the parent strain, were used as model organisms to explore the biological activity of newly synthesized metal complexes of α-N-heterocyclic carboxaldehyde thiosemicarbazones (HCT) in vitro. Studies of metal complexes of thiosemicarbazones have shown that they can be more active in cell destruction, as well as in the inhibition of DNA synthesis, than the uncomplexed ones. Thiosemicarbazones are a class of compounds obtained by condensing thiosemicarbazide with suitable aldehydes or ketones and are well known to possess antitumor [2], antiviral [3], antibacterial [4], antimalarial [5], and other activities. They belong to a large group of thiourea derivatives, the biological activity of which are a function of parent aldehyde or ketone.Extensive research has been carried out on aliphatic, aromatic, heterocyclic and other types of thiosemicarbazones and their metal complexes [6–13]. Efforts to evaluate structural features essential for biological activity have included:Exchange of the sulfur atom of thiocarbonyl group by oxygen or selenium.Changing the point of attachment of the thiosemicarbazone moiety in the parent aldehyde or ketone.Substitution on the terminal 4N position.Variation of the parent aldehyde or ketone.The thiosemicarbazone moiety (Figure 1) acts as a chelating agent for metal ions by binding through the sulfur atom and the hydrazine nitrogen atom. Thiosemicarbazone molecule can exist in tautomeric form in solution as thione or thiol (Figure 2). The thione form can act as a neutral bidentate ligand, while the thiol form can be a singly charged bidentate ligand due to loss of its proton. In metal complexes, thiosemicarbazones can act as a tridentate ligand with a donor atom apart from thione/thiol sulfur atom and azomethine nitrogen. The possible ligation in metal complexes can be either as a neutral molecule or as a monobasic anion due to loss of hydrogen from azomethine nitrogen atom.The molecular features essential for such activities can be ascertained by designing synthetic routes to modify, replace, or substitute the derived thiosemicarbazone ligand. In this direction, the phenyl and ethyl derivatives of cyclopentanone thiosemicarbazones were synthesized and characterized [14]. Recently, attention was focused on the derivatives of di-2-pyridylketone thiosemicarbazones as a ligand for tin [15] and other metals. As only a few tin (IV) complexes of these ligands have been reported, the synthesis and chelating behavior of dibutyl and diphenyl tin (IV) along with platinum (IV) complexes of di-2-pyridylketone phenyl thiosemicarbazones and its reactivity in Shewanella oneidensis. It is anticipated that data obtained from this study will provide information on the interactions of metal complexes in vivo and the development of better therapeutic agents for translational research.A solution of di-2-pyridyl ketone (1.84g, 10 mmol) in 50mL ethanol was slowly added with stirring to a solution containing 4-phenyl-3-thiosemicarbazide (1.67g, 10mmol) in ethanol (50 ml) and five drops of concentrated hydrochloric acid. The reaction mixture was refluxed for 2 hours. A yellow solid resulted on cooling. The crude sample was filtered and re-crystallized in ethanol (yield ca.70%) with a melting point 138–140°C. Melting points were determined using mel-temp II apparatus. IR spectra were obtained in 4000–400 cm−1 range in KBr pellets on a Nicole 670 FT-IR spectrophotometer [νN-H 3150 cm−1 (br); νC=N 1588 cm−1(sh), 1531 cm−1(sh), 1480 cm−1(sh),1455 cm−1(sh); νC=S 999 cm−1 (sh), 917 cm−1 (sh)]. The absorption spectra recorded on Cary 3E UV-Vis spectrophotometer using 1 cm quart cuvettes in DMSO as solvent. The visible spectra of the ligand showed strong absorbance at 345 nm, and 280 nm indicating π-π* transition of pyridyl ring and imine function of thiosemicarbazone moiety, and their n-π* transitions.A typical method of preparation involved either refluxing or stirring the metal salts dissolved in anhydrous ethanol in 1:1mole ratio.The tin complex of di-2-pyridylketone 4N-phenyl thiosemicarbazone was prepared by the following procedure: to a hot ethanolic solution (50 mL) of the ligand (5 mmol, 1.67g) was added a solution in ethanol (50 mL) of diphenyltin dichloride (5 mmol, 1.72g). The mixture was refluxed for about 30 minutes and the resulting solution cooled to obtain a yellow solid (yield ca. 65%) with a melting point 138°C–140°C. A comparison of IR spectrum with the ligand showed a shift of the band from ~3300 cm−1 to 3450 cm−1(br) is attributable to νN-H. This indicates that the ligand is coordinated in its deprotonated form. The shift of imine νC=N band from ~1530 cm−1 to ~1550 cm−1 indictes the coordination of azomethine nitrogen N(2) to the metal ion. The band νC=S at about ~850 cm−1 of the ligand is shifted to lower frequencies by ~40–50 cm−1 suggests the coordination of metal through sulfur atom. The bands at about ~510 cm−1 can be assigned to the metal nitrogen binding. The absorption spectrum of the tin complex was similar to that of the ligand. Conductivity measurements of the metal complexes were determined in DMSO using YSI model 35-conductance meter. The low molar conductance values (< 10Smol−1cm2) of the tin and Pt-complexes indicate the non-electrolytic nature.To a suspension of platinum (II) dichloride (g) in methanol (50 mL) was added 5.0 mmol methanolic solution (50 mL) of the ligand (1.67g), and the mixture was refluxed for about 1 hour. The resulting mixture was left to air at room temperature to obtain a dark red colored solid (yield 80%) with a melting point 102°C–103°C. The IR spectrum has shown a shift of 10–30cm−1 for ν(C=N) absorption band indicating the positive involvement of the azomethine nitrogen in bonding to the metal ion. Similarly the ν(C=S) band shifted by 5–10cm−1 indicating the coordination of thiocarbonyl sulfur to the metal ion. The molar conductivity of the complex in DMSO indicated that the platinum complex (~50Smol−1cm2) is nonpolar in nature.Wild-type Shewanella oneidensis MR-1 and DSP-010, a spontaneous rifampicillin derivative of the parent strain (a kind gift from the Environmental Science Group at Oak Ridge National Laboratory) were grown under aerobic condition in Luria-Bertani (LB) medium at 37ºC. This species of Shewanella is gram-negative, facultative anaerobes. Bacterial cell concentrations used in the experiments were equivalent to an optical density of 0.4 at 600 nm (approximately 8 × 108 cells/mL = 0.18 g/L). The optical densities of the cell suspensions were measured spectrophotometrically at 600 nm using a Bio-RAD Smart Spec instrument. To an exponentially growing culture of MR-1 and DSP-010 cells, uncomplexed ligand, Pt- and Sn- complexes were added at 0, 5, 10, 15, and 25 ppm (μg/ml). Negative controls containing media alone or solvent controls (methanol) were included in each growth curve experiments. Cells were incubated for 48 h at 37ºC and the bacterial growth was measured at 600nm. Mean cells per milliliter and optical densities at 600nm [OD600nm] were plotted against Ligand, Pt- and Sn-metal concentrations for Shewanella oneidensis MR-1 and DSP-010 cells.A one-way analysis of variance (ANOVA) test was used to determine whether there are significant differences in mean numbers of S. oneidensis cells exposed to the test chemicals. The Fisher’s PLSD test was used for pair-wise comparisons among treatment groups. The Student’s t-test was applied for comparing paired data sets (significance at a p-value ≤ 0.05).In this study, wild-type Shewanella oneidensis demonstrated biological activity when exposed to newly synthesize Pt-, and Sn- complexes of α-N-heterocyclic carboxaldehyde thiosemicarbazones. Wild type (MR-1) grown in the presence of increasing concentrations of Sn- thiosemicarbazone complexes was comparatively more sensitive (mean cell number = 4.8 × 108 ± 4.3 × 107 SD) than the DSP-010, a spontaneous rifampicillin derivative of the parent strain (mean cell number = 5.6 × 108 ± 6.4 × 107 SD) under comparable aerobic conditions (Unpaired t-test, t-value=−4.003, df =28; p=0.0004; n=15). No differences were observed in the sensitivity of the wild and mutant types when exposed to various concentrations of diphenyl Pt- thiosemicarbazone complex (Unpaired t-test, t-value = −0.81; p= 0.425) or the thiosemicarbazone ligand (Unpaired t-test, t-value=−1.03, p=0.313).Growth of MR-1 in the presence of diphenyl Sn- thiosemicarbazone was significantly different (ANOVA, df=6,14; F-value=4.259; p=0.012) among treatment groups (Figure 3). MR-1 cell numbers were significantly higher at 5ppm than at 10 to 20ppm (Fisher’s PLSD test, p = 0.05). The mean number of DSP-010 variant strain cells also differed among diphenyl Sn- thiosemicarbazone complex treated groups (ANOVA, df=6,13; F-value = 2.897; p = 0.051) (Figure 3). No differences were observed in the number of cells in the negative control and methanol control groups (Fisher’s PLSD test, p > 0.05). In general, there was an increasing trend in the number of cells from about 5.0 × 108 cells (methanol control group) to about 6.0 × 108 cells (25ppm). The number of cells in methanol control group was significantly lower than cell numbers at 20ppm and 25ppm (Fisher’s PLSD test, p = 0.05), and numbers at 5ppm treatment were lower than at 20 and 25ppm (Fisher’s PLSD test, p = 0.05).Furthermore, a marginally significant difference in the number of MR-1 cells was observed among diphenyl Pt- thiosemicarbazone complex treatment groups (ANOVA, df=6,14; F-value=2.464; p = 0.077), and an increasing trend in the number of cells was noted from ~5.0 × 108 cells (methanol control group) to ~5.8 × 108 cells (20ppm) (Figure 4). In contrast, the DSP-010 variant strain showed no significant differences in cell numbers (Figure 4) when treated with various concentrations of diphenyl Pt- thiosemicarbazone complex (ANOVA, df=6,14; F-value=1.492; p = 0.251).MR-1 cells showed no significant differences in numbers (Figure 5) when treated with various concentrations of the uncomplexed thiosemicarbazone ligand alone (ANOVA, df=6,14; F-value=0.235; p = 0.235). A similar observation was noted when DSP-010 variant strain was treated with various concentrations of the ligand (ANOVA, df=6,14; F-value=2.004; p-value = 0.133) (Figure 5).Micro-organisms require the presence of a number of metals that play essential biochemical roles such as catalysts, enzyme co-factors, activity in redox processes and stabilizing protein structures [16]. Metals may accumulate above normal physiological concentrations by the action of unspecific, constitutively expressed transport systems, whereby they become toxic. Intracellular metals can exert a toxic effect by forming coordinate bonds with anions blocking functional groups of enzymes, inhibiting transport systems, displacing essential metals from their native binding sites and disrupting cellular membrane integrity [17]. There are five basic mechanisms that convey an increased level of cellular resistance to metals: (1) efflux of the toxic metal out of the cell; (2) enzymic conversion; (3) intra- or extracellular sequestration; (4) exclusion by a permeability barrier; and (5) reduction in sensitivity of cellular targets. Our study indicates that S. oneidensis MR-1 is susceptible to growth inhibition by tin complex. In contrast, DSP-010, a spontaneous rifampicillin derivative of the MR-1 parent strain, demonstrates marked growth in the presences of di-2-pyridylketone 4N phenyl thiosemicarbazone as a ligand and tin or platinum complexes. For diphenyl Sn- thiosemicarbazone, untreated and methanol solvent controls groups were not significantly different (Fisher’s PLSD test; p=0.114) therefore, the observed differences of DSP-010 cell growth among treatments as compared to MR-1 are most likely due to the effect of Sn on Shewanella mutant bacteria.Growth comparison of Shewanella oneidensis MR-1 and DSP-010 strains in the presence of thiosemicarbazone ligand, diphenyl Pt- and Sn- thiosemicarbazone complexes demonstrated strain specific differences under aerobic conditions. No significant change occurred with MR-1 or DSP-010 strains grown in the presence of the uncomplexed ligand. Similarly, wild type and mutant strains of S. oneidensis did not demonstrate significant growth differences in the presences of diphenyl Pt- thiosemicarbazone. However, an unpaired t-test indicated significant difference between MR-1 and DSP-010 after 48-hour growth in the presence of increasing concentrations of diphenyl Sn- thiosemicarbazone. Differences in response to Sn- metal complex, between MR-1 and DSP-010 growing cultures indicate that biological activity to thiosemicarbazone metal complexes may be strain specific.The possible reduction of inorganic Sn(IV) to Sn(II) may explain the growth advantage of DSP-010 mutant strain over wild-type MR-1 with di-2-pyridylketone thiosemicarbazone derivatives. In literature, microbial resistance mechanisms have been widely studied [18,19]. However, very little is known about the reduction of inorganic Sn(IV) to Sn(II) in microbes. It may be hypothesized that heavy metal complexes such as diphenyl tin (IV) thiosemicarbazone are unable to escape sequestration, metal efflux or metal reduction systems thereby causing toxicity in Shewanella oneidensis MR-1. In contrast, DSP-010 may be able to effectively reduce inorganic Sn(IV) to Sn(II) thus allowing growth in increasing concentrations of Sn- metal complexes. Alternately, DSP-010, may be able to mobilize and remove di-2-pyridylketone 4N-phenyl thiosemicarbazone, diphenyl tin and platinum complexes more efficiently through membrane transport systems. Future binding studies should address the uptake and binding affinities of diphenyl thiosemicarbazones in Shewanella oneidensis MR-1 and DSP-010.Thiosemicarbazone uncomplexed ligand moiety.Thiosemicarbazone molecule as thione or tautomeric forms in solution.Effects of various concentrations of diphenyl Sn- thiosemicarbazone metal complex on Shewanella oneidensis MR-1 and DSP-010 grown aerobically in LB broth. Mean cell/ml was determined at 37ºC after a 48 hour time period for three independent experiments. Error bars are for standard deviation.Effects of various concentrations of diphenyl Pt- thiosemicarbazone metal complex on Shewanella oneidensis MR-1 and DSP-010 grown aerobically in LB broth. Mean cell/ml was determined at 37ºC after a 48 hour time period for three independent experiments. Error bars are for standard deviation.Effects of various concentrations of uncomplexed thiosemicarbazone ligand on Shewanella oneidensis MR-1 and DSP-010 grown aerobically in LB broth. Mean cell/ml was determined at 37ºC after a 48 hour time period for three independent experiments. Error bars are for standard deviation.This work was financially supported in part by a grant from “ONR Interns in Biomolecular Sciences” grant #00014-03-1-0317, National Institutes of Health-Research Centers for Minority Institutions (NIH-RCMI) Core Facilities grant #1G12RR12459-01, and Jackson State University Center for Scholar’s Award Program. Shewanella oneidensis MR-1 and DSP-010 strains were a kind gift from The Environmental Science Division-Oak Ridge National Laboratory (ESD-ORNL), Oak Ridge, Tennessee.
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+ In an effort towards adapting new and defensible methods for assessing and managing the risk posed by microbial pollution, we evaluated the utility of oligonucleotide microarrays for bacterial source tracking (BST) of environmental Enterococcus sp. isolates derived from various host sources. Current bacterial source tracking approaches rely on various phenotypic and genotypic methods to identify sources of bacterial contamination resulting from point or non-point pollution. For this study Enterococcus sp. isolates originating from deer, bovine, gull, and human sources were examined using microarrays. Isolates were subjected to Box PCR amplification and the resulting amplification products labeled with Cy5. Fluorescent-labeled templates were hybridized to in-house constructed nonamer oligonucleotide microarrays consisting of 198 probes. Microarray hybridization profiles were obtained using the ArrayPro image analysis software. Principal Components Analysis (PCA) and Hierarchical Cluster Analysis (HCA) were compared for their ability to visually cluster microarray hybridization profiles based on the environmental source from which the Enterococcus sp. isolates originated. The PCA was visually superior at separating origin-specific clusters, even for as few as 3 factors. A Soft Independent Modeling (SIM) classification confirmed the PCA, resulting in zero misclassifications using 5 factors for each class. The implication of these results for the application of random oligonucleotide microarrays for BST is that, given the reproducibility issues, factor-based variable selection such as in PCA and SIM greatly outperforms dendrogram-based similarity measures such as in HCA and K-Nearest Neighbor KNN.As the number of beach closings and advisories continue to rise, so does the public’s concern regarding microbial pollution in recreational waters. In a survey of more than 230 U.S. coastal and Great Lake communities, there were at least a total of 13,410 days of beach closings or advisories during 2001 [1]. The majority of beach closings and advisories were based on the presence of elevated levels of fecal contamination as measured by fecal bacterial indicators, such as Escherichia coli and Enterococci. Under section 303(d) of the 1972 Clean Water Act, states, territories, and authorized tribes are required to develop pollutant-specific lists of impaired waters and may be required to establish a total maximum daily load (TMDL) for those impaired waters [2]. TMDLs specify the maximum amount of a pollutant that a water body can receive and still meet water quality standards. Fecal coliforms are frequently listed as impairment on many states 303(d) list of associated water-quality impairments [3]. While TMDLs have historically focused on chemical impairments, more attention is now being focused on microbial impairments. Recently, the EPA published an extensive protocol for developing pathogen TDMLs [2]. Currently, there are several regional pilot projects underway aimed at establishing fecal coliform TMDLs for impacted watersheds [4].Reducing the loads of fecal contamination can be problematic because often the pollution sources are not known or have non-point sources. Non-point sources of microbial fecal pollution are mobilized by rain/snow events and can include urban litter, agricultural runoff, failing sewer lines, malfunctioning septic systems, and domestic and wildlife excrement. Implementation of best management practices (BMPs) for TMDL compliance is dependent upon accurately identifying the source(s) of the impairment. Source tracking of non-point sources of microbial pollution, specifically indicator bacteria, has been generically referred to as bacterial source tracking (BST) [5] or microbial source tracking (MST) [6,7] and can be accomplished using a collection of multidisciplinary bacterial sub-typing methods. In addition to determining the origin of fecal contamination, BST methods can differentiate between human and non-human sources of microbial pollution [6,7], which can aid in generating more accurate risk assessments for managing the risk posed by microbial pollution.BST methods can be divided into two general groups, 1) phenotypic or biochemical-based methods, and 2) genotypic or molecular-based methods [7]. Of the phenotypic methods, multiple antibiotic resistance (MAR) analysis has been reported the most and has been shown to be successful in 1) discriminating human and animal sources of E. coli or fecal streptococci [8, 9, 10] and, 2) further discriminating animal sources by animal type [11]. This method involves isolating and culturing target indicator organisms from various sources and locations to create a reference library. These isolates are subsequently replica plated on selective media containing multiple antibiotics at a range of concentrations. Antibiotic susceptibilities are characterized, subjected to discriminant analysis and compared to a reference antibiotic susceptibility library to determine identity. Reliability of the method is determined by analyzing isolates as both standards and as unknowns. The number of isolates assigned to the correct categories divided by the total number of isolates is referred to as the average rate of correct classification (ARCC) [12]. ARCC values for this method range from 62% to 94% when individuals are compared. Despite the success of this method in simple watersheds [11], some researchers have indicated that MAR lacks the sensitivity, reproducibility, and host specificity that is needed for BST [13].In contrast to the limited number of phenotypic sub-typing methods, numerous genotypic methods have been described including ribotyping [14, 15, 16], length heterogeneity polymerase chain reaction (LH-PCR), terminal restriction fragment length polymorphism (T-RFLP) PCR [17, 18], repetitive PCR (rep-PCR) [19], denaturing gradient gel electrophoresis (DGGE) [20, 21], pulsed-field gel electrophoresis (PFGE) [22, 23, 24], and amplified fragment length polymorphism (AFLP) [25]. Most of these molecular methods rely on PCR to interrogate a fraction of the target organisms’ available genetic information. PCR amplification products are subsequently resolved by gel-electrophoresis and the resulting banding pattern may be compared to a reference library to determine the identity of the organism. ARCC values can approach 100% when using some of these methods, such as rep-PCR [19]. Despite the success of genotypic methods, there is an ongoing need in BST for increased resolving power to discriminate between closely related microorganisms. Newer technologies, like DNA microarrays, which have been employed for various environmental microbiology applications [26], could potentially increase the resolving power of BST analysis [27]. For example, DNA microarrays interrogate DNA samples at the DNA sequence level. In contrast, gel-based methods rely on DNA fragment sizing; a method in which co-migration of heterogeneous DNA sequence populations of similar sized fragments is possible. Unlike gel-based methods, which rely on size fractionation of banding patterns that are subject to positional variation, DNA microarray profiles are comprised of physically immobilized, addressable spots. In addition, the resolving power of the microarray can be further improved by increasing the amount of oligonucleotide elements on the micro array.The methods and data analysis algorithms for the application of DNA microarrays towards BST are just starting to be developed. Recently, oligonucleotide microarrays were evaluated for their ability to differentiate 25 closely related Salmonella isolates [27]. Previously, the same authors used a similar microarray approach to discriminate closely related Xanthomonas pathovars [28]. In this study, we aim to build upon these findings and further the development of oligonucleotide microarrays for use in BST. Here we report the application of a microarray, consisting of 198 oligonucleotide elements, to discriminate 17 unique environmental isolates of Enterococcus sp. based on the host source of the bacteria.A collection of 51 Enterococcus sp. isolates originating from bovine, deer, gull, and human sources were provided by Dr. Shiao Wang (University of Southern Mississippi; Hattiesburg, MS). Details of the isolation and characterization of these strains have been described in detail elsewhere [29]. Isolates were routinely propagated in brain heart infusion liquid media (Becton Dickenson, San Jose, CA). High molecular weight genomic DNA for PCR analysis was obtained from each isolate using Qiagen’s DNeasy Tissue Kit (Qiagen, Valencia, CA).PCR primer BOX A 1R 5′ CTA CGG CAA GGC GAC GCT GAC G 3′, was custom synthesized by Qiagen and targeted repetitive extragenic palindromic BOX sequences [19]. Primer BOX A 1R was used to amplify select portions of the Enterococcus sp. isolate genomes to be used as target DNAs for microarray analysis. All PCR reactions and their subsequent microarray analysis were carried out in triplicate. Final reaction conditions were as follows: 10mM Tris, pH 8.3, 50mM KCl2, 4.5mM MgCl2, 0.001 (w/v) gelatin, 0.2mM dNTP’s, 2μM BOX A 1R primer, and 5U Taq polymerase (Promega, Madison, WI) in a final reaction volume of 100μl. A total of 100ng of genomic DNA was used as template for each reaction. Amplification was carried out in a MJ Research Tetrad thermocycler (MJ Research, Inc., Waltham, MA) programmed as follows: initial step at 95°C for 2 min followed by 35 cycles of: 94°C for 3 sec, 92°C for 30 sec, 50°C for 60 sec, 65°C for 8 min and finally cooling to 4°C at the end of the last cycle. Ten microliter portions from each reaction were electrophoresed through a 1.0% agarose gel in 1x TAE (40mM Tris-Acetate, 1mM EDTA) running buffer and stained with Sybergold (Molecular Probes, Inc., Eugene, OR) for visualization to confirm amplification. The remaining portions of each amplification reaction were ethanol precipitated with sodium acetate [30] and the resulting air-dried DNA pellets were re-suspended in 20μl Millipore water.PCR products were aminoallyl(aa)-labeled as described previously [31]. Briefly, 3.3μl (3μg/μl) of random hexamers (Invitrogen, Carlsbad, CA) were added to each of the re-suspended PCR products and the final volume brought up to 39μl. The sample was heated to 100°C for five minutes and immediately placed in an ice bath. Twenty units of DNA polymerase I Klenow fragment (New England BioLabs, Beverly MA), 5μl of EcoPol (Klenow) buffer (New England Biolabs), and 2μl of 3mM dNTP/aa-labeling mix [100mM each dNTP, 50 mM aa-dUTP (Ambion, Austin TX)] were added to the reaction and the reaction was incubated at 37°C overnight. The reaction was stopped by adding 5μl of 0.5M EDTA. Unincorporated aa-dUTPs and free amines were removed from each reaction using the QIA quick PCR purification (Qiagen) kit with the following modifications: PE wash buffer was replaced with a 5 mM KPO4, 80% ethanol solution and elution buffer was replaced with a 4mM KPO4 solution. Purified PCR templates were dried down in a vacuum centrifuge and resuspended in 4.5μl of 0.1M Na2CO3 buffer, pH 9.3. DNA samples were labeled with a Cy5 dye by adding 4.5μl of a Cy5 mono-Reactive Dye Pack solution (Amersham Biosciences, Piscataway, NJ) and allowing the reaction to proceed in the dark at room temperature for two hours. The reaction was stopped by the addition of 35μl of 100mM NaOAc. Free dye was removed from the samples by using the QIA quick PCR purification kit (Qiagen) according to the manufacture’s instructions. DNA samples were dried down and immediately processed for microarray analysis.One hundred ninety eight 9mers (Table 1), with an amine-modification at the 5′ end, (Sigmagenosys, Woodlands, TX) were randomly selected from a list of 102,403 9mer sequences that conform to criteria described previously [28]. Briefly, 9mer sequences had GC contents between 44–55%, could not have: 1) four nucleotide (or higher) repeats, 2) inverted repeats three nucleotides (or higher), 3) dual-terminal inverted repeats of 3 nucleotides (or higher), and 4) single-terminal inverted repeats of three nucleotides or higher. In addition to these criteria, all 9mer sequence combinations that occurred in Enterococcus sp. rRNA genes present in GeneBank as of 5/03 were eliminated. A Cy3-labeled control oligonucleotide, 5′ TTG GCA GAA GCT ATG AAA CGA TAT GGG 3′, with an amine-modification at the 5′ end, was used as a positional reference and hybridization control.Microarrays were fabricated on aldehyde-coated glass microscope slides (Telechem International, Inc., Sunnyvale, CA) using the BioRad VersArray ChipWriter (BioRad, Hercules, CA) equipped with SMP3 Stealth microspotting pins (Telechem Internation, Inc.). Prior to fabrication, amine-modified oligonucleotides were transferred to a 384-well plate (Whatman, Clifton, NJ) and diluted to a concentration of 80 μM in 50% dimethyl sulfoxide (DMSO). Probes were printed in duplicate, using a 2-pin configuration, at a relative humidity of 60%. The resulting grid pattern and corresponding oligonucleotide probe location is illustrated in Fig. 1. After printing, slides were baked for 45 minutes at 80°C, briefly washed with 0.2% SDS, and subsequently rinsed with reagent grade water. Free aldehyde groups were chemically blocked by soaking printed slides in a fresh NaBH4 solution [0.75g NaBH4 (Sigma, MO), 225 ml phosphate buffered saline (pH 7.0), 66.5ml 100% ethanol] for five minutes. Following chemical blocking, printed slides were momentarily dipped 3 times in 0.2% SDS, washed for one minute in reagent grade water, and individually spun dried in 50ml Falcon conical tubes (Fisher Scientific, MO) at 700rpm for 10 minutes in a tabletop centrifuge. Microarray substrates were stored at room temperature in a desiccator.Prior to hybridization, printed slides were pre-hybridized in 0.1% SDS, 4X SSC (1X SSC, 0.15M NaCl, 0.015M trisodium citrate, pH 7.0), and 10mg/ml bovine serum albumin (BSA) in 50ml Falcon conical tubes at 40°C with slight agitation for 2 hours. Pre-hybridized slides were rinsed 5 times in reagent grade distilled water and chilled to 4°C on a solid metal platform. Cy5 aminoallyl-labelled DNA targets were resuspended in 15μl of 4X SSC, heated at 95°C for 5 min, and immediately placed on ice. The Cy3 labelled oligonucleotide, 5′CCC ATA TCG TTT CAT AGC TTC TGC CA 3′, was also included in the hybridization reaction (final concentration 0.6μM) as a control to hybridize with the control oligonucleotide attached to the microarray. Chilled hybridization reactions were pipetted on prechilled printed microarray slides, covered with array cover slips (PGC Scientifics, Gaitherburg, MD), and incubated overnight at 4°C as described previously [28]. Hybridized microarrays were gently rinsed in 4°C 4X SSC 5 times for 1 minute intervals followed by a final 30 second rinse in reagent grade water. Microarray slides were spun dried in 50 ml conical tubes as described above prior to scanning slides.Processed microarray slides were scanned at 532nm and 635 nm using the VersArray Chipreader system (BioRad, Hercules, CA) configured at a 5μm resolution. Spot intensity data from the resulting 16-bit TIF images were initially extracted using the ArrayPro Analyzer software (Media Cybernetics, Silver Spring, MD). Background signal was determined locally for each spot using the “local corners” option. Individual spot intensities, minus local backgrounds, were normalized to total spot intensity for all of the spots on each micro array. The mean-normalized datasets were transformed by taking the logarithm of these values. An empirical data reduction process was employed (see Results) to identify which of the 198 probe spots had the most information (example: spots that were always “on” or “off” for all isolates would have no information for this dataset) and which of the spots that were too variable within the replicates of the same isolates. Principal Components Analyses (PCA) and cluster and classification analyses were run on the remaining dataset using Pirouette (Infometrix, Inc., Bothell, WA).Oligonucleotide microarrays were evaluated for their ability to resolve BOX PCR amplification products derived from environmental sources of Enterococcus sp. isolates originating from deer, bovine, gull, and human. Purified genomic DNA from Enterococcus sp. isolates was subjected to BOX PCR amplification and the resulting amplification products were visualized by agarose gel electrophoresis. The results of a typical experiment can be seen in Fig. 2, which represents the subset of samples originating from deer. Agar gel electrophoresis confirms amplification as well as consistency of the BOX PCR reaction. PCR products were fluorescently labelled with aminoallyl dUTP and Cy5 then resolved by hybridization to in house fabricated 9mer oligonucleotide microarrays (see Material & Methods). The results of a representative microarray experiment can be seen in Fig. 3, in which replicate BOX PCR reactions from Enterococcus sp. deer isolate 49.1.1 were hybridized to replicate oligonucleotide micro arrays. A histogram of fluorescent spot intensities indicates that these randomly selected nonamer intensities follow a lognormal distribution (data not shown). Of the 17 environmental isolates analysed, not all replicate microarrays were usable. For six of these isolates (4 human and 2 deer) a single microarray hybridization replicate, consisting of duplicate microarray spots, was available for analysis. For the remaining 11 isolates and their replicates, spots that exhibited extreme variability in normalized spot intensities among replicates within a specific source were identified and subsequently eliminated from analysis for all isolates. Normalized spot intensities with above median standard deviations > or = 0.7 within source-specific datasets (i.e. bovine, deer, etc.), were eliminated leaving 45 of the 200 probes. The remaining 45 probes were then used for analyzing all 17 isolates.The dendrogram of a complete Euclidean distance Hierarchical Cluster Analysis (HCA) did not project good origin-specific clustering of the isolates. In particular, the bovine-origin replicates were spread among several clusters (example part of dendrogram Fig. 4). A K-Nearest Neighbour classification confirmed the HCA, misclassifying 8% of the deer, 16% of the human, and 50% of the gull isolates as bovine isolates. The PCA was visually superior at separating origin-specific clusters, even for as few as 3 factors (Fig. 5). A Soft Independent Modelling (SIM) classification confirmed the PCA, resulting in zero misclassifications using 5 factors for each class. Numerical descriptions of the SIM classification model for bovine-origin Enterococcus sp. are presented in Table II. These factors describe the multidimensional subspace within the PCA projection in which the various microarray source profiles exist. Factor numbers indicate the relative linear weights of each probe in each factor. For instance probes 2 and 16 have the highest weights for the most important factor, Factor 1, which accounts for 30% of the variability. Thus for this set of isolates, SIM classifications based on 5 factors for each class and 5 linear combinations of the 45 probes sufficed to distinguish the origins of Enterococcus sp. isolates.In an effort towards adapting new defensible methods for assessing and managing the risk posed by microbial pollution, we evaluated the utility of oligonucleotide microarrays for bacterial source tracking. Specifically, we evaluated the ability of oligonucleotide microarrays to visually discriminate 17 unique environmental isolates of Enterococcus sp. based on host origin, i.e. gull, bovine, deer, and human. As observed in an earlier study by Kingsley et al. [28], many of the microarray oligonucleotide probes exhibited high variations in fluorescent spot intensities within a series of replicates. A strong down selection for reproducible spot intensities within replicates produced a set of 45 probes, and this reduced set proved useful for classifying isolates by source. It should be reiterated that this data reduction was performed in order to improve reproducibility, and had the side effect of improving the classification fit. This is the opposite of the familiar problem of model over fitting, in which the addition of extra variables improves classification at the expense of robustness and reproducibility.Following data reduction, a number of multivariate statistical analysis procedures are available for evaluating the relationships among microarray hybridization profiles. Previously, PCA was successfully used to visualize relationships among microarray hybridization profiles derived from closely related Xanthomonas pathovars [28]. In this study, PCA and HCA were compared for their ability to visually cluster microarray hybridization profiles based on the environmental source from which the Enterococcus sp. isolate originated. Classification of Enterococcus sp. isolates by source using a Soft Independent Modelling of class analogies consisting of 5 factors was more accurate than classification based on K-Nearest Neighbour calculations. This difference is apparent when comparing the PCA, which is a visualization of some of the SIM calculations, to the HCA, which is a visualization of some of the KNN calculations. The implication of these results for the application of random oligonucleotide microarrays for BST is that, given the reproducibility issues, factor-based variable selection such as in PCA and SIM greatly outperforms dendrogram-based similarity measures such as in HCA and KNN. Given any sample based strictly on the microarray intensity values, the SIM model outputs the best fitting class for that sample, with zero misclassifications for the dataset. Further optimization of source classifications may result from the application of information theory to detect patterns in microarray profiles. In particular, bacterial source tracking may benefit from several measures of classification utility, such as those based on mutual information that have been developed as part of information theory [32]. However, successful application of information theory for microarray analysis will be dependant upon accurately understanding, capturing, and modelling sources of variation in the microarray experimental process. Some of these sources of variation, such as PCR amplification and microarray fabrication have been described previously [27]. Once improved microarray experimental protocols and statistical methods have been developed, it will be possible to incorporate microarray technology into the growing toolbox of technologies that is rapidly defining bacterial source tracking. While there is currently no one best method that accomplishes the ambitious goal of source tracking as demonstrated in the latest study by Stoeckel et al. [33], it is likely that a combination of methods will lead to effective source tracking.Configuration of the printed microarray spots and the physical location of the corresponding oligonucleotide probes as referenced in Table 1. Control oligonucleotide designated by C.BOX-PCR agarose gel fingerprints run in triplicate from Enterococcus sp. isolates originating from deer. A HindIII digested Lambda marker was included in the gel run as a size standardOligonucleotide microarray replicate hybridization profiles resulting from hybridization with BOX-PCR amplification products from deer isolate 49.1.1.Hierarchical Cluster Analysis of normalized microarray spot intensities of replicates of 17 environmental isolates of Enterococcus sp. The dendrogram does not show good clustering by host origin at reasonable similarities. The bovine-origin replicates were most spread.Principal Components Analysis of normalized microarray spot intensities of replicates of 17 environmental isolates of Enterococcus sp., colored by host origin: deer is red, bovine is yellow, human is green, gull is purple. For this 3D view only the first 3 components can be plotted, but clustering is evident.Microarray Oligonucleotide Probes5′ amine-modified 9mer oligonuceotide microarray probes and corresponding I.D. numbersThe 5-factor oligonucleotide microarray SIM classification model for bovine-origin EnterococcusThis work was funded by the Long-term Effects of Dredging Operations (LEDO) program at the U. S. Army Engineer Research and Development Center (ERDC). Permission was granted by the Chief of Engineers to publish this information.For list of impairment by state visit http://www.epa.gov/OWOW/tmdl/Term first used by Hagedorn and Wiggins at http://www.bsi.vt.edu/biol_4684/BST/BST.html
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+ Despite consented efforts in prevention, mycotoxins remain a problem of human health concern in several parts of the world including developed countries. Within the same range of toxins concentrations in the blood some people develop a disease while others do not. Could this inequality in front of mycotoxins effects be explained by environment factors and/or genetic predisposition? Among recent advances in environmental health research Correlation between chronic diseases and mycotoxins in humans deserves attention through several questions: Are genetic factors involved in disease causation of mycotoxins? How much are these factors currently taken into account for mycotoxins risk assessment and how much should we involve them? Answers are still to come. Genetic and environment factors deserve therefore more attention when dealing with regulatory limits, since among the general population, those who are at risk and will develop specific diseases are likely those bearing genetic predispositions. We have addressed these questions for the specific case of ochratoxin A in humans by investigating in Tunisia, county of Jelma, in four rural families forming a household of 21 persons all exposed to ochratoxin A in diet. Our results confirm that ochratoxin A induces chronic tubular nephropathy in humans and mainly point at those having the HLA haplotype A3, B27/35, DR7 to be more sensitive to the disease for quantitatively similar or lower exposure. Persons with such haplotype were found to bear chronic interstitial nephropathy with tubular karyomegalic cells while others were apparently healthy. Godin et al. (1996) in France have also found in sibling (a sister and her brother from urban area) that have similar HLA haplotype B35-patern, OTA-related renal tubulopathy with mild proteinuria including β2-microglobulinuria. Several mechanisms are discussed that could be put ahead to explain how the HLA haplotype could lead to tubular cells lyses and renal failure. In the mean time it is urgent to search for mass screening biomarkers for mycotoxins in humans and related genetic factors to set-up more appropriate regulation.The most frequent toxinogenic fungi in the world are Aspergillus, Penicillium and Fusarium species. They produce Aflatoxin B1 transformed into Aflatoxin M1 found in milk, Ochratoxins and Zearalenone, Fumonisin B1, T-2 toxin, HT-2 toxin and Deoxynivalenol (vomitoxin) that are increasingly of human health concern [1–2]. Some of their metabolites are still toxic and may be involved in human diseases, they are not destroyed by normal industrial processing, [2–7]. Their toxic effects, liver, kidney and hematopoetic toxicity, immune toxicity, reproduction toxicity, foetal toxicity, teratogenicity, and mainly carcinogenicity are mostly known in experimental models, the extrapolation to humans being sometime impossible due to lack of appropriate tools and biomarkers [2, 8–11]. The inaccuracy of extrapolation of experimental data to human may be explained by several reasons: (i) discrepancy between human and animal exposure, (ii) lack of precise health risks associated with specific proposed limits, (iii) possibility of synergism or antagonism with other mycotoxins present in the same diets [2, 4, 8, 12], (iv) coexistence of other pathologies such as viral hepatitis, immune or hormonal deficiencies or organs dysfunction. In contrast to experimental animals, humans are placed in more complex environment in which their genetic imprinting may be influenced, modulated and/or modified by many factors including mycotoxins.Even when a specific biomarker of a given mycotoxin is identified in humans, it remains difficult to establish the relationship with a given illness because of above quoted reasons and because of the genotoxic and epigenetic effects of mycotoxins. These can be strongly influenced by the metabolome as the expression of individual genetic polymorphism. Actually for example, in anyone continuously exposed to aflatoxin B1 hepatocarcinoma would be expected. However, according to environment factors and to his individual genetic polymorphism, some other cancers or diseases may be found. Possible beneficial influence of diet may mask or counterbalance the expected pathology as previously observed in experimental models.People may develop diverse degree of the same disease for similar exposure to mycotoxins. It is clear that individuals do not react the same way for similar exposure to harmful mycotoxins. If one compares OTA blood levels from certain Balkan Endemic Nephropathy (BEN) areas and some regions in Western European countries, it appears that while some people have BEN there is apparently no disease in Western Europe for similar blood concentrations [2–3, 13]. It is presently believed that this discrepancy is explained by diets and eating habits. Is that enough?How these facts influence the risk assessment in humans and how much should we take them into account? The toxic effects of mycotoxins are currently explained by several mechanisms, (i) direct cytotoxicity of the parent compound and derived metabolites, (ii) DNA damages leading to mutagenesis, (iii) indirect DNA bases modifications such as oxidation and/or methylation leading to epigenetic promoter pathways This applies to aflatoxin B1, ochratoxin A, Zearalenone and partly to fumonisin B1 since no direct genotoxic effects have been demonstrated for the latest, [2, 13–15].In these genotoxic and epigenetic promoter pathways, several enzymes are involved such as repair enzymes. Modification occurring in the encoding genes may modulate and/or reduce their activity. Furthermore it is accepted that mutations occurring in key genes such as p53 gene mutation, frequent in human tumours, or in the genes encoding the proteins Bax or Bcl2 or in oncogenes [16, 17], will strongly influence cell survival versus cell death thus promoting casually cell transformation and tumour genesis.All these pathways cannot exclude either the participation of endogenous effectors such as hormones (some mycotoxins being moreover hormonal disrupters such as zearalenone) or that of xenobiotics present in our environment, (air, water, and diet) possibly acting in concert with mycotoxins [2, 18]. Several mechanisms of action have been demonstrated in experimental models and in humans that do not take into account genetic pre disposition of people either before exposure, and/or after exposure to environment factors. This applies actually to the rationales for the establishment of limits and regulations for mycotoxins [19]. For all the above reasons we have to address the question: how much should we involve genetic and environmental factors in the risk assessment of mycotoxins in humans?Taking as example the case of human intake of ochratoxin A in which it is obvious that not all heavily exposed persons are ill, the implication of genetic pre disposition has been investigated in Tunisia (specially the county of Jelma) that is known to be a high spot of OTA exposure in Northern Africa. Ochratoxin A (OTA) (Fig. 1) is produced by several species of fungal genera [1] that is widespread in human food, animal feed and detected in blood and tissues in Tunisia as well as in the Balkans. This mycotoxin has several adverse effects the most prominent being nephrotoxicity and carcinogenicity [2–3, 7, 9].Exposure to OTA has been linked with BEN, a chronic kidney disease associated with urothelial tumours, [2, 4, 20]. In Tunisia, a Chronic Interstitial Nephropathy (CIN) of unknown aetiology bearing striking similarities with BEN in its pathogenic characteristics was under investigation for more than one decade for implication of OTA [2, 7, 21]. Taking into account the high prevalence of ochratoxin A in blood of Tunisian people and the above quoted disease as compared to other kidney diseases and, considering that healthy populations are equally exposed, links were established that tend to demonstrate the causative role of OTA in ill persons [5, 7, 21].Whether the causative role of OTA in inducting this human nephropathy is correlated to some specific genetic imprinting of ill people is an obsessing question for risk assessment. To enlighten this point, investigations were designed on a rural household, consisting of 21 people for kidney impairment, OTA assays in food, blood and urine in addition to HLA (Human Leucocyte histocompatibility locus A) fingerprinting.People presenting the clinical signs of CIN related to OTA similarly to BEN that are lombalgia, polyuria, and absences of oedema, hypertension and hematuria. Echography and radiography also reveal small size-kidneys with asymmetry and irregular renal contours [3, 7, 21]. They live in a rural environment in the western centre of Tunisia (Jelma). This area belongs to a region that was found to be one of most affected Tunisian areas of high OTA contamination, according to previous investigations and even considered as an area of endemic OTA-related nephropathy [2, 5, 7]. Indeed, high concentrations of OTA in blood were found. This region is warm and humid.The prospected community is made up of 21 members (11 women and 10 men, aged from 32 to 65 years) of four families (see genealogical tree, Fig. 2). For each subject, blood and urine samples were collected for OTA assays and for the evaluation of the renal function. For this nephropathy, several etiological agents have been incriminated seriously considered and disqualified. The cases that remained without any possible aetiology are categorised as CIN of unknown aetiology presumably related to OTA since only this causal agent remained correlated to the disease [5, 7].β2-microglobulin was quantified in urines using an immunoenzymatic competitive test in microplate coated with monoclonal antibody, in the presence of a tracer consisting in a β2-microglobulin conjugated with alkaline phosphatase. After washing to remove excess tracer bound enzymatic activity is measured by the colour developed by a chromogenic substrate in a plate reader. The concentration of β2-microglobulin is then obtained by a standard curve built using the material provided with the ELISA kit by Immunodiagnostik, Benscheim, Germany.Blood samples were collected in heparinized tubes. After centrifugation, (850g/10min at 4°C) the plasma samples (2ml) were frozen at −80°C until analysis. The plasma and urine samples were acidified with acetic acid 1M (pH 4.5) then cleaned on C18 Sep-Pak cartridge (10 × 20mm), Waters (Supelco, Sigma-Aldrich, France). After washing with acetic acid 2% (v/v), OTA was eluted with HCl 1% in methanol (v/v). Samples were then separated by HPLC system and quantified by fluorescence detection (Varian Pro Star) with an excitation wavelength of 340 nm and an emission wavelength of 465nm at ambient temperature according to known quantities of standard OTA as previously described, [2, 5, 7, 21]. Detection and quantification limits were confirmed to be 0.1ng/ml and 0.5ng/ml [7, 9].For each family, three types of daily-consumed basic food (cereals, cereal made-foods and beans) were collected for OTA assays. OTA-positive samples were confirmed by hydrolysis of OTA by 100μl of carboxypeptidase (Bovine pancreas, Sigma) (100 IU/ml) in the buffer Tris-HCl 0.04 M (pH 7.6) and NaCl 1M during 2h at 37°C and analysis by HPLC using the conditions described above. Ochratoxin alpha (OTα) appears instead of OTA [5, 7, 21].Light microscopy observations were performed with magnification of 400 on sections of kidney biopsies stained using trichrome de Maçon dye to confirm karyomagaly in renal tubules [7, 22] that are BEN-like lesions.On blood samples, HLA-A and HLA-B antigens have been typed by a standard microlymphocytotoxicity technique using a panel of standard sera (One Lambda, CA).The 21 subjects were selected from the rural population of Tunisia in the region of Jelma, known for high OTA exposure.OTA assays in blood showed that 19 persons are contaminated among the 21 prospected in the range of 8ng/ml to 1468ng/ml, (Table 1). High OTA concentrations were found in blood of the three cases (n°1, n°2 and n°3), respectively 505.83ng/ml, 102.63 ng/ml and 1023ng/ml. Higher OTA blood concentrations were found in the blood of other members of the family I (n° 7, 1332ng/ml) and from family II, (n° 9 and n°11, respectively 1348 and 1334ng/ml), (Table 1). The ill persons are three siblings, two brothers and their sister which kidney biopsies showed karyomegaly in renal tubules, Fig 2. They have altered renal function, high uremia and creatinemia, without hypertension, β2-microglobulinuria was respectively 10400μg/l, 440μg/l and 360μg/l. All these signs are characteristic of BEN, (Table 2). Some others have more OTA in food and blood although they do not show any sign of BEN, (Table 2).The 21 poeple were tested for their HLA haplotype because this is known to be implicated in renal tubulopathy and karyomegaly in human by several groups working on the subject [23–25]. The three cases showing BEN-like nephropathy with high OTA concentrations in blood and urines belong respectively to the haplotype, A 3/28; B 27/35; DR 7/52; A 3/28; B 27/35; DR 7/11 - A 3/40; B 27/35; DR 7/52. They share the following elements of phenotype A3, B27/35 and DR7.In these investigations, it is obvious that all people from rural areas enrolled are exposed to ochratoxin A. The two individuals considered to be negative have shown in their blood and urine OTA values that fall in detection and quantification limits and were not confirmed further as previously described. They need to be followed-up to know whether their low exposure remained low or not. The urine OTA-concentrations of other apparently healthy people fall also in this range. However their case needs to be discussed according to the knowledge we have concerning the elimination of OTA in urine. OTA is actually eliminated by glomerular filtration, tubular re-absorption and tubular secretion, [2]. Those who have limited amount of OTA in urine have likely no alteration at any of these processes. Those who have appreciable OTA concentrations in urine without symptoms of tubulopathy have likely asymptomatic glomerular disease that leads to urinary elimination of proteins that bind more than 90% of available OTA in blood [2]. They are under investigation for that reason. All asymptomatic subjects also need to be seriously surveyed. Only three of OTA-positive people are ill, showing BEN-like nephropathy although others have similar OTA concentrations in blood and urine. Could this be explained by environment factors only or genetic factors only or both? It is to be emphasised that for ethical reasons no biopsy has been performed on those who were apparently healthy.The ochratoxin A concentrations found in either food or biological fluids during the present investigations in Tunisia are even higher than those observed in BEN areas in the Balkans. These should then normally elicit kidney tubules damages, provided that people are exposed for long enough. There is no doubt that most individuals enrolled in the present investigations have been exposed for many days or weeks according to the half-life time of OTA in human body (more than 30 days) and to the concentrations found that cannot likely be ingested in a single dose. The two people who do not have OTA in blood could have ingested also substances bearing the capability of preventing OTA tissue distribution, metabolism and of enhancing its elimination such as OTA-binding peptides, Aspartame, antioxidant such as tea, inhibitors of prostaglandin synthetase, etc., [2].The renal lesions in ill people are clearly established, since β2-microglobulinuria is high and karyomegaly is histologically demonstrated. This has been also found in north of France in sibling (a sister and her brother from urban area) that bear similar HLA hoplotype B35-patern, [25] and also mild proteinuria including β2-microglobulinuria. This implicates a particular genetic imprinting and evidence for genetic defects, [24, 25]. The two individuals (n°8 and 15) who have elevated β2-microglobulinuria seem to be very atypical cases for a reason we now ignore. Their HLA haplotype is under verification.Ochratoxin A is nephrotoxic to all animal species tested so far including pigs and induces karyomegaly in proximal tubules in rats, [11, 22]. These data have to be put in parallel with data in humans, [5, 7, 20, 21, 25] and karyomegaly has to be considered as a marker of pathology in both humans and animals exposed to OTA. Two mechanisms could be put ahead to explain how the HLA haplotype could lead to tubular cells lyses and renal failure. First, Hanada et al. 2004, propose that cytotoxic T lymphocytes (CTLs) could detect and destroy cells displaying modified class I molecules of the major histocompatibility complex, [26]. OTA is known to be mutagenic and may induce modification in HLA complex favouring renal tubule cell lyses and/or karyomegaly in tubular regenerating cells, [2, 3, 15, 20–22]. Second, Morozov et al. 1991, propose a possible mechanism for the release of beta 2-microglobulin from HLA complex to make it available to exert toxic effects on the kidney as a potential pathogenesis of Balkan endemic nephropathy, [27]. Both mechanisms may be co-acting on the renal tubules of people having a given HLA haplotype and high beta 2-microglobulinuria. One may then hypothesize that genetic pre-disposition are involved in this disease causation.Other genetic polymorphisms or pre-dispositions that also can influence human genotoxic risk are those of glutathione s-transferase and cytochromes P450, suggesting involvement of oxidative stress, [2, 10, 11, 28]. Recent findings have demonstrated that oxidative damages contribute to the cytotoxicity and carcinogeneticy of mycotoxins such as ochratoxin A, aflatoxin B1 (AfB1), fumonisin B1 and Zearalenone [2, 10–12, 16]. The case of aflatoxins is comparable to that of ochratoxin A in several aspects. Aflatoxins occur in peanut butter, cereals and their metabolism requires allelic polymorphic enzymes such as glutathione-S-transferases encoded by glutathione-S-transferase gene (GSTM1) and glutathione-S-transferase gene (GSTT1) and microsomal epoxide hydrolase encoded by epoxide hydrolase gene (EPHX). The rate at which aflatoxins become activated or detoxified may depend on polymorphisms in the encoding genes. GSTM1 homozygous deletion was indeed found to modify the association between peanut butter consumption and hepato-cellular carcinoma (HCC) [6, 10]. Possible roles of GSTT1 and EPHX polymorphisms in this relationship are well documented [2, 6, 10, 11, 17, 28]. Whether humans vary in their ability to detoxify the active intermediate metabolite of AfB1, AfB1-exo-8, 9-epoxide, is not certain but may explain why all exposed individuals do not develop HCC [6]. The situation is similar for ochratoxin A. If it becomes obvious that in an exposed population to mycotoxins, those who will develop specific diseases are those bearing some genetic pre-disposition or defaults, it would be valuable and indeed it becomes urgent, to take these factors into account for prevention, risk assessment and management. In the mean time, while waiting for more appropriate regulatory limits, the common sense suggests that prevention involves first, reduction of mycotoxins levels in foodstuffs and further implication of diet components such as vitamins, antioxidants and substances known to prevent carcinogenesis.Currently, environmental factors are those mostly evaluated to determine human exposure. This procedure appears largely insufficient to definitely establish the aetiology (links) of a given intoxication and/or disease with probable toxic agents. The present findings suggest that biomarkers might be more useful tool for epidemiological research in mycotoxins exposure in humans in relation to diseases provided that genetic factors are also taken into account in addition to environmental factors. The whole would then represent a valid instrument for group evaluation. To this objective it is necessary to demonstrate a stronger association with diseases that perhaps the future development of genetic and molecular epidemiology will make possible.Since there is a real lack of biomarkers for mycotoxins, a review of main methodological questions regarding biomarkers will be needed. This should focus on protocols finalized for the study of multiple panels of cytotoxicity and genotoxicity biomarkers. These should take into account the influence of environment toxicants at low doses, biomarkers associated to risk and genetic polymorphism. In this context it should be suggested to evaluate by genotyping and phenotyping the proportion of the population that bears similar haplotype since they are those who seem really to be at risk for diseases related to ochratoxin A according to life-style. The opportunity is given in northern African to follow up, in cohort study, ochratoxin A-contaminated populations and those who are already in the danger list.Structure of ochratoxin AGenealogical tree of people included in the study.Ochratoxin A contaminations, in food, blood and urine of the 21 exposed people, and β2-microglobulinuria.HLA haplotype of patients CIN BEN-like and daily intake of OTA for the subjects calculated from the plasma clearance, blood concentration and bioavailability of OTA in human according to the following equation,
2
+
3
+
4
+
5
+
6
+ K
7
+ =
8
+ Cpl
9
+ ×
10
+ Cp
11
+ /
12
+ A in which K= daily intake, Cpl= OTA plasma clearance ml/min, Cp=OTA plasma concentration and A= bioavailability.
Med-MDPI/ijerph_1/ijerph-02-02-00194.txt ADDED
@@ -0,0 +1 @@
 
 
1
+ The complexity of interactions in socio-ecological systems makes it very difficult to plan and implement policies successfully. Traditional environmental management and assessment techniques produce unsatisfactory results because they often ignore facets of system structure that underlie complexity: delays, feedbacks, and non-linearities. Assuming that causes are linked in a linear chain, they concentrate on technological developments (“hard path”) as the only solutions to environmental problems. Adaptive Management is recognized as a promising alternative approach directly addressing links between social and ecological systems and involving stakeholders in the analysis and decision process. This “soft path” requires special tools to facilitate collaboration between “experts” and stakeholders in analyzing complex situations and prioritizing policies and actions. We have applied conceptual modeling to increase communication, understanding and commitment in the project of seven NGOs “Sustainable Regional Development in the Odra Catchment”. The main goal was to help our NGO partners to facilitate their efforts related to developing sustainable policies and practices to respond to large-scale challenges (EU accession, global changes in climate and economy) to their natural, economic and socio-cultural heritages. Among the variety of sustainability issues explored by these NGOs, two (extensive agricultural practices and “green” local products) were examined by using Adaptive Management (AM) as a framework that would link analysis, discussion, research, actions and monitoring. Within the AM framework the project coordinators used tools of systems analysis (Mental Model Mapping) to facilitate discussions in which NGO professionals and local stakeholders could graphically diagram and study their understanding of what factors interacted and how they affect the region’s sustainability. These discussions produced larger-scale Regional Sustainability Models as well as more detailed sub-models of particular factors, processes, and feedback loops that appear critical to a sustainable future. The Regional Sustainability Model was used to identify a subset of key interacting factors (variables). For each variable, several sustainability indicators were suggested. The growing understanding and acceptance of the AM framework and systems analysis created a momentum both locally and within the region, which makes continued successful use of these indicators quite likely. In contrast to expert-driven projects that inject outside knowledge into a local context, this project established a broad basis for stakeholder-driven discussion that is articulated into goals, objectives, conceptual models, and indicators. The ability to learn and adapt in the AM framework increases the capacity to innovate and find policies and practices that enhance resilience and sustainability in a world in transition.The complexity of interactions in socio-ecological systems makes it very difficult to plan and implement policies successfully. One of the main reasons for this is the uncertainty emerging not only from complex interactions within different sectors (for example, academia, government, business), but also from the tangle of relations across ecological, economic and socio-political domains. The challenge to understand and manage complex systems emerges in a history of surprising reversals of initial policy success [1, 2]. At first, attempts to eliminate, and then to merely control disturbances (flood, fire, and pests) have often promoted larger and more profound disruptions. Stubborn resistance to most policy remedies has earned such problems the title of “wicked problems” [3], as if evil intention is a metaphor for how intractable, unknowable and uncooperative the world is. Blame for rising flood statistics or declining river valley economies cannot simply be pinned on “the usual suspects”: exogenous drivers or ignorant human actors or policies. Analysis of the underlying complexity continues to improve, but understanding, and more importantly the capacity to adapt, remains woefully behind the evolving reality. The move from the “hard” and narrow technical approach to a more adaptive and comprehensive “soft” path [4] requires not so much methods of analysis or management intervention, but their integration.Coping with uncertainty requires the sustained capacity to learn and to flexibly manage. For thirty years a decision making process has been evolving to address the challenge of learning while managing. This process, Adaptive Environmental Assessment and Management (AEAM), also known as Adaptive Management (AM), offers a framework to integrate research, policy and local practice that has been developed over three decades of experimental applications to understand and manage crises of collapsed fisheries, agriculture, forestry and rangeland grazing [5–11]. AM increases adaptive capacity by shifting linear decision making processes (crisis - analysis - policy) to a cyclic learning process that iteratively integrates how we modify assessment, policy formulation, implementation and monitoring in order to track and manage change in the world (Figure 1).The AM learning cycle usually starts with an Assessment phase wherein stakeholders explore a range of assumptions and ideas in order to formulate a suite of equally plausible hypotheses that provide separate predictions of why the problem in question occurs [8, 12]. Modeling can serve as a useful exercise for stakeholders participating in the AM process to bound the problem and examine the key variables and interactions they consider crucial to the dynamics of resilience and vulnerability in the system. Conceptual models facilitate discussion and comparison of different interpretations of the system’s structure. Such models allow participants to graphically analyze and discuss which variables are involved and how are they linked, including identification of reinforcing and balancing feedback loops and delays that affect system dynamics [1, 13]. Graphic tools such as diagrams and mental maps open the discussion of complex systems to include people who find verbal descriptions too complicated or too long and involved. Often a single map replaces pages of text required to describe all of the variables and their interactions.Monitoring and Evaluation constitutes the link closing the AM cycle. Monitoring is usually done through defining and measuring different indicators. Indicators can be developed in a top-down expert driven way such that a uniform set of indicators is equally measured in different locations involved. Quite often, however, the impact is more academic than practical and does not lead to any meaningful change on the ground but only to comparing the indicators’ evolution over time. Alternatively, indicators can also be prepared as a means of facilitating local communities’ learning about sustainable development [16–19]. This approach usually requires public or stakeholder participation in creating and using the indicator database [20–28]. The level of participation can vary – from manipulative, wherein it is only pretence and the local stakeholders have no influence on the decision process, to self-mobilization, wherein local people initiate actions themselves. In the latter case, when people design and monitor their own indicators, they develop a better common trust and experience with which to interact with professionals and higher authorities [29, 30]. However benignly intended, the interference of “expert” external agencies can stifle the development of trust and cripple the long-term acceptance and implementation of innovative policies.In this article, we describe the modified Adaptive Management Framework that we used in the Barycz Valley, first from the conceptual, and then the practical point of view.In our approach we have adopted the Adaptive Management Framework (AMF) and enhanced it using advances in conceptual modeling and sustainability indicators practice (Figure 2). The framework can be classified as a “soft approach” in the following ways.An open, participatory and recursive process both for policy formulation and indicators selection is used instead of top-down control.Systems analysis including many feedbacks between sectors is performed, instead of narrow technical analysis.Conceptual, qualitative modeling is used instead of formal, quantitative modeling.Conceptual systems thinking techniques have been developed as a reaction to the failure of quantitative systems analysis to cope with the so-called “messy” problems, where it was difficult to identify a clear goal to attain. Different methodologies have been developed to tackle these problems [31–33]. Their usefulness has been verified in a variety of different contexts and applications [34]. The soft approach has made it easier to engage the client or the public in the process of group model building.Bell and Morse [20] in their work on sustainability indicators suggested the usefulness of a participatory approach linked to the broader perspective of participatory learning using systems thinking. One of the most important conclusions of their work was that sustainability indicators have to be developed in an open and participatory way which helps the community directly learn about its performance and thereby improves the decision making process. The indicator sets are more flexible and adapted to the specific stakeholders’ needs. Moreover, the participative process makes any potential review and continual improvement of the indicators not only possible but also a desired part of the process.There are however some drawbacks. The main problem they encountered with the soft approach was that it was not “easily reportable or demonstrable to auditing authorities” [21]. Another problem may arise when a team working on indicators feels the soft systems approach is not rigorous or professional enough (because it does not provide quantitative results).We tried to avoid or mitigate these weaknesses in two ways. Firstly, we used a structured and recursive learning process within an AMF to identify and address errors in a transparent way. Secondly, the same framework allowed us to integrate different phases, using them to challenge and reinforce one another. For example, we meshed the phases of bounding and then measuring the problem by linking the results of the conceptual modeling phase with the process of defining indicators. As conceptual modeling provides the “big picture” of the problem and helps to overcome human information processing limitations, it also makes the process of indicators selection more rigorous and leads to more comprehensive results.We have adapted the general AMF to suit the needs of the project (Figure 2). The modified framework consists of the following steps.In this first step the group identifies, discusses and agrees on the issues of most concern, such as crime, education and the environment. Conflicting opinions and uncertainty are used as valuable signals as to what areas need investigation most. Because Adaptive Management is designed to review and revisit every phase, it is not necessary to deal with every possible issue, but simply make a good start that unites participants in what they agree that they know, and what they understand that no one knows.Identifying Variables and Interrelationships. Here conceptual modeling is used to map the underlying assumptions about the linkages and causality in the system. We have used the qualitative system dynamics methodology with causal loop diagrams as mapping tool.Assessing Major Uncertainties and Unknowns. Disagreements reveal gaps in understanding. Uncertainties are pondered to the point when they can be clearly stated as hypotheses.Identifying Key Variables. Using the conceptual model developed in 2a, most important (key) variables are selected by considering the number of interactions and/or delays as well as employing a conservative rule that each feedback loop should be represented in the set of indicators by one of its variables.Deriving Indicators for Each Variable. Each key variable should be represented by at least one indicator. Often multiple indicators are needed to capture the range of values and qualities associated with a variable.Scoring Indicators with Three Sets of Criteria. The scoring process must be streamlined and simple enough to be easily understood and relatively rapid to accomplish. Criteria should also help one examine what makes an indicator useful and convincing. To meet these goals a set of three criteria was employed: importance (work group’s perspective), compellingness (stakeholders’ perspective) and measurability.Selecting a Final Set of Sustainability Indicators (based on cumulative scoring).Plausible competing policies are formulated (treated as hypotheses) aimed to achieve objectives chosen by stakeholders and to achieve targets identified in the process of defining indicators. Based on a common agreement regarding prioritization of needs, only a few policies are chosen (a subset that is small enough to be practically and thoroughly applied) for further implementation.Actions necessary to realize the chosen policies are planned and implemented.Information is gathered to further review the appropriateness of the indicators chosen in 3. The output and impact of chosen policies are measured and evaluated as to whether action plans and management interventions have achieved the targets specified in policies.The Barycz River, with a total length of 133 km, is one of the largest tributaries of the Odra River, which in turn is the second largest river in Poland. A basin topography that combines a flat lower valley with steep slopes in the surrounding hills results in diversified habitats with a mixture of forests, meadows and ponds, which occur both in the form of large and small complexes. The total river valley covers an area of 2600 km2 and it is administratively divided among 17 local communities (municipalities). The hydrological system supports Europe’s biggest fishponds – the Milicz Ponds. Large migratory flocks of birds concentrate seasonally in the whole Barycz Valley. As many as 276 bird species (166 breeding) have been recorded in this area.Changing political and economic conditions resulting in rising unemployment pose many threats to biodiversity in the Barycz Valley. The challenge is to preserve biodiversity and at the same time improve the local economy. To this end a broad coalition of NGOs active in this region has been established to explore a range of policies and practices to promote biodiversity and environmental quality in the Barycz Valley. The specific challenge arising from this is to help organize and integrate such a variety of organizations pursuing a diversity of different objectives and activities in a way that different projects individually contribute to progress toward sustainability.Efforts to use external experts to inject knowledge into local situations may add valuable experience from similar situations elsewhere, but the full practicality of such information is seldom realized if it is paternalistically handed down to “clients”. Frustration over failures for local people to accept or utilize such expert knowledge, let alone innovate, has led to the deeper question - how to empower local professionals and stakeholders to better use their own experience to create their own policies and practices, to measure their own progress, and to continue to learn and innovate in this regard in order to adapt to a world of changing climate and economy? People thus empowered might be better prepared to adapt ideas and experience from abroad to their local reality, because they are capable of detecting failures and improving them. In short, they can learn by themselves and, once secure on the local level, can responsibly interact with higher levels in ways that will sustain development without violating their heritage. Our project was supposed to provide local professionals and other stakeholders with tools and methods that would help them solve these problems.This project pursued the goal of establishing a framework for different NGOs to discuss their diversified visions and approaches and develop their own indicators and targets by which they could measure their progress toward sustainability. In this way they can share and enrich each other’s experience and continue to develop a common vision that unites actions in the region. Adaptive Management has been used as a general platform to link discussion and research during the whole project. It created the red thread through the steps of the project and helped participants to answer the question “where are we?” whenever they felt lost. Within our AMF, Conceptual Modeling has been used for mapping mental models of participants and stakeholders. Together with the project participants, we came up with an initial list of Sustainability Indicators to demonstrate to participants their use and potential effectiveness. The indicators can be improved in the ongoing process of formulation, measurement and revision.The project proceeded in a series of workshops with NGOs. Local stakeholders and students occasionally joined the project group meetings and discussions. Professionals in local NGOs are good contributors to the start of an adaptive discussion framework, helping initially to absorb new ideas and methods, and subsequently to act as bridges of understanding to local stakeholders with whom they have established trust over the years. The NGOs were then responsible to pass the new knowledge on to the local people and thus, also involve them in the project. Direct involvement of stakeholders would make the project far more costly and time consuming.At the beginning, a common language was established (with variables, and links between the variables, as its basic elements) and was used to develop mutual understanding shared by all participants. This graphic language enabled us to successfully join diverse participants’ experiences and backgrounds into a common model exploring regional sustainability issues. First, a list of potential variables was elicited. Secondly, the initial list was winnowed to narrow it down to a practical range (< 25) of key variables. Finally, we used causal loop diagramming [1, 14] as a discussion guide in linking variables and slowly developing a graphic image of the system structure. As the web of relations took shape, certain sections became more understandable as identification of reinforcing and balancing feedback loops reveals the system macrostructure. In this way, the vast and dense “thicket” of links was reduced to a smaller set of clusters of variables that tend to interact with each other. The group’s desire to focus on specific parts of the model often generated sub-model diagrams that clarified some of the causal details underlying the more aggregate variables and relations in the general model. Causal Loop Diagrams enabled us to elicit from the participants their underlying assumptions and mental models and to express them graphically in the form of a “map” containing key factors and processes in the region. The model functioned as the knowledge container; open and easily modifiable when new facts or ideas were provided or revealed during the process. In every discussion the model presented on the wall plainly showed the complex relations between nature and society in rural landscapes. This model proved to be easily understandable not only for the project participants but also for other local stakeholders and students.In summary using the model disciplined the group discussions in a positive sense as follows.Differences (and agreements) in opinions were articulated much more precisely.Gaps in understanding were discovered more efficiently.The conceptual model of regional sustainability issues (Figure 3) was developed following the overall goal of the NGO participants in this project to analyze and measure how their projects contribute to improve or sustain the quality of the regional environment. It should be stressed that this model, created during the first cycle of AMF, contains mostly the assumptions of the NGO professionals about the analyzed system. These assumptions should be challenged and refined during consecutive cycles. The model contains four main parts: Environmental Quality, Environmentally Friendly Farms, Green Local Products and Green Tourism.We assumed that Environmental Quality depends on the level (intensity and extent) of Environmentally Friendly Practices. In order to achieve the appropriate level of Environmentally Friendly Practices (EFP) we need to introduce Environmental Standards. In this top-level model the term “environmental standards” can represent both legally enforced regulations and/or standards voluntarily chosen by the producers to increase competitiveness. Success in establishing EFP depends on what Perceived Environmental Benefits are evident to the community in the region. All of these parts are heavily interconnected, and for each of them a detailed sub-model was developed. As an example, the Environmentally Friendly Farms sub-model is described in more detail in Appendix 1.The final version of the model was used for choosing key variables, which constituted the first step to choosing sustainability indicators.Based on our understanding of regional sustainability issues gained in developing the conceptual model, steps were taken to obtain the instruments for measuring progress – sustainability indicators. The regional sustainability model was the basis to identify the most important variables (marked in red color in Figure 4) and then derive sustainability indicators (see Table 1).Systems concepts and methods, such as the analysis of mental models and causal loop diagrams, are impressive for their power to clarify complexity but can be intimidating when first encountered. While the outcome is a useful simplification, along the way one must face and digest far more complexity than when filtering the world through the lens of a single discipline. Systems science rigorously engages and integrates multiple disciplines and experiences, and it takes years of training and application to master. However, with the help of experienced systems scientists such methods can be practically employed by both professionals and lay people to develop regional sustainability strategies and indicators. For this to happen, much attention has to be paid to clarifying the basics, including the specific language of systems analysis, and to providing the participants with an opportunity to use the knowledge they acquire during exercises. Also, causal loop diagrams have to be discussed gradually, starting from a single loop and building the whole diagram around it.Systems methods help people see what they normally do not consciously think about or discuss in an open forum: feedback loops with complex interactions and delays that create long and medium-term impacts. The transition to sustainability requires that stakeholders grasp the structure of systems. These methods help in that transition by exposing counter-intuitive links between natural, economic and social processes and by showing how delays distort our understanding of change. Stakeholders in the Barycz Valley approached by the NGOs directly participating in the project, as well as the NGO representatives, found these methods and ideas powerful in opening new ways to capture values, qualities and relationships related to the sustainability of their community and environment. They also felt that systems science methods were good complements to traditional verbal descriptions of sustainability. Briefly, stakeholder-driven processes that use graphical maps of sustainability engage a far wider group and build broader acceptance of novel ideas than do words alone.Professionals in local NGOs are good contributors to the start of a discussion within the Adaptive Management Framework, helping initially to absorb new ideas and methods, and subsequently to act as bridges of understanding to local stakeholders with whom they have established trust over the years. Once an Adaptive Management Framework has established an open, trusting exploration of ideas among such professionals, the discussion can much more easily be extended to include (and be challenged to improve by) the experience of the wider community of concerned citizens.Conceptual modeling involves parallel efforts to examine regional as well as local sets of processes associated with individual variables and/or feedback loops. These processes complement one another in improving the view at one scale by adding perspectives from other scales. The process of starting at a regional scale and then focusing down to specific questions led to a much-improved regional model applicable to a range of sustainability questions.The essential point of the AMF is to use the insight gained in assessment to act, then monitor the impacts of action and then start the cycle again by using the results for reassessment. Collaborative re-assessment of policies and practices will require closer cooperation with local authorities, professionals, NGOs and concerned citizens in future projects, because such processes cannot be forcibly applied by administrative rulings. They must be understood and accepted by a majority of stakeholders such that an AMF process is voluntarily engaged by society. The engagement of NGO professionals is crucial for success because their work has established the trust that will bring local stakeholders into such a discussion and help them fully contribute to the discussion, to monitoring impacts and to subsequent reassessments and new actions. University education cannot teach how AMF should be applied and cannot create the basis of experience by which to advance its concepts and methods. Only through application in the real world can broad coalitions of lay people and professionals successfully learn how to use AMF and the project described in this article provides a good example of how this can be carried out. That wide stream of experience will provide the foundations to devise new concepts and methods through which AMF dialogues can evolve.Adaptive management process as a structured learning cycle that iteratively links four phases: starting from assessment, through policy formulation to implementation, and monitoring used as an input for the assessment phase in the next cycle.Adaptive Management Framework used in the Barycz Valley. Conceptual Modeling was incorporated into Assessment phase (steps 1 and 2).Regional sustainability model with four main parts aggregated.Variables and causal links in the Environmentally Friendly Farms sub-model. Variables in red are the key variables, which were used to find sustainability indicators.Key variables and indicators that describe themThe research was realized within the project funded by DBU (Deutsche Bundesstiftung Umwelt) and WWF Germany: “Sustainable Regional Development in the Odra Catchment” (Grant Number: DBU AZ 18902). Collaborative research in the field requires generous commitments of time, experience and good will over long periods. In particular, we wish to thank Peter Torkler (WWF Germany), Dorota Chmielowiec (Lower Silesian Foundation for Sustainable Development) and Andrzej Ruszlewicz (Green Action) for sharing their rich knowledge of the region and its people. Their insightful comments and questions are at the heart of the understanding built by this project within the region.Regional Sustainability Model was developed using Causal Loop Diagrams. This qualitative system dynamics approach emphasizes the system structure described in terms of balancing and reinforcing feedback loops. Interactions between loops and shifting dominance provide the explanation for the complex system behavior. Here we describe one of the diagrams, Figure 4, which focuses on Environmentally Friendly Farms; however interactions with other sections of the model are also present.When Environmental Quality deteriorates Social Support for Environmental Standards increases as people, seeing the poor state of the environment, are more eager to support some standards to improve it. This support enables the introduction of better Environmental Standards, which in turn raises the level of Environmentally Friendly Practices. After some time (delay) it leads to improvement of environmental quality. These relations create the balancing loop (B2 - Rescue Environment) which operates to keep Environmental Quality in a good state.So far we have assumed that better Environmental Standards increase the level of Environmentally Friendly Practices. However, this is true only if we keep the number and area of Environmentally Friendly Farms unchanged. The immediate result of introducing new Environmental Standards may be the decrease of Profits from Environmentally Friendly Crops, which may discourage some farmers and decrease the number or area of Environmentally Friendly Farms. This shows that introducing new Environmental Standards can affect the level of Environmentally Friendly Practices in two different ways and one cannot be certain whether this level will raise or fall. In effect this may cause the loop B2 (Rescue Environment) not to operate in a desired way.When Profits from Environmentally Friendly Crops go down Perceived Environmental Benefits (any benefits coming from environment perceived by community in the region) will also go down. This will lower Social Support for Environmental Standards and in turn will make more difficult to keep Environmental Standards which is our main tool for improving Environmental Quality. This describes another balancing loop (B1 - Environmental Standards Raise Costs and Lower Crops, as a consequence of lower productivity), which operates to keep Environmental Standards at low level. This loop describes the resistance, which environmental NGOs quite often encounter. The source of this resistance is the mental model, which states that good environment means poor economy. The further loops in our model show why this does not always have to be true, however in the short term it quite often happens this way.The above analysis shows that in order to start the process of improving Environmental Quality through increasing the number of Environmentally Friendly Farms the impulse from outside is needed. Such an impulse can be provided through Organizational Support for Environmentally Friendly Farms. This in practice is mostly done by environmental NGOs but it can also be done by local or regional authorities or other institutions. Usually these stakeholders decide on the target – Desired Level of Environmentally Friendly Practices and then execute the pressure until a gap between the actual level of Environmentally Friendly Practices and Desired Level of Environmentally Friendly Practices disappears. This process creates another balancing loop (B3 - Organizational Pressure for Environmentally Friendly Practices), which is extremely important for achieving environmental goals. It is aimed to make the system operate without external support, which means that ultimately environmentally friendly farms should be economically self-sufficient. But in order to achieve it, there must be a certain period when sufficient support is provided to those farms. It should be emphasized that institutional support for creating environmentally friendly farms, as well as producing and promoting green local products and developing green tourism is one of the crucial factors of success in this process. Recent research on innovation implementation [15] shows that one of the main reasons for collapses of improvement initiatives is cutting the external support too early. The balancing loop B3 must operate for a sufficiently long time to enable the reinforcing loops, which amplify innovation, to operate in the right direction. The role of environmental NGOs (or other institutions) is to keep this process operating until critical thresholds are reached and reinforcing loops can amplify both economic and environmental goals at the same time.The first reinforcing loop (R1 - Revenues through Agri-Environmental Programs) connects Environmentally Friendly Practices with Profits from Environmentally Friendly Crops. The Environmentally Friendly Farms become more profitable, which encourages other farmers to increase Environmental Standards and generates more Environmentally Friendly Practices which finally makes possible to obtain even more Revenues from Agri-Environmental Programs. It should also be noted that increasing Profits from Environmentally Friendly Crops makes it easier for people in the region to perceive environmental benefits, and raises Social Support for Environmental Standards. We have seen that short-term drop in Profits from Environmentally Friendly Crops is corrected through longer-term Revenues from Agri-Environmental Programs.The other long-term process is connected with “green tourism” opportunities, which are only possible, when Environmental Quality is sufficiently good. Improving Environmental Quality influences Touristic Attractiveness of the region but after significant time delay. Raised Touristic Attractiveness makes possible to obtain additional Profits from Green Tourism, which raises Perceived Environmental Benefits. This process closes another reinforcing loop (R2 - Nature Attracts Tourists). It affects also the balancing loop B1 (Environmental Standards Raise Costs and Lower Crops) making easier to keep Environmental Standards.Environmental Standards not only define the standards for farmers but also can be used to introduce the local brand for “Green Local Products” (GLP). If GLP Production brings profits the part of it can be reinvested to increase or diversify the production bringing even more Profits from Green Local Products. These links create the reinforcing loop R6 (GLP Growth through Reinvestment). Profits from Green Local Products raise Perceived Environmental Benefits so it also contributes to better environmental standards.The relation between production and profits from green local products is obviously influenced by many factors such as: Attractiveness of Green Local Products to Consumers, external Support for Green Local Products, Regional Food Processing Capacity and Local Cultural Identity. Brand Attractiveness constitutes another important factor as the successful introduction of a brand may greatly help in marketing and sales of Green Local Products.Environmentally Friendly Farms and Green Local Products are tightly connected. GLP Production is mainly based on crops from Environmentally Friendly Farms. Increased demand on crops from environmentally friendly farms leads to bigger Profits from Environmentally Friendly Crops. This makes being an “environmental farmer” more attractive and leads to the growth of Environmentally Friendly Farms. This process creates another reinforcing loop R5 (Environmentally Friendly Farms and Green Local Products Reinforce Each Other).Green Local Products are also connected with tourism. Many Green Local Products will be sold through green tourism facilities. In this way Touristic Attractiveness affects Profits from Green Local Products. This link closes another reinforcing loop (R3 - Revenues from Local Products Sales to Tourists). Touristic Attractiveness also improves the Region Image, which makes the local brand much more recognized and attractive. These links close reinforcing loop R4 (Revenues from Local Products through Green Image).To sum up, the balancing loop B1 operates in the short term hampering the introduction of Environmental Standards aimed to improve Environmental Quality. In the long term, reinforcing loops R1 to R6 make environmental benefits much more obvious for the community in the region. This means that introducing Environmental Standards is the process which requires patience – it is necessary to wait to overcome the initial negative economic effects. The model helps to understand that “environment or economy” is a false dichotomy if we look at the situation with sufficiently long time horizon.For a historical review of community indicators, see: GahinR.PatersonC.National Civic Review2001904347361A lot of information about sustainability indicators is available in the internet and the good selection of websites was provided by: HechtJ. E.Environment200345134
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+ Excessive amounts of arsenic (As) in the groundwater in Bangladesh and neighboring states in India are a major public health problem. About 30% of the private wells in Bangladesh exhibit high concentrations of arsenic. Over half the country, 269 out of 464 administrative units, is affected. Similar problems exist in many other parts of the world, including the Unites States. This paper presents an assessment of the health hazards caused by arsenic contamination in the drinking water in Bangladesh. Four competing hypotheses, each addressing the sources, reaction mechanisms, pathways, and sinks of arsenic in groundwater, were analyzed in the context of the geologic history and land-use practices in the Bengal Basin. None of the hypotheses alone can explain the observed variability in arsenic concentration in time and space; each appears to have some validity on a local scale. Thus, it is likely that several bio-geochemical processes are active among the region’s various geologic environments, and that each contributes to the mobilization and release of arsenic. Additional research efforts will be needed to understand the relationships between underlying biogeochemical factors and the mechanisms for arsenic release in various geologic settings.Arsenic is a natural component of the Earth’s crust. It can be found in soil and water that have interacted with arsenic-rich rocks. Arsenic can also be introduced into the groundwater anthropogenically through the application of the arsenic-rich herbicides and pesticides that are frequently used on agricultural lands.The maximum acceptable levels of dissolved arsenic in drinking water is 0.01 mg/l and 0.05 mg/l according to the World Health Organization and the United States Environmental Protection Agency, respectively [1, 2]. Acceptable levels of dissolved arsenic in drinking water in Bangladesh are 0.05 mg/l. Drinking water that has elevated levels of arsenic for a prolonged time period is unsafe, and specific health risks are well documented [3, 4].Arsenic-contaminated groundwater was first documented in groundwater in 1984 [4]. Since then, several systematic studies have been carried out to determine the extent and severity of this contamination [5, 6]. The magnitude of the problem is not yet known, however, the number of affected tube wells and arsenic-related health complaints is on the rise. About 30% of 10 million private wells are highly contaminated. Most people in Bangladesh use groundwater water as a source of drinking water because the surface water has unacceptable level of bacterial contamination. A total of 269 out of Bangladesh’s 464 administrative units, called upazilla, more than half the country, contain high levels of arsenic-contaminated drinking water (Fig. 1).In addition to Bangladesh, several countries in the world have identified excess amount arsenic in drinking water including Argentina, Bangladesh, Chile, China, Hungary, India, Japan, Mexico, Mongolia, Poland, Taiwan, and the United States. However, it is noteworthy that the number of people affected by arsenic pollution in Bangladesh exceeds the total number of people affected by arsenic pollution in all other countries combined.Most of Bangladesh and West Bengal, India, fall within the Bengal Basin, one of the world’s largest geosynclinal troughs (a large elongate or basin-like structure on Earth’s surface). The Ganges-Brahmaputra-Meghna river system occupies this extensive basin. As it nears the Bay of Bengal coast, this system spreads out forming many distibutary channels. Along the coast, the mean tidal range is high, and wave energy is low. As a result, the Ganges-Brahmaputra deltaic complex serves as a textbook model of a tide-dominated delta.The Bengal Basin comprises a wide variety of fluvial (e.g., channels, floodplains, natural levees, back swamps) and marginal marine (e.g., distributary channels, mangrove forests, tidal flats, beaches) sedimentary environments. Distinctive sediment, biological activity, and chemical and physical processes characterize each of these environments. They inter finger laterally as well as vertically and continuously change their position in response to tectonic disturbances as well as global variations in climate and sea-level fluctuations.Although sea level has varied considerably throughout geologic time, the changes over the last 1.6 million years (Pleistocene and Holocene epochs) have been dramatic. During this time, the sedimentary environments of the Bengal Basin shifted both seaward and landward relative to the present coastline in response to global sea-level changes. It is likely that the southern boundaries of the Pleistocene highlands, terraces such as the Barind Tract, Madhupur Tract, and Tippera Terrace, marked the position of the paleoshoreline (shoreline position at a given time in the past) during the last interglacial period (125,000 years ago).As sea level fell to its lowest stand during the last glacial maximum (~18,000 years ago), the shoreline gradually advanced ocean-ward (to the south). Since then, the shoreline migrated landward as sea level rose. Sea level in the Bay of Bengal was about 7 m below its present position about 7000 years ago [9]. As a consequence, the various depositional environments migrated back and forth between the Pleistocene highlands and the position of the present shoreline, leaving behind their characteristic sedimentary features, now buried beneath the recent sedimentary cover. For example, the gradient of rivers must have been higher during the low sea stand, facilitating formation of incised river valleys. Lower parts of these incised river valleys must have been filled with fine-grained and organic-rich sediments during subsequent sea-level rise. It is important to analyze the spatial relationship between these valley fills and high concentration of arsenic in underlying aquifers.Arsenic concentration data throughout the basin are highly variable on both spatial (map view; refer to Figure 2) and temporal (at depth in cross sections and drill holes) scales. The reason for this variability is not well-understood. The source of arsenic in this region is dependent on the geology of the upstream region and land-use practices in the catchment’s areas of the major rivers that carry and distribute the sediment and anthropogenic waste materials.However, the geological, hydro geological and bio-geochemical processes operating during sediment deposition played a significant role in controlling the mobility of arsenic. Dissolved arsenic exists as several chemical species, the stability of which depends on the oxidation-reduction potential of the environment, as well as the microbial activity present at various depths in the sedimentary deposits [10]. The amount of dissolved organic content and dissolved ions likely to play a role in arsenic mobility. For example, arsenic-bearing iron can form chemical complexes with other anions, which, in turn, can enhance the solubility and mobility of arsenic species. The degree of complexation of arsenic-bearing iron cations is likely to be a higher in groundwater that has high concentration of dissolved organics and other anions, such as the groundwater found at mid-depth aquifers in Bengal Basin [5–7]. In order to better understand the complexity of the origin, pathways, and sinks of dissolved arsenic in groundwater, a detailed analysis of subsurface geologic history and bio-geochemical processes is necessary. Constructing paleogeographic (distribution of land and water at a given time in the past) maps of the Bengal Basin based on the depositional environments at depth can aid in the interpretation of the data and suggest mechanisms responsible for the mobility of the dissolved arsenic.Figure 3 shows the regional tectonic map (portrayal of large geologic structures on Earth’s surface). The distribution of high arsenic concentrations appears to have a strong correlation with certain tectonic elements in Bangladesh. For instance, over 75% of tubewells contaminated with high arsenic concentrations are found in Bengal Foredeep (low-lying elongated depression) area, which consists of Faridpur Trough, Barisal Gravity High, Hatiya Trough, and Sylhet Trough.Four hypotheses are proposed to account for the origin and mechanism of the arsenic pollution (i.e. an amount sufficient to cause health problems) of groundwater in Bangladesh. The proposed mechanisms are as follows:Oxidation of arsenical pyrite;Reductive dissolution of FeOOH (hydrous ferric oxides or HFO) resulting in the release of sediment-bound arsenic;Anion (competitive) exchange of sediment-bound arsenic with phosphate from fertilizers;Release of arsenic from the degradation of pesticides* and fertilizers. We used published data as well as the knowledge of socio-environmental and geologic conditions in Bangladesh in critically reviewing these mechanisms to explain arsenic release to groundwater.According to this hypothesis, the arsenic pollution results from oxidation of authigenic pyrite that is concentrated in deposits of organic matter due to lowering of the water table. This hypothesis implies that arsenical pyrite oxidizes in the vadose zone (drawdown zone around pumping wells) releasing arsenic attached on iron hydroxides [11, 12, 13]. The proponents of this hypothesis also suggest that increased drawdown during the dry season and the subsequent recharge of groundwater facilitates this oxidation process. They argue that the arsenic pollution is a recent phenomenon, which is triggered by the diversion of surface water in upstream regions of the Ganges River. The following chemical reaction is proposed as a mechanism for the release of arsenic from arsenical pyrite [10]:Questions addressing this mechanism include:Why is arsenic found in areas that are not affected by upstream diversion of surface water, such as in the Meghna basin (NE Bangladesh)?Why are arsenic concentrations not highest in the unsaturated zone beneath the surface where the groundwater table fluctuates between dry and wet seasons?Does arsenic pollution increase during the wet season?Are the Eh-pH field changes in the aquifer due to drawdown?Why is the arsenic pollution so severe in West Bengal and Bihar where upstream diversion of surface water did not occur?Why does the arsenic pollution occur in areas where drawdown is not a likely cause for oxidation, such as in Wisconsin, USA?According to this hypothesis, arsenic contamination results from microbial reduction of organic-rich, fine-grained sediments that contain arsenic-coated hydrous ferric oxides (HFO) which were deposited in low-lying floodplains and coastal plains during the last several thousand years [6, 14–19]. This hypothesis implies that arsenic is released from HFO surface when iron-reducing bacteria oxidize organic carbon for their metabolism. The following chemical reaction is proposed as a mechanism for arsenic release from HFO [6]:If this is true, then arsenic would be found in groundwater that passes through organic rich sediments. However, arsenic concentration, though most common in shallow aquifers at depths ranging from 15 to 75 meters (m) in low-lying (i.e. elevation less than 10 m) coastal plain regions (i.e. Faridpur Trough, Hatiya Trough, and Barisal Gravity High in Fig. 3), is not exclusive to this geologic setting and depth range. For instance, high concentrations of arsenic are also found in Sylhet Trough, Bangladesh (Fig. 3), Bihar and Chattishgarh, India, and Nepal, which do not represent low-lying floodplain or coastal plain settings. Moreover, the sea level has risen about 7 m over the last 7,000 years [9], however the high concentration of arsenic is not limited to aquifers that were deposited during last several thousand years. Aquifers that are deeper than 100 m and are separated from the shallow aquifers also contain high concentrations of arsenic.Additionally, the following questions relative to this hypothesis demand answers:Is there a peat layer above the aquifer in the entire area that is polluted with arsenic?Even if peat layers and/or organic rich sediments occur in aquifer materials, at what depth do they occur?Why is there arsenic pollution along the Brahmaputra riverbanks, where sediments are predominantly sandy in nature?The peat layer associated with the low sea level position in the Bengal Basin is found at depths ranging from 2 to 7 m below mean sea level and about 80 to 120 km landward of the present shoreline location. Why then is the aquifer located at depths ranging from 15 to 75 m polluted with arsenic? How does the arsenic produced in the peat layer located at 2 to 7 m move to a greater depth?If organic-rich sediments are deposited in tidal flat or coastal plain environments, why are there no high concentrations of arsenic in most of today’s coastal plain, such as in Bhola, Barguna, and Patuakhali (south central parts in Figures 2 and 3)The Madhupur Tract and Barind Tract are also very similar in their geologic origin, i.e. they represent ancient delta plain and floodplains. Why are there no organic-rich sediments and high concentrations of arsenic in the aquifers in those areas?Can the hypothesis account for the variability in arsenic concentrations in terms of time-dependent field data? [20].Depending on the redox potential and acidity of the environment, arsenic can exist in several anionic forms. Arsenic is attached on surfaces of fine-grained clay and hydrous ferric oxides. When other anions, such as phosphates (from fertilizers and other sources), exist in excess, they can replace arsenic anions and release the arsenic to groundwater. Phosphate anions are relatively immobile and get attached to mineral grains near the surface. Accordingly, the “overuse” of phosphate fertilizers during the last few decades must have played a role in dislodging arsenic attached to mineral grains and introducing it into groundwater [21].The competitive ion-exchange mechanism, however, would not be very effective in deeper aquifers. Arsenic concentrations appear to be highest at depths ranging from 15 to 75 m. These are not explained by this hypothesis. Additionally, fertilizer usage is very common in all parts of Bangladesh. It follows that if phosphate anions from fertilizers were responsible for the release of arsenic to the groundwater, then arsenic would be present in groundwater throughout Bangladesh; however, arsenic concentrations do not follow this pattern (Figures 2 and 3). Also, over 75% of the high concentrations of arsenic are found in fine-grained, organic rich (marsh clay and peat) sediments (alluvial/deltaic silt and clay) in tidal and deltaic environments, a fact that cannot be explained by this hypothesis.Relative to this hypothesis, the questions that beg answers are as follows:Why do not the phosphate anions from fertilizers force the release of arsenic in all geologic settings equally?Why are arsenic concentrations not high in surface water where phosphate anions are supposed to be in high concentration?Is there any relationship between fertilizer application and elevated arsenic concentrations in Bangladesh or elsewhere?Apparently, some of the chemical fertilizers and pesticides used in Bangladesh contain high amounts of arsenic, which might have been introduced into groundwater [22, 23]. According to this study [23], there is no evidence that arsenic originated from any natural source in Bangladesh. It is also claimed that a large amount of arsenic trioxide is kept in stock at Ghorasal Fertilizer Factory in Bangladesh for use or disposal to unknown destinations. About 30% of the fertilizers used in Bangladesh are lost through surface run-off. It is also suggested that in some parts of Bangladesh there is an apparent correlation between the amount of fertilizers used in crops and the amount of arsenic present in groundwater [23].It has not been demonstrated that fertilizers and pesticides have been applied in sufficient quantities to degrade groundwater quality in Bangladesh [24]. Other questions relative to this hypothesis that need to be answered are as follows:What are the pathways and sinks of pesticides and fertilizers as they relate to groundwater movement?How do the pesticides and phosphate fertilizers move through the fine grained sediments into the deep aquifers?Why is the amount of arsenic in groundwater not proportional to the amount of pesticides and fertilizers used Bangladesh?Are there any seasonal variations in the amount of arsenic observed in groundwater that reflects the usage of pesticides and fertilizers?Why are the high concentrations of arsenic predominantly associated with the low-lying coastal/deltaic environments that are characterized by fine-grained sediments (Fig. 2)?The lack of understanding of the origin, pathways, and sinks for arsenic contamination in Bangladesh’s groundwater is taking a terrible toll on human health. The arsenic contamination has reached an epidemic proportion. Millions of people are suffering from various arsenic-related diseases, and millions more are exposed to the possibility of contacting these diseases. Arsenic is also finding its way into the food chain. In order to better understand the affect that arsenic has on human health and the state of the epidemic in Bangladesh, it is imperative that the link between the food and arsenic uptake by people in Bangladesh be examined thoroughly before any remedial actions can be taken to mitigate this problem. To illustrate the nature and severity of arsenic-related diseases, in the next section entitled “Effect of Arsenic Poisoning on Health”, we have included the results of an investigation that one of the authors (Mitra) carried out with other researchers [36].Chronic arsenic poisoning, arsenicosis, can increase the risk of several health hazards including skin lesions, cancers, restrictive pulmonary disease, peripheral vascular disease (blackfoot disease), gangrene, hypertension, non-cirrhotic portal fibrosis, ischemic heart disease, and diabetes mellitus [25–35]. A study of 150 patients (75 males and females) visiting the dermatology outpatient department of the Sher-e-Bangla Medical College Hospital, Barisal district, Bangladesh, in 2000, provides substantial evidence for this link. In addition, results confirm the connection between arsenic toxicity and malnutrition.Skin changes due to arsenic poisoning include a raindrop pattern of pigmentation and depigmentation that is particularly pronounced on the extremities and the trunk. Although less common, other patterns include diffuse hyperpigmentation (melanosis) and localized or patchy pigmentation, particularly on skin folds. Hyperkeratosis (hardened skin) appears predominantly on the palms and the planter surface of the feet. In the early stages, the involved skin might have an indurated, gritlike character that can be best appreciated by palpation; however, the lesions usually advance to form raised, punctuated, 2–4 mm wart-like keratosis that are readily visible. Occasional lesions might be larger (0.5 to 1 cm) and have a nodular or horny appearance occurring in the palm or dorsum of the feet. In severe cases, the hands and soles display diffuse verrucous lesions [26]. All the patients of our study displayed raindrop skin pigmentation, and more than 80% had hyperkeratosis with or without nodular skin lesions (Fig. 4). One hundred and twenty-three (82%) patients had moderate or severe skin lesions. Sites of the skin lesions were trunk, including chest (38%), hands only (18%), both hands and feet (15%), feet only (13%), and chest only (11%).Skin cancer resulting from chronic arsenicosis is quite distinctive. Multiple lesions are common and involve covered areas of the body, contrary to non-arsenical skin cancers which usually appear as a single lesion and which occur in exposed parts of the body. In our study, biopsy specimens from skin nodules of one person, aged 42 years, showed squamous cell carcinoma (Fig. 5). Other types of cancers reported in significantly higher number among patients with chronic arsenic poisoning include cancers of lung, urinary bladder, kidney, prostate, and liver.About 18% of our study patients did not complain of any clinical symptoms, except that their skin lesions were ugly-looking. One hundred and fifteen (77%) subjects had multiple symptoms, including weakness, chronic cough, joint pain, itching, abdominal pain, chest pain, loss of appetite, insomnia, shortness of breath, and frequent urination with burning (Table 2).High blood pressure (systolic BP ≥ 140 mm Hg or diastolic BP ≥ 90 mm Hg) and depression (lowered mood or sadness, loss of interest, and anxiety) were the two most frequent complications, each observed in 20 (13%) of our study subjects. The other important findings included raised serum alanine aminotransferase (n = 13), palpable liver (n = 8), X-ray features suggestive of pneumonia, interstitial lung disease and lung abscess (n = 11), peripheral vascular problems in the form of intermittent claudication, in the absence of history of smoking (n = 2), pulmonary tuberculosis (n = 2), diabetes (n = 1), and decreased libido (n = 1). Anemia (hemoglobin: <135 g/L in males, and <120 g/L in females) was a major clinical finding observed in 88 (58%) subjects; 14 had pedal oedema. Ten (7%) patients were admitted to the hospital due to complications or associated major illnesses, including severe anemia, hepatitis, hepatic cirrhosis, renal failure, lung abscess, interstitial lung disease, intermittent claudication with uncontrolled diabetes, and skin cancer. One of them died due to heart failure from chronic obstructive pulmonary disease and lung abscess.Several clinical symptoms that appeared in our subjects had conformity with previous reports [23–35, 37, 18]. A dose-effect relationship has been reported between arsenicosis and patients with skin cancer, blackfoot disease, cardiovascular disease, hypertension, and diabetes [25, 33, 34, 39]. We found a significant direct relationship between the mean arsenic concentration in water and the severity of the clinical disease.Eighty-nine percent of our patients were underweight and 11% were weight-appropriate; none were overweight or obese. After controlling for age, the duration of disease varied inversely with BMI (r = −0.298, P = 0.013, n =70). Their body weight, height, and BMI did not differ by the severity of disease. Evidence suggests that poor nutritional status may increase toxicity to arsenic retained in the body, probably by diminished ability to methylate inorganic arsenic [40]. Patients with protein-energy malnutrition are particularly deficient in methionine. Studies by Vahter and Marafante found that a low amount of methionine or protein in the diet decreased methylation of inorganic arsenic in rabbits [41]. Deficiency of certain other dietary trace elements, including zinc and selenium, also associated with malnutrition, may contribute to the toxic effects of accumulated levels of arsenic in the body [42]. Our study confirmed previous reports of the relationship between malnutrition and increased arsenic toxicity [43]).Generally, in unpolluted environments, ordinary crops do not accumulate enough arsenic to be toxic to man. However, in arsenic contaminated soil, the uptake of arsenic by the plant tissue is significantly elevated, particularly in vegetables and edible crops [44]. There is, therefore, concern regarding accumulation of arsenic in agricultural crops, vegetables, and fishes grown in the arsenic-affected areas of Bangladesh.In a recent study of 100 samples of crop, vegetables and fresh water fish collected from three different regions in Bangladesh. Arsenic concentrations were not increased in samples of rice grain (Oryza sativa L.) [45]. The results of the study were consistent with a previous study in Bangladesh [46]. However, rice plants, especially the roots had a significantly higher concentration of arsenic (2.4 mg/kg) compared to stem (0.73 mg/kg) and rice grains (0.14 mg/kg). While not covered by food hygiene regulations, rice straw is used as cattle feed in many countries including Bangladesh. The high arsenic concentrations in straw may have the potential for adverse health effects on the cattle and an increase of arsenic exposure in humans via the plant-animal-human pathway. Further studies are needed to measure arsenic concentration in meat from cattle fed on contaminated rice straw.In the study by Das et al. [45] it was found that arsenic contents of vegetables varied; those exceeding the food safety limits included Kachu sak (Colocasia antiquorum) (0.09–3.99 mg/kg, n = 9), potatoes (Solanum tuberisum) (0.07–1.36 mg/kg, n = 5), and Kalmi sak (Ipomoea reptoms) (0.1–1.53 mg/kg, n = 6). Lata fish (Ophicephalus punctatus) (n = 9) did not contain unacceptable levels of arsenic. These results indicate that arsenic contaminates some food items in Bangladesh. Further studies with larger samples are needed to demonstrate the extent of arsenic contamination of food in Bangladesh.In addition to carrying out research to fully understand the health implications of arsenic contamination of groundwater and food sources, it is necessary to determine the sources, chemical behaviors, hydrogeologic control, and pathways of arsenic in groundwater in Bangladesh. As discussed above, there is no clear understanding of the geology and its relationships to arsenic contamination in Bangladesh. In the next section entitled “Future Research and Policy Implications” we have offered specific suggestions for policy makers and researchers who are involved in arsenic mitigation and remediation projects.Based on critical review of the four competing hypotheses, we conclude that none of these hypotheses can solely explain the nature of variability of the arsenic concentration found in groundwater in Bangladesh. Additional research will be needed before any prediction can be made as to the occurrence, mobility, and concentration of arsenic in aquifers in different geologic settings. To this end, new directions for future research are suggested by the authors involving:A comprehensive study of the geologic history including the compilation of paleogeographic interpretations for various time intervals.The identification of the biogeochemical processes active in different geographic and geologic settings.The investigation of regional land-use practices.The design of appropriate conceptual models for determining groundwater flowpaths.The design of appropriate geochemical models of reaction mechanisms.A comprehensive study of the epidemiological aspects of the problem. These studies need to be undertaken in order to establish relationships between biogeochemical factors responsible for the release of arsenic and the observed concentrations of arsenic in groundwater in Bangladesh.Our research analyzed four existing hypotheses addressing the arsenic problem in Bangladesh. We found that none of the hypotheses listed above solely can explain the nature of variability of the reported arsenic concentrations. Further studies are needed in the following areas:Detailed lithologic descriptions of well logs (types and characteristics of units) to establish a 3-D framework of the geologic strata to identify and characterize those units acting as aquifers (porous and permeable layers that hold and transmit water to tubewells) aquitards (impervious layers that impede water movement through them) in the arsenic-affected areas.The geologic history of evolution of the Bengal Basin and the resultant sediments as a function of time and space.Reconstruct paleogeographic maps of Bangladesh, which will illustrate the types of sediments deposited at various locations and depths at a given time period.Develop conceptual models of groundwater flowpaths, with evidence supported by computer simulations.Use geochemical data collected from the existing pumping wells, as well as from newly drilled wells, to develop geochemical models, which can be used to determine the types of chemical reactions and products, including possible chemical complexes. The best way to determine complexes is to utilize geochemical computer codes, such as MINTEQ or PHREEQE [47].Compile geologic information for sediments and chemicals observed at various depths to decipher spatial and temporal relationships between sediments and arsenic concentrations.Multidisciplinary studies to better understand biogeochemical environments and epidemiological relationships to the natural environments.To combat the health problems due to arsenic poisoning, it is necessary to:Enhance public awareness of the health problems from contaminated water.Take short- and long-term intervention strategies to curb the exposure.Strengthen rapid diagnostic facilities.Establish effective treatment facilities in rural areas.Improve the nutritional status of the people.Arsenic concentrations in groundwater in Bangladesh and West Bengal, India [4].Geologic units and arsenic concentrations in groundwater in Bangladesh [7, 8].Map showing the distribution of arsenic in the context of the generalized tectonic map of Bangladesh [7, 8].A 42-year old woman having characteristic raindrop skin lesions on her both extremities and a nodule on her thigh [36].Biopsy specimen from the nodule showed anaplastic squamous epithelial cells in interconnected sheets and islands in the dermis. Well-defined keratin horn pearls are also demonstrated (Hematoxylin-Eosin Stain, x122) [36].Summary of four competing hypotheses regarding the mechanism for arsenic occurrence and mobilization in groundwater in Bangladesh.-Requires pyrite as source of As-Requires oxygen associated with lowering of groundwater table-Known occurrence of As do not always correspond with oxidizing environments in aquifers-Presence of pyrite in all polluted aquifers is not documented-Presence of As in areas not affected by groundwater withdrawal (e.g. Nepal, Bihar and Chattishgarh, India)-As is also present in geologic settings that are different from those in Bangladesh-Not supported by data collected from aquifers at different depths and geologic environments-Not demonstrated to work with hydro geochemical modelling at large scales-Pyrite is not stable in the known oxidation-reduction-potentials of aquifers in Bangladesh-Results from microbial reduction-Requires presence of organic-rich sediments in aquifers-Requires reducing environment in aquifers containing HFO-As release mechanism must have been present over geologic time-Known occurrence of As does not always correspond to known presence of organic-rich layers-Oxidation-reduction potential of aquifer is not always reducing where As has been reported-As also present in geologic settings other than coastal/delta plain (e.g. in northern areas of Bangladesh, India, and Nepal)-Not supported by data collected from aquifers at different depths and geologic environments-As contamination appears to be a relatively recent phenomenon-Not demonstrated to work with hydrogeochemical modelling at large scales-Requires high concentrations of phosphates in groundwater-Amount of phosphate anions available from fertilizers do not correspond with the calculated amount of As present in groundwater-High concentrations of phosphates are not demonstrated where As is found-No relationships between fertilizer applications and As occurrence has been established-Phosphate is immobile and is not common in groundwater (15–75 m) where As is found-Amount of reported As in groundwater cannot be accounted for by the amount of potential arsenic released from cumulative use of fertilizers and pesticides-Pesticide and fertilizers are applied most everywhere in Bangladesh, but the high As concentration does not match the pesticides/fertilizer application patternsRanking of clinical symptoms of arsenic-related illnesses in 150 patients admitted to the Sher-e-Bangla Medical College Hospital, Barisal, Bangladesh, January-December 2000.Total percentage exceeds 100 because several patients had multiple symptoms and complications [36].
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+ Two approaches are distinguished in modern ecological monitoring. The first one is physicochemical analysis of environmental objects with respect to maximum allowable concentrations (MACs) of chemical substances, which is performed by standards methods in accordance with state regulations. The second approach (biological monitoring) is based on the methodology of biotesting and bio indication. The task of this work is to create biotests for estimation of Al and other metals toxicity in ground water and to compare these results with physicochemical analysis dates. Risk assessment for heavy metals contaminated groundwater was also performed. Risk assessment was performed accordingly EPA US recommendation and gave results about 90 per 100000 citizens for Al and 402 per 100000 for mixture of different heavy metals. For comparison: risk for earth background radiation for Middle Russia is (Individual dose 1 millisivert per year) consist 5 per 100000 people. It was shown that groundwater consist HCO3− ions (360 mg/l), sometimes Al compounds 0.21–0.65 mg/l (MAC for Al is 0.5 mg/l for Russia). Other groundwater contain Hg – 0.004 mg/l (MAC – 0.0005 mg/l); Cr – 0.072 mg/l (MAC – 0.05 mg/l); As – less than 0.03 mg/l (MAC – 0.05 mg/l). We developed plant biotest for estimation of groundwater quality with barley roots, tradescatia and others. Some biotests parameters correlate with HCO3−, Cl−, SO42− and metal ions content positively, for another biotest this correlation is strongly negative. The quality of groundwater near Obninsk and in Kaluga Region is very different but hasn’t been changed since the year 1998.Now the traditional ecological monitoring involves not only physical and chemical analysis of environmental objects connected with a system of maximum permissible concentrations (MPC) and classes of danger of chemical substances, maximum permissible doses and levels (MPD and MPL) of ionizing radiation and electromagnetic fields, but also risk assessment which allows qualitative and sometimes quantitative determination of the impact of anthropogenic factors or their combinations to be performed [1].At present health risk assessment, i.e. assessment of expected consequences as a result of exposure to any hazardous substance or dangerous effect is one of the essential medical-ecological problems.Contaminated drinking water is one of the expected sources of human health risk. Intensive operation of water intakes and anthropogenic contamination of groundwater are considered to be the reason for this kind of risk. This caused disorder in the natural regime of groundwater and resulted in deterioration of its quality.The situation observed in Obninsk and other water intakes of Kaluga region may serve an example of these changes. The MPC excess even in one water sample was established from the following parameters: mineral and microcomponent composition (nitrates, nitrites, ammonium, iron, strontium, manganese, fluorine, barium, cadmium, copper), content of toxic substances (benz(a)pyrene, oil- products), radioactive isotopes (tritium, strontium-90) as well as organoleptic (turbidity, color) and bacteriological properties [2].Increased concentrations of the above elements and compounds in drinking water harmfully affect human health and cause toxic, mutagenic or carcinogenic effects depending on the substance properties. The expected effect of these substances may be illustrated by risk assessment. Therefore the paper seeks to assess the health risk for inhabitants of Obninsk using drinking water.Health risk assessment in the analysis of environmental quality involves four basic stages [3].Hazard identificationExposure assessmentEvaluation of the “dose-response” relationshipRisk characterizationThe following objects were chosen for studies:Water from springs near the nuclear power enterprise SRC RF Institute of Physics and Power Engineering near Obninsk; these springs are widely popular among citizens.Water from three intakes of the centralized water services of Obninsk.Information of the State Sanitary InspectionAt the first stage the ecological diagnostics has been performed of water from the chosen springs by biotesting. Water quality was estimated from the following biological parameters: barley germination, biomass and the length of sprouts and barley root parts, the mitotic index and chromosome cell aberrations in the barley root meristem, wheat coleoptile increment. The diagnostics allowed these springs to be divided into three groups.The first group includes the springs (No.16 as an example) most distant from city and relatively satisfactory. The second group includes the springs (No.2 is typical) located near Obninsk where according to biotesting water practically doesn’t differ from that in the town water pipelines.Springs with water considered to be the worst both from hygienic and ecological points of view are referred to the third group (e.g. spring No.7). The smallest increment of wheat coleoptiles and a 2 fold excess in a number of aberrant cells relative to the reference level (pipeline cold water) are observed in phytotests (plants) grown in this water. So, it was assumed that water from the third group exhibits mutagenic and, probably, carcinogenic activity.Only one spring was chosen from each group. The content of chloride-, sulphate-, nitrate- and carbonate-anions, calcium and magnesium cations and pH, mineralization, general rigidity and gamma-radioactivity were determined by physical and chemical methods. Besides, concentrations of aluminum, beryllium, boron, arsenic, mercury, lead, silver, chromium and zinc were assessed. Table 1 shows some of these data.The studies have shown that the excess was registered for such substances as beryllium, cadmium, mercury and chromium. According to biological and chemical analysis, water from spring No.7 was found to be the worst. In the future when estimating the risk just this spring will be used because of the greatest number of factors capable of causing adverse human impact.At the second stage the results of chemical analysis of water from three Obninsk water intakes (1990–2003) have been investigated. The analysis has shown that during last several years the excess of iron (up to 1.9 mg/l), the increased rigidity (up to 7.8mg - eqv/l) and fluorine content close to MPC (1.5mg/l) were observed in all three water intakes. Besides, there were variations in mineralization values (from 280mg/l to 500mg/l), pH (from 6.2 to 7.6) and such components as sulphates (from 14 to 136), nitrates (from 0.04 to 6.3), chlorides (from 3 to 35), but their concentration did not exceed MPC.So, the assessment of adverse factors has shown that heavy metals are the most essential pollutants of spring water. Other pollutants are not analyzed. For pipeline water the worst parameters are associated with iron content, general rigidity as well as with cadmium and chromium. These factors are taken into consideration at the next stages of risk assessment.According to the Environmental Protection Agency (US EPA) technique we should take into consideration the following. Risk is calculated under the condition of this water consumption every day during the whole human lifetime. The water quality standard for calculating risk is also specified for the same period. The average amount of water used every day for drinking is about 3 liters, the mean body weight is 70 kg. So, under these conditions the dose of a chemical substance taken by a person with drinking water every day is:where,ADDd is the dose taken with drinking water;BW is the body weight, 70 kg;C is the substance content in water, mg/l;DW is the mean volume of water drank every day, 3 liters.Let us estimate a daily mean dose of the individual’s uptake of contaminants with water. Calculations have been performed using analyzed data on spring (No.7) and pipeline water. Table 4 presents the data obtained.This stage of activities seeks to represent the quantitative relationship between impact and risk appeared. A linear model suggested by US EPA for risk assessment in our case can be expressed by the following formula:where,Risk is the risk of adverse health effect estimated as the probability of this effect under given condition;ADDd is the daily substance dose, mg / kg;UR is the risk unit specified as a factor of risk proportion depending on the available concentration (dose). The risk unit UR accepts the true value depending on the impact (carcinogenic, non-carcinogenic) which this substance has and the substance itself.Table 2 shows the calculated health risk in cases of using spring and pipeline water. Water contains many different components and the risk of a combined impact of contaminants (see Table 4) can be determined from the formula:where,Risksumis the risk of a combined impact of contaminants;Risk1 … Riskn is the risk of impact of each isolated contaminant.As a result of spring water consumption the risk of oncological diseases will come to 3.95 × 10−3 and non-oncological diseases to 0.98 × 10−3. In terms of per capita (approximately 1000 persons) it means that 4 persons would be in danger of oncological diseases and 1 person of non-oncological diseases.In case of pipeline water resources the carcinogenic risk of a combined impact amounts to 5 × 10−6 and non-carcinogenic risk to 34 × 10−6, that for the inhabitants of Obninsk means 0.54 and 3.7 cases of carcinogenic and noncarcinogenic effects, respectively.Until recently it was taken as an axiom that aluminum was harmless for people, animals and plants. In fact, in its content in the Earth’s crust (8.8 %) aluminum is the third after oxygen and silicon and the most widely spread among metals. Thanks to its extremely high reactivity, aluminum quickly forms insoluble compounds and becomes practically safe for plants and animals as it doesn’t penetrate into their cells and tissues. Under certain conditions, however, e.g. under the action of acid rains, aluminum can pass to the ionic state and react with biological objects changing them or suppressing their function [4].Data published in the 1990s have shown that high concentrations of inorganic (free) aluminum were found in the surface and ground water owing to acid rains or any other reasons.Recent observations show that aluminum content in ground water near Obninsk has increased. It is possibly caused by the above reasons. So, aluminum content in springs in the late 1990s was as follows (table 3).The chemical risk will be estimated from these data using the technique proposed (table 4).The obtained data show that human health risk when drinking spring water with the increased aluminum content ranges from 0.94 to 2,80 cases (per 1 thousand of spring water consumers) of non-carcinogenic effects depending on the number of a spring where water is taken. Though the scientific literature does not cite the unit values to calculate the potential carcinogenic peroral risk for aluminum, the information is available that the increased aluminum content in drinking water may cause chromosome aberrations in barley meristem cells [4]. Therefore further studies of mutagenic and carcinogenic impact of aluminum on living organisms are to be carried out.Processed data show that water with the increased aluminum content can cause approximately three times as many negative non-carcinogenic effects for human health than other substances contained in spring water and possessing non-carcinogenic effects. At the next stage we tried to associate the physical and chemical parameters of water with the population sickness rate to demonstrate clearly the effect of water quality on human health.High levels of general rigidity of drinking water, typical of Obninsk can have a direct bearing on human health. Russian and foreign scientists disagree on the impact of this factor on human health. Besides, there are no criteria in literature for quantitative assessment of the expected risk associated with increased drinking water rigidity. Therefore we tried to associate the rigidity parameter with the sickness rate on the basis of ecological and epidemiological risk assessment of drinking water contamination [5].Harmful environmental impact on human health reveals itself as the increased sickness rate. To study human health we used an ecological approach where temporal variations in the sickness rate were compared with those in the environmental contamination levels. This approach is based on the information from “Report on the cases of diseases registered for patients living in the area of Obninsk medical services». Among a great variety of nozological forms of diseases the approach takes account of those which are most widely spread, adequate to the harmful effect and pollutants; considered are also the informative sanitary parameters available.In our case we tried to establish the relationship between communication of diseases and drinking water rigidity in the centralized water supply. Epidemiological data on the disease structure and cases for inhabitants of Obninsk served the basis for these studies.Comparison of the water rigidity and the disease structure allowed certain relationships to be established. So, genito-urinary, stomach ulcer and duodenum diseases, cerebrovascular diseases and diseases of skin and hypodermic tissue are closely connected with this parameter of drinking water. The correlation coefficient for them is 0.70–0.94.Based on the correlation analysis it is possible to conclude that the rigidity value has 4 of 11 important correlation coefficients with the population sickness rate indices in Obninsk. It means that probably each third sickness rate index most probably is associated with drinking water rigidity.Our data confirm that the risk of diseases in case of spring water consumption is hundreds times higher as compared to pipeline water. Uncontrolled consumption of spring water unfavorable in its composition may be dangerous for human health. According to these studies, pipeline water is more preferable to people, but some risk of carcinogenic and non-carcinogenic effects for inhabinats of Obninsk is also not excluded, however, it is 1000 times lower than the real number of cancer cases in Obninsk.Chemical composition of some Obninsk springsElement excess over MPC.Calculated carcinogenic and noncarcinogenic risk of groundwaterAluminium content in spring water near ObninskCalculated health risk for drinking water containing aluminum
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+ We investigated the changes in the properties of water when exposed to sunlight for 40 days. We hypothesize and prove that solar irradiation to water entraps electromagnetic radiation as potential energy, which becomes kinetic energy in various systems. It is postulated that photochemically-induced energy transfers, associated with individual spectral emission of visible spectrum of solar light, exert diverse influences on biological systems. Bottles of distilled water, individually wrapped in spectral-colored cellophane were exposed to sunlight and compared to an unwrapped bottle to determine chemical and physical changes as well as modifications of biological properties. Each bottle of water was named according to the color of cellophane paper with letter E (stands for exposed) as a prefix with (E-violet, E-indigo, E-blue, E-green, E-yellow, E-orange, and Ered). E-control (without wrap) was exposed to polychromatic sunlight. This study addresses two main issues viz., the chemical and physical changes in E-water and its effect on biological activities. Chemical and physical composition analysis using inductively coupled plasma atomic emission spectrometry; physical conductance by a Wheatstone Bridge type conductivity meter; osmolarity by a vapor pressure osmometer; and, salt solubility profile of 10% sodium bicarbonate were determined. Furthermore, testing the effect of E-waters on human lymphocyte proliferation, mosquito larvae hatching and seed germination determined the functional role of solar radiation through specific spectrum/s of visible light on various biological processes. We found that water exposed to visible spectral emissions of sunlight had an altered elemental composition, electrical conductance, osmolarity and salt-solubility, as well as differences in bio-modulatory effects. A gradual increase in leaching of Boron from E-violet to E-red was noted. E-indigo showed maximal increase in electrical conductance and maximal salt solubility of sodium bicarbonate. E-blue inhibited phyto-hemagglutinin-induced immune cell proliferation and mosquito larvae hatching. E-orange stimulated root elongation in seed germination. We conclude that 40-day exposure of water to specific solar spectrum changes chemical and physical properties and influences on biological activity.Water is an essential constituent of all living organisms, and sunlight is our main source of energy [1]. There is an absolute necessity in daily life for water, especially uncontaminated water. In order to keep water free from life-threatening contaminants, an elaborate purification system is in place in western countries that is not found feasible in some Third-World countries where the water sources usually have fecal contaminants. Exposure of contaminated water to sunlight can be sterilizing [2–5]. The process is dependent upon the composition of the vessel, as well as intensity and time of exposure [2]. Placing bottled (plastic/glass) water in sunlight for a minimum of 5 hrs and up to a maximum of 48 hrs has been scientifically shown to detoxify and decontaminate water infected with bacteria and viruses [3–5]. Extended times of exposure to sunlight for approximately six weeks, has been used for solarization of soil [6].Water alone has also shown potential for healing. Healing with water containing various additives as calcium chloride, magnesium sulfate, sodium metasilicate and sulfated castor oil has been patented by Willard [7]. In ancient times, water, after exposure to direct sunlight and filtered through colored glass, was used as a therapeutic modality [8]. The action of sunlight on bottled water, as used in the sterilization study, had been reported as irreversible and documented by the inability of coliforms to re-grow after placing the exposed bottled water in the dark storage [4]. However, biomodulatory and physico-chemical effects of individual spectrums of visible light on water have not been characterized.It is our hypothesis that solar irradiation to water entraps electromagnetic radiation as potential energy which becomes kinetic energy in various systems. The central hypothesis tested in this paper assumed that photochemically-induced energy transfers, associated with individual spectral emission of visible spectrum of solar light, exert diverse influences on biological systems. To test this hypothesis, sterile water in unopened capped plastic bottles (which allows 50% light transmission as per the information provided by the manufacturer) was used. Dye-impregnated cellophane paper covering clear glass or translucent (50%) plastic was used as an absorption filter to selectively allow exposure of semi-monochromatic light to water in these bottles. The cellophane wrapped bottles were exposed to sunlight through violet, indigo, blue, green, yellow, orange, and red colored cellophane paper in an open environment in an upright position. Each of the exposed (E) water was named according to the color of the cellophane paper, e.g., E-violet, E-indigo, E-blue, E-green, E-yellow, E-orange, E-red (Figure 1). E-control was exposed to polychromatic sunlight. Thus, the sun provides the heat in the daytime while in the absence of sun lower temperatures occur during nighttime. In essence, this provides a constant cycle of evaporation and condensation in the enclosed plastic bottle. Further, water when irradiated by the sun is exposed to decreasing amounts of energy of radiation; violet being high frequency of light energy whereas red is of lower frequency and higher heat index. After the exposure of 40 days, the bottled water was wrapped with aluminum foil to protect from exogenous irradiation.The overall aim of this study is to scientifically investigate the changes in the properties of water when exposed to sunlight for 40 days. This study specifically addresses the chemical and physical changes in exposed (E) water, and its effect on biological activity. The potential biological significance of changes in the E-waters was determined by testing their effect on: 1) lymphocyte (human peripheral blood) proliferation, 2) mosquito larva (Anopheles) viability and 3) seed (Diolichos Uniflorous) germination.Sterilized distilled water in sterile translucent (50% transparent) plastic (a co-polymer of 98% polypropylene and 2% polyethylene) bottles was obtained from Baxter (Deerfield, Illinois). Double-distilled water was obtained from the laboratory, and well water was obtained from a well in Columbus, Texas. The double-distilled water was placed in sterile 500-ml clear glass media bottles. Each of the glass bottles was then heat-sterilized in an autoclave at 121°C for 15 minutes, and wrapped in cellophane of a different color. Well water was used to fill cleaned 16-ounce clear glass sterilized bottles, and heat sterilized. Bottles were cellophane-wrapped corresponding to the visible individual spectral colors of violet, indigo, blue, green, yellow, orange, and red. The colored cellophane color distribution in the red (600–700nm), green (500–600nm) and blue (400–500nm) filter was measured on TQC90, a color transmission and reflection densitometer, Electronic Systems Engineering Company, Inc. (Cushing, OK). Violet wrap had 38% transmission (T) in the red filter, 12% T in the green filter and 55% T in the blue filter. Similarly indigo wrap had 4% T in red filter, 54% T in green and 79% T in blue; blue wrap had 27% T in red filter, 41% T in green filter and 52 % T in blue filter; green wrap had 6% T in red filter, 50% T in green filter and 10% T in blue filter; yellow wrap had 49% T in red filter, 44% T in green filter and 13% T in blue filter; orange wrap had 83% T in red filter, 54% T in green filter and 3% T in blue filter; red wrap had 74% T in red filter, 0% transmission in green filter and 4% transmission in blue filter. Controls included one unwrapped bottle of each type of water; i.e., well water and double-distilled. The distilled water and well-water were exposed to sunlight from July through August in Columbus, Texas. The double-distilled water in plastic bottles was exposed to sunlight from October through November in Jackson, Mississippi. Bottles (minimum of two bottles for each cellophane paper) were exposed to sunlight for 40 days and thereafter the bottles were placed in the dark to avoid unwanted light exposure without removal of cellophane. All experiments used water obtained from plastic bottles except for studies in germination.For chemical composition analysis, eight E-water samples were placed in 15-ml sterile tubes wrapped with aluminum foil to protect them from further exposure to radiation. These samples were sent to NASA-Marshall Space Flight Center, Huntsville, AL for analysis using inductively coupled plasma atomic emission spectrometry (ICP) [9].Salt solubility of 10% sodium bicarbonate was evaluated in each of the eight E-waters. Eight 15 ml sterile polystyrene tissue culture tubes were filled with 1 gram of sodium bicarbonate and 10 ml of the different E-waters at room temperature. The tubes were vortexed to dissolve the solute. After 24 hours, solubilization was determined macroscopically.Physical conductance of each E-water was measured using a Wheatstone Bridge type conductivity meter. Osmolarity of E-waters was determined using the Wescor 5100 vapor pressure osmometer, Wescor, Inc. (Logan, UT).Lymphocytes were incubated with phytohemagglutinin (P), a T cell mitogen, in presence or absence of E-waters. In a cell proliferation assay [10], peripheral blood (10 ml) was obtained from healthy subjects and lymphocytes were purified using ficol-hypaque gradient. Lymphocytes were removed, washed and resuspended in RPMI 1640 containing 1% penicillin, 1% streptomycin and 10% fetal bovine serum using standard laboratory procedures. A mixture of 2 × 105 cells in 100μl was plated into each well of a flat-bottom 96-well plate and incubated both with media alone and media supplemented with phytohemagglutinin at 10 ng per ml. Ninety μl of E-water group was plated into their respective wells in triplicate. Ten μl of 10x phosphate buffered saline (PBS) was placed in each well to make the solution isotonic. The samples were incubated for 3 days at 37°C, pulsed with 1μCi of tritiated thymidine, harvested after 18 hours onto filter paper, and radioactivity was determined in a beta counter.Mosquito larvae, particularly the IV-in stared stages of the Anopheles species, were removed from an outdoor stagnate water supply and ten larvae were placed in 12-well plates for viability testing. These ten mosquito larvae were mixed with 500μl of pond water plus 500μl of E-water covering the full spectrum of light. The E-waters were diluted 1:10 and 1:100 where the dilutions were made with unexposed distilled water (Baxter, Deerfield, IL). The plates were incubated for 10 days (normal hatching time for mosquito larvae) at room temperature for normal morphogenesis to take place and the viability was determined at 24, 48, 72, 96 and 240 hours. A mosquito was considered viable if the wings had emerged. The experiments were performed twice in triplicate. Data shown in the Table 2 were observations from t = 24 hrs, 48 hrs, 72 hrs, 96 hrs. The experiment was carried out for 10 days (240 hrs), however, no difference was noted between 96 and 240 hrs (data not shown). Three conditions were tested for each time period, i.e., the mosquito larvae were incubated with undiluted E-water plus two dilutions of E-water at 1:10 and 1:100.Eight petri dishes were filled with 7 grams of Horsegram (Dolichos Uniflorous) seeds [11]. Each petri dish with a 4 × 4 cm cotton gauze was filled with a different E-water (25 ml), and the seeds were allowed to incubate at room temperature for up to 48 hours. In this experimental design, the effects of energy on the material of the container (transparent versus translucent), season of incubation (summer versus fall), and the type of water (well water versus distilled water) were tested. Three types of water were tested viz., a) distilled water in plastic bottles (DWPB; fall), and b) well water in glass bottles (WWGB; summer) and c) distilled water in glass bottles (DWGB; summer). The cotyledons of the seeds were measured (in millimeters) from their point of visible emersion. The seeds were then incubated for an additional 24 hours and measured again. Three separate experiments were performed in duplicate, one using well water, one using sterilized double-distilled water in sterilized transparent glass bottles, and one using sterilized distilled water in sterilized translucent plastic bottles (Commercially available from Baxter).Statistical analysis was performed by one-way ANOVA for multiple samples by multiple comparison (turkey test) or by Student’s-t-testing with matched pairing if appropriate, as well as regression analysis.First, the chemical analysis was performed on 72 elements (Table 1A) of which only boron, calcium, magnesium and rubidium showed coefficient of variances (CV) greater than 10 when compared to E-control (Table 1B). In particular, boron concentration in parts per billion (ppb) steadily increases from E-Violet water (13,137 ppb) to E-Red water (27,563 ppb), while E-control corresponded to 12,536 ppb. The compositional distribution of Boron according to the photon energy in electron volts (eV) has been depicted in Figure 2. The Binomial equation for the compositional distribution of boron which gives the best fit is:wherey = photon energy in electron volts (eV), andx is the atomic mass (ppb).The pH of different E-waters was determined and found to be about 5.5. The addition of sodium bicarbonate to water changes the solubility profile. The solubility showed a variable pattern where the sodium bicarbonate solubility of E-control was between E-green and E-blue. Interestingly, E-indigo exhibited maximal solubility while E-orange exhibited greatest insolubility (Figure 3).Physical conductance of each E-water as measured, using a Wheatstone Bridge type conductivity meter, was found to be five-fold higher in E-Indigo than in E-control (p<0.01; Figure 4).A gradual increase in osmolarity between E-Violet and E-Yellow; and between E-Orange and E-Red was noted (Figure 5A). However, E-Yellow, E-Orange and E-Red waters exhibited higher osmolarity when compared with E-control. The binomial equation for the distribution of osmolarity which gives the best fit is:wherey = photon energy in electron volts (eV), andx is the osmolarity (mmol/Kg) (Figure 5B).Second, the effect of E-water on T cell proliferation was assessed. T cells in media alone and in the presence of P, are considered as base value (Figure 6). While most E-water tests showed some degree of suppression of T-cell proliferation, E-blue showed complete inhibition of PHA stimulated T cell proliferation. The viability of cells was ≥90% in all conditions as measured by trypan blue, suggesting that the inhibition in proliferation was not due to some toxic reagent.Third, the impact of the irradiated water on the growth of mosquito larvae was assessed. Data shown in Table 2 are observations from t=24 hrs, 48 hrs, 72 hrs, 96 hrs. The experiment was carried out for 10 days (240 hrs), however, no difference was noted between 96 and 240 hrs (data not shown). These studies indicate that E-violet and E-indigo showed similar values to E-control for viability when tested undiluted, while E-blue and E-green greatly inhibit larval growth for up to 72 hours. E-yellow, E-orange and E-red promoted the hatching of mosquito larvae, except E-red in dilutions of 1:10 and 1:100. E-control showed only 1/3 hatching compared to E-green after 96 hours (Table 2). E-blue completely abolished larvae growth in the three concentrations tested (undiluted, 1:10, 1:100).Finally, the effect of E-water on seed germination was investigated (Table 3). After 48 hours of incubation, the seeds in the DWPB showed significant enhanced growth (p<0.05) with E-indigo, E-yellow, and E-orange, while seeds in the DWGB showed enhanced growth with all colors when compared to E-control. WWGB showed growth in E-violet, E-indigo, E-green, E-yellow and E-orange after 48 hours but not in E-blue or in E-red. In contrast, after 72 hours, the seeds in the DWPB showed the most significantly enhanced growth with the E-yellow and E-orange water (p<0.05, <0.01 respectively). The WWGB showed enhanced growth with E-yellow and E-orange and the DWGB showed enhanced growth with E-violet, E-indigo, E-yellow (p<0.05) and E-orange (p<0.001). E-red water suppressed seed germination at 72 hours in all the three water groups tested.Plastic vessels allow only 50% transmission of light while glass vessels allow about 90% transmission. This means that there is more solar energy delivered in glass versus plastic. In germination experiments we demonstrated that maximal root elongation and minimal root elongation were independent of the vessel i.e., plastic vs glass but were dependent on the wavelength of light rather than intensity. Thus, implying that the energy transfer is dependent on the wavelength rather than the amount of sunlight the vessel receives. Also the effect of organic material like plasticisers for example dibutyl pthalate and other esters used in the synthesis of plastic bottles is not known. Whether the variation of conductance, osmolarity, salt-solubility and chemical composition of the different types of water had any impact on the germination process, mosquito larva survival, or cell proliferation requires further investigation. Likewise, the influence of lunar irradiation, if any cannot be ascertained. Classically, electronic transition like n→n, π→π* and n→π occur in the ultraviolet region which can lead to photolytic effect of water but the outcome of these transitions in the visible spectrum is unknown. It is possible that the photochemical effect observed could have been a thermal effect based on solar spectral irradiance ([2]. This postulate cannot be authenticated because the temperature of the E-waters in the bottles was not recorded during the experiment. The conductivity, mosquito hatching, T cell-data, some ICP data and the osmolarity data seem to mimic the solar energy spectrum with a maximum effect in the blue-green region. Furthermore, the E-waters were tested at least three months after being at room temperature where the thermal effect if any, would have been equilibrated.Since boron is constantly increased in concentration from violet to red, the change in elemental concentration could possibly be due to the overall excess heating by red cellophane paper, which allows more leaching. It is not known whether the leached elements were present due to the process of the manufacture of the polypropylene-polyethylene plastic. On the other hand, we know that the distilled water was handled in the same manner for all test-bottles. Since the elemental composition is different after 40-day exposure at certain wavelengths of sunlight, there is an implication that there is a solar effect causing leaching of some components contained in the plastic. Further investigation is on going to determine the source of boron and other elemental composition after irradiation. Swiss studies, however, have shown that photoproducts are formed after solar exposure on the outer surface of the plastic bottle and that no substances hazardous to the health are leached into the water through the bottle [4]. In contrast to the Swiss study, in our study we report leaching of elements into the water [3].The exposure of water to sunlight has no influence on the pH of water. It has been reported that a solution of sodium bicarbonate in 10 parts of water is soluble at 25°C with a resulting pH of 8.3 when freshly prepared [13]. Thus, using sodium bicarbonate solubility as an example at room temperature, we are at the edge of solubility in which the solute must be fairly soluble. Whether the greater insolubility of the E-orange water as compared to E-control was due to density clouds of water molecules increasing the forces of repulsion, or whether the forces of attraction of the sodium bicarbonate were reduced is not known. Attractive forces in the molecular structure of sodium bicarbonate may have undergone changes in order to sediment rather than solubilize as observed with E-orange. The ion-dipole interactions seem to be predominant in terms of increasing ΔH of solubility. In other words there is maximal enthalpy of solution (ΔH) for E-indigo while E-orange has minimal enthalpy of solution in our study. This is a clear-cut indication that exposure to specific solar spectrum energizes the water resulting a change in the pattern of salt solubility.Leached chemicals such as boron, calcium, magnesium and rubidium appear in such minute quantities (ppb) as would not cause difference in conductance. Thus, the increased conductance such that we observed with E-indigo must be due to some modification by the irradiation process. Because there is no significant contribution of the concentration of elements, it is reasonable to assume a constant current, implying that the voltage generated by the red cellophane-wrapped bottled water would be greater than the voltage generated by the indigo cellophane-wrapped bottled water (voltage × conductance = current). Thus, the differences in voltages would cause a potential difference, meaning that energy is trapped in the water and manifesting as a potential difference. It is very unlikely that the difference in conductance is due to the impurities that create differences in ionic composition due to the use of distilled water which is devoid of contamination. It is noteworthy to mention that in a separate study; only E-yellow reacted with prostate related substance in the urine of cohort of patients with a serum prostate specific antigen in the range of 0.21–4.0 ng/ml [14].Elements present in the water contribute to the osmolarity of E-water. The added total of all elements in the system such as the leached chemicals contribute minimally. Considering boron by itself, concentration ranges from 1.16 mM baseline (control) to 2.55 mM in E-red while the sums of the other elements are in the nanomolar range. Such low concentration of the leached chemicals as measured in these waters would not account for this difference in osmolarity, which was found after the irradiation. Thus, the alteration in osmolarity that we observed is related to some unique phenomenon which is not defined by the elemental composition of the tested waters.Healing is associated with the immune system. In disease eruptions, suppression of an immune response via the T cell network, is of utmost importance in various instances as in transplant immunology and autoimmune diseases. Data on T cell proliferation demonstrated that E-blue showed maximal inhibition of PHA stimulated T cell proliferation by a non toxic mechanism. The immunosuppression of T cells lends credibility to the possibility of a particular E-water to be of importance in a disease paradigm.Solar energy has been used to detoxify water infected with viral and bacterial contamination [2–5]. However, solarized water has not been used to address the problem of mosquito biology. Since mosquitoes can spread disease the issue of mosquito control is of great importance. All mosquitoes spend their larval and pupal stages in water and our study has used this observation to study mosquito viability. These studies indicate that E-violet and E-indigo showed similar values to E-control for viability when tested undiluted, while E-blue and E-green greatly inhibit larval growth for up to 72 hours. E-yellow, E-orange and E-red promoted the hatching of mosquito larvae, except E-red in dilutions of 1:10 and 1:100. E-control showed only 1/3 hatching compared to E-green after 96 hours (Table 2). The most interesting finding was that E-blue completely abolished larvae growth in the three concentrations tested (undiluted, 1:10, 1:100). The abolishment of mosquito larva in E-blue water at different dilutions provides evidence to the possibility of spraying stagnant waters with the E-blue water to destroy mosquitoes.Water is the most abundant ingredient of active plant cells [10]. Seed germination does not require sunlight or soil, but it requires water [10]. Germination aids in the hydrolysis of the complex into simple sugars that are readily utilized in the synthesis of auxins and proteins [15]. The auxins help to soften cell walls to facilitate growth and the proteins are readily utilized in the production of new plant tissues [15]. Thus, germination when accelerated is capable of improving crop yield and boasting farmers income in agricultural industry involved in growing crops such as tomatoes (Lycopersecum esculentum Mill) and tamarind (Tamarindus indica L) [15, 16] Our germination studies showed that regardless of the vessel at 72 hours, E-orange enhanced while E-red inhibited seed germination when compared to E-control. Results indicate that a change took place independent of the source of water (well water versus double distilled), type of container (plastic or glass), or the season of the year (summer versus fall). Distilled water when compared to well water has a reduced load of microorganisms. The change in photochemical property of well water in comparison to distilled water maybe due to the presence of microorganisms in the well water. The mode of energy transfer however, as these investigations indicate, seems to be governed by a process independent of the presence of bacterial and viral contaminants. In both well and double distilled water at 72 hours with E-orange there was a maximal activation of seed germination. Thus, change in the pattern of seed germination by E-water appears to be solely dependent on the wavelength of sunlight rather than the type or source of water used. E-orange showed enhanced growth both in plastic containers and glass containersIn conclusion, in this paper we have shown that: 1) exposure of water to specific solar spectrum for a period of 40 days alter a) boron concentration, b) Δ H of solubilization with sodium bicarbonate, c) E-indigo induced five-fold greater conductivity in the water when compared to the effect of E-red, d) osmolarity, e) T cell proliferation by total suppression of mitogen stimulated T cell proliferation by E-blue, f) mosquito larva hatching abolished by E-blue but not other E-waters, and g) germination with E-orange giving the maximal elongation of bean roots. Additionally, our results (Figure 5A and B) suggest for a unique phenomenon of osmolarity change without involvement of the concentration of the solutes present in the water. This is the first study to show that the penetration of sunlight through specific colored cellophane wrapped bottles changes chemical, physical and biological properties of water.The results of these changes are summarized in Figure 7. There is strong indication that 40-day incubation in sunlight results in storing solar energy as potential energy, which can be utilized as kinetic energy in various biological systems. These results are novel and open new possibilities for the investigation and understanding of the role of exposed water through different colors of the solar spectrum for plant, insects and human life.Diagrammatic representation of the Chemical, Physical and Biological modifications in properties investigatedDistribution profile of boron leeching from the plastic bottle into the water after exposure to specific photon energy of visible spectrum of lightEach tube contains 1.0 gram of sodium bicarbonate with different E-water. Macroscopic evaluation exhibits maximum solubility with E-I and maximal insolubility with E-O.The Conductivity of different E-waters using conductance meter (Wheatstone Bridge). Measurements were taken in triplicate and are represented as Mean ± SD.Osmolarity study of different E-waters by Osmometer. Measurements were taken in triplicate and are represented as Mean ± SD.Mathematical Relationship between photon energy and osmolarity of different E-watersThe proliferative responsiveness of peripheral blood T cells with different irradiated waters to phytoheamagglutinin (P). Cells were incubated with media, media + P and with media + P + E water passed through 0.22μ filter. The experiments were done thrice in triplicates and are represented as Mean ± SD.Maximum changes in properties of water as specific wavelength range.List of 72 elements analysed by Inductively Couple Plasma Spectroscopy.Variation of elements contained in spectral irradiated distilled water as measured by Inductively Couple Plasma Spectroscopy (PPB).El = elementMean ± S.D. of % mosquito larvae survival in differing concentrations of E-waters at t=24, 48, 72 and 96 hoursMean ± SD in millimeters of the growth of cotyledons of gram beans following incubation with E-waters at t=48 and 72 hrs. Mean of E-control is compared to other E-waters for enhanced growth by multiple comparison procedure (Turkey test). Three types of water tested were viz; Distilled-water in plastic bottles (DWPB), Well-water in glass bottle (WWGB) and Distilled-water in glass bottle (DWGB).We thank Dr. Manoj Shukla, Dr. Ramaiyer Venkatraman, Dr. Johanne Bauer, Dr. Ingrid Glurich, Dr. Cindy Cohly and Taylor Schwalenberg for their valuable criticisms in the preparation of the manuscript, and Dr. A. K. Markov for helpful discussions.
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+ Phytoextraction is gaining acceptance as a cost-effective and environmentally friendly phytoremediation strategy for reducing toxic metal levels from contaminated soils. Cognizant of the potential of this phytoremediation technique as an alternative to expensive engineering-based remediation technologies, experiments were conducted to evaluate the suitability of some plants as phytoextraction species. From one of our preliminary studies, we found that tall fescue (Festuca arundinacea Schreb. cv. Spirit) can tolerate and accumulate significant amounts of lead (Pb) in its shoots when grown in Pb-amended sand. To further evaluate the suitability of tall fescue as one of the potential crop rotation species for phytoextraction, a study was conducted to determine whether the addition of ethylenediaminetetraacetic acid (EDTA) alone or in combination with acetic acid can further enhance the shoot uptake of Pb. Seeds were planted in 3.8 L plastic pots containing top soil, peat, and sand (4:2:1, v:v:v) spiked with various levels (0,1000, 2000 mg Pb/kg dry soil) of lead. At six weeks after planting, aqueous solutions (0, 5 mmol/kg dry soil) of EDTA and acetic acid (5 mmol/kg dry soil) were applied to the root zone, and all plants were harvested a week later. Results revealed that tall fescue was relatively tolerant to moderate levels of Pb as shown by non-significant differences in root and shoot biomass among treatments. An exception to this trend however, was the slight reduction in root and shoot biomass of plants exposed to the highest Pb level in combination with the two chelates. Root Pb concentration increased with increasing level of soil-applied Pb. Further increases in root Pb concentrations were attributed to chelate amendments. Translocation index, which is a measure of the partitioning of the metal to the shoots, was significantly enhanced with chelate addition especially when both EDTA and acetic acid were used. Chelate-induced increases in translocation indices correspondingly led to higher shoot Pb concentrations.Metals are anthropogenically released into the environment at increasing rates by mining, industry, and agriculture, causing serious problems for environmental and human health [1–4]. In the USA alone, more than 50,000 metal-contaminated sites await remediation, many of them Superfund sites [5]. In spite of the ever-growing number of toxic metal-contaminated sites, the most commonly used methods dealing with heavy metal pollution are either the extremely costly process of excavation and burial or simply isolation of the contaminated sites. Such cleanup is practical only for small areas, often a hectare or less [6]. Recent U.S. remediation costs have been estimated at $7 billion to $8 billion per year, approximately 35% of which involves remediation of metals [7].Recently, heavy metal phytoextraction has emerged as a promising, cost-effective alternative to the conventional engineering-based remediation [8–12]. The objective of phytoextraction is to reduce heavy metal levels below regulatory limits within a reasonable time frame [9]. To achieve this objective, plants must accumulate high levels of heavy metals and produce high amounts of biomass. Early phytoextraction research dealt with hyperaccumulating plants, which have the ability to concentrate high amounts of heavy metals in their tissues [10, 13]. However, hyperaccumulators often accumulate only a specific element and are slow-growing, low-biomass-producing plants with little known agronomic or horticultural attributes. Moreover, there is no known hyperaccumulating plant for Pb, one of the most widespread and toxic metal pollutants in soils.Previous hydroponic studies revealed that uptake and translocation of heavy metals in plants are enhanced by increasing heavy metal concentration in the nutrient solution [14]. The bioavailability of heavy metals in the soil is therefore, of paramount importance for successful phytoextraction. Lead has limited solubility in soils, and its availability for plant uptake is minimal due to complexation with organic and inorganic soil colloids, sorption on oxides and clays, and precipitation as carbonates, hydroxides, and phosphates [15–16]. Therefore, successful phytoextraction must include mobilization of heavy metals into the soil solution that is in direct contact with the roots. In most soils capable of supporting plant growth, the readily available levels of heavy metals, especially Pb, are low and do not allow substantial plant uptake if chelates are not applied. Chelates have been shown to desorb heavy metals from the soil matrix into soil solution [17], facilitate metal transport into the xylem, and increase metal translocation from roots to shoots of several fast-growing, high-biomass-producing plants [18–27].Using a Pb-amended sand [28], tall fescue (Festuca arundinacea Schreb. cv. Spirit) was identified as a potential phytoextraction species because of its high biomass yield under elevated Pb levels and its ability to translocate high amounts of Pb into its shoots. The main objective of this study was to further evaluate the effectiveness of tall fescue as a phytoextraction species. We envisioned that this species can be used in a crop rotation scheme during the colder months, and also serves as cover crop for an otherwise barren metal-contaminated soil. Specifically, this experiment was conducted to determine whether pre-harvest amendments of ethylenediaminetetraacetic acid (EDTA) alone or in combination with acetic acid can further enhance the shoot accumulation (i.e., translocation index) of Pb by tall fescue grown on a Pb-contaminated soil.Plants were maintained under a naturally-lit greenhouse with 31°C/20°C day/night temperatures. Supplemental light for 12 hours were provided by high intensity super halide lamps (1000 W H.Y. Lites Horizontal System, High Yield, Inc., Camas, WA). The photosynthetically active radiation (PAR; 400–700 nm) measured at the canopy level was no less than 1400 μmol photons m−2 sec−1 as measured with a LI-COR 6200 portable photosynthesis system (LI-COR, Inc., Lincoln, NE). Tall fescue (Festuca arundinacea Schreb. cv. Spirit) seeds were obtained from Hutto’s Garden, Jackson, MS. Unless otherwise specified, six seeds were sown in each 3.8 L plastic pot containing a growth medium composed of sieved silty clay loam soil (pH 8.2; 1.5% organic matter), peat, and sand mixed in 4:2:1 volumetric proportions. Emerged seedlings were thinned out to 5 plants per pot at 5 days after planting. Using a hand trowel, three concentrations (0, 1000, 2000 mg Pb/kg dry soil) of lead (supplied as lead nitrate) were thoroughly mixed with the soil. The Pb levels used are within previously reported Pb concentration ranges found in various contaminated sites [12, 18, 19]. The Pb-spiked soils with moisture contents maintained at field capacity were then allowed to equilibrate inside the greenhouse for 21 days prior to planting. Using previously described sequential extraction procedures [29], this Pb-spiked soil had the following percentages of Pb distributed in the various soil fractions: exchangeable (34.5%), carbonates (43.5%), Fe-Mn oxides (11.5%), organic matter (4.3%), residual (6.2%). Plants were watered every 2 to 3 days, depending on the evaporative demand, with full strength nutrient solution [30–31]. EDTA (0 or 5 mmol/kg dry soil) was applied as a 100 mL aqueous solution one week before harvest (pre-harvest). Moreover, a 100 mL aqueous solution of acetic acid (5 mmol/kg dry soil) was also added to some of the treatments one week before harvest. On average, 100 mL of nutrient solution were added to each pot to ensure that soil moisture content was maintained at field capacity and that no excess soil moisture drained from perforations at the bottom of each pot. A 17.8-cm plastic saucer was placed beneath each pot to prevent cross contamination among treatments.Any symptoms of metal toxicity (e.g., discoloration, pigmentation, yellowing, necrosis, stunting) exhibited by plants were visually noted during the experimental period. All plants were harvested at seven weeks after planting. For dry biomass determinations, shoots and roots were separated during harvest then oven-dried at 70°C for 48 hours. Prior to oven drying, roots were washed with distilled water to remove any adhering debris.Dried samples were weighed and ground in a Wiley mill equipped with a 425 μm (40-mesh) screen. Lead contents of each 200 mg dry, ground plant tissue were extracted using previously described procedures [32] with slight modifications [30–31]. Briefly, 40 ml of 50% aqueous nitric acid were added to a 250 ml Erlenmayer flask containing a representative sample of ground tissue. The acidified sample was heated to 95 °C, refluxed for 15 minutes without boiling and then allowed to cool. Another 10 ml of 50 % aqueous nitric acid were added and the sample was again heated and refluxed for 30 minutes. The heated sample was allowed to cool, then completely oxidized in 5 ml concentrated nitric acid. The oxidized solution was allowed to evaporate to approximately 5ml without boiling. To initiate the peroxide reaction, 2 ml of deionized distilled water and 3ml of 30% hydrogen peroxide were added to the concentrated digestate and then heated until effervescence subsided. Another 7 ml of 30% hydrogen peroxide were added continuously in 1 ml aliquots as the digestate was again heated. The digestate was heated until effervescence was minimal and its volume reduced to approximately 5 ml. After cooling, the final digestate was diluted to about 100 ml with deionized, distilled water. The digestate was filtered through a filter paper (Whatman No. 1) and the final volume was adjusted to 100 ml with deionized, distilled water.Lead concentrations were quantified using atomic absorption spectrometry (Thermo Jarrell Ash Model AA Scan 4) and expressed as mg Pb/kg dry weight of plant tissue. This analytical system had a 98% recovery efficiency and detection limit of 5 parts per billion (ppb) Pb. Per cent translocation index (T.I.) was calculated using the formula described previously by Athalye et al. [33]: T.I. = (shoot Pb accumulation) × 100/ (shoot + root Pb accumulation).In this experiment, each treatment replicate consisted of one pot containing 5 plants. Treatments were arranged in a completely randomized design (CRD) with four replications. Data were analyzed using Statistical Analysis System (SAS). Treatment comparisons were done using Fisher’s Protected Least Significant Difference (LSD) test. In this study, a probability p≤0.05 was considered to be statistically significant.The Pb treatments alone did not significantly affect root biomass of tall fascue plants (Table 1). However, root biomass of plants grown at 1000 and 2000 mg Pb/kg were reduced by 24% and 28%, respectively with the addition of EDTA alone, and in combination with acetic acid. Both Pb and chelate amendments did not affect shoot biomass (Table 1). This observation was also supported by the absence of any discernible phytotoxic symptoms (e.g., stunting, chlorosis, necrosis, discoloration, pigmentation) exhibited by the shoots.Pb accumulations in the roots increased in a dose-response manner with increasing levels of Pb treatments (Fig. 1A). Such root Pb accumulations were enhanced with EDTA amendments. Further increases in root Pb concentrations occurred due to the synergestic effects of both chelates that were applied simultaneously. When no chelates were applied, shoot Pb concentrations slightly increased with increasing levels of soil-applied Pb (Fig. 1B). Dramatic increases in shoot Pb were observed with the addition of EDTA alone. However, when both chelates were amended, increases in shoot Pb were even more remarkable.One of the attributes of an effective phytoextraction species is its ability to maximize the amount of metal that is partitioned to the above-ground, harvestable biomass (i.e., shoots). This is indicated by the translocation index. In the absence of chelate(s), translocation indices of tall fescue plants grown at 1000 and 2000 mg Pb/kg were minimal, accounting for only 3.3% and 11.4%, respectively (Table 2). However, translocation indices dramatically increased to 66% and 73%, respectively when EDTA alone or both chelates were added simultaneously.Since total metal removal is a function of the metal concentration in the harvestable biomass (e.g., shoots), the first requisite in phytoextraction is the production of high plant biomass yield. This means that the plant species must be able to grow successfully at the contaminated site. Generally, F. arundinacea not only produced high biomass but was able to tolerate elevated levels of soil-applied Pb and chelates. There were no noticeable phytotoxic effects of Pb and/or chelates on tall fescue except for a slight reduction in root biomass (Table 1). It is known that metal phytotoxicity causes stress to the plant resulting in a reduction in biomass and eventual death (in some cases). Cunningham and Ow [34] described the presence of specific high-affinity ligands as one of the metal-resistance mechanisms existing in some plants. These ligands, which are natural metal-binding peptides known as phytochelatins and metallothioneins, make the metal less toxic to the plant, and at a certain EDTA threshold, these ligands may be activated. We are not certain whether this resistance mechanism also exists in tall fescue hence, further study is warranted. Vassil et al. [21] also demonstrated that free protonated EDTA (H-EDTA) was more phytotoxic to Brassica juncea than a Pb-EDTA complex. From earlier studies [25], we also observed the relative tolerance of wheat to Pb-EDTA complex. These previous findings support our present observation that a Pb-chelate complex was relatively nonphytotoxic to tall fescue, a monocotyledonous plant like wheat. Monocotyledonous species are usually more tolerant to metals than dicotyledonous species [35].Root Pb accumulation increased with increasing levels of soil Pb treatments (Fig. 1A). The addition of EDTA, one week before harvest, improved Pb accumulation by the roots. However, when both EDTA and acetic acid were applied a week before harvest, there was a significant increase in root Pb accumulation especially at the highest soil Pb treatment. Majority of the absorbed Pb remained in the roots when no chelate was applied (see Fig. 1A vs. Fig. 1B). This could be due to Pb binding to ion exchangeable sites on the cell wall and extracellular deposition mainly in the form of Pb carbonates deposited on the cell wall as previously demonstrated [36]. Root depth and density are important factors in phytoextraction. It was observed that during the duration of the study, the root system of tall fescue was extensive similar to the fibrous root systems of most monocots including wheat in our previous studies [25]. Roots provide a large surface-to-volume ratio to maximize the total uptake of various elements and compounds from the soil [37]. Using hydroponics systems, Dushenkov et al. [36] concluded that root Pb absorption is a rapid process and may be the fastest component of metal removal by plants. But, the ability of plants to translocate Pb to the shoots varies much more than their ability to accumulate metals in the roots [24, 38].One of the requisites contributing to the success of phytoextraction is the enhancement of Pb accumulation in the harvestable biomass (e.g., shoots). Vassil et al. [21] demonstrated that coordination of Pb transport by EDTA enhances the mobility within the plants of this otherwise insoluble metal ion, allowing plants to accumulate high concentrations of Pb in shoots. In this study, shoot Pb accumulation in tall fescue increased with increasing concentrations of Pb applied to the growth medium. This increase was especially remarkable in plants grown at 2000 mg Pb/kg with pre-harvest EDTA/acetic acid amendments (Fig. 1B). Lead, being a soft Lewis acid, forms a strong covalent bond not only with the soil, but with plant tissues as well [19]. It is believed that since the xylem cell walls have a high cation exchange capacity, the upward movement of metal cations are severely retarded [12]. Bringing the Pb into solution with a chelating agent, not only makes more Pb bioavailable for root uptake [18–19, 39]) but also moves the Pb that is sequestered in the xylem cell wall upwards and into the shoots.In an earlier study, Blaylock et al. [18] demonstrated with Indian mustard (Brassica juncea) that induced phytoextraction (i.e., equivalent to pre-harvest chelate amendment in our study) brings more of the Pb ions into solution and decreases the binding of Pb by the root tissue, thereby facilitating some of the desirable characteristics of a hyperaccumulator, such as high metal uptake by the roots, and translocation of the metal from the root to the above ground shoots. We believe that EDTA enhanced Pb desorption from soil to soil solution and facilitated transport from roots to shoots as previously demonstrated in EDTA-mediated phytoextraction studies using corn, peas and Brassica juncea[18–19, 21]. Corollary studies relating the available Pb levels in soils and chelate-mediated shoot Pb uptake are currently being investigated in our laboratory. Our preliminary results indicated that with EDTA application, there is a positive correlation between bioavailable Pb levels in soil and shoot Pb accumulation (data not shown).In a previous study by Blaylock et al. [18] using EDTA and acetic acid, the pH of the soil was decreased only slightly from 8.3 to 7.8. Similarly in our experiment, the pH of the soil before planting was 8.2, and decreased to 7.4 at harvest. Soil pH not only represents an easily determined feature of soil but is an easily managed agronomic parameter as well. Several plant nutrients become less available to plants at the extremes of pH values and other elements become available in toxic amounts [40]. Likewise, the bioavailability and plant uptake for Pb (free lead) can be accomplished by lowering soil pH. In this study, it was observed that root Pb accumulation increased as the soil pH was decreased (data not shown).The results of this study indicated that tall fescue can be an efficient Pb-accumulating plant when coupled with other phytoextraction strategies such as lower pH, and the use of a chelate. Chelates, however, may pose environmental risks and possible contamination of the groundwater if allowed to stay long in a polluted soil, a serious concern raised in many phytoextraction studies reviewed by Lasat [39]. It is therefore likely that further technical refinements are needed on chelate-assisted phytoextraction particularly the EDTA threshold requirements for efficiency of Pb uptake. Other engineering control measures will have to be provided to prevent leaching of soluble Pb into the ground water, thereby preventing a secondary source of Pb contamination [19]. Also, limiting the resident time of the chelate in the soil by applying it a few days before harvest lessens the mobility of bioavailable metals that can potentially migrate and serve as sources of secondary pollution to the ground water [41].Root (A) and shoot (B) Pb concentrations of tall fescue grown at various levels of Pb and chelates. Means with a common letter do not differ significantly using Fisher’s Protected LSD test (P=0.05). An error bar indicates the standard error of the mean of 4 replications. (* indicates that an aqueous solution of acetic acid was added at the same time as the aqueous solution of EDTA).Effects of various concentrations of Pb and chelates on root and shoot dry biomass of tall fescue. For each organ, means with a similar letter do not differ significantly using Fisher’s Protected LSD test (p ≤0.05).indicates that an aqueous solution of acetic acid (5 mmol/kg dry soil) was added at the same time as the aqueous solution of EDTA; (SEM= standard error of the mean of 4 replications).Effects of various concentrations of Pb and chelates on per cent translocation indices of tall fescue. Means with a similar letter do not differ significantly using Fisher’s Protected LSD test (p ≤0.05).indicates that an aqueous solution of acetic acid (5mmol/kg dry soil) was added at the same time as the aqueous solution of EDTA; (SEM= standard error of the mean of 4 replications).This research was made possible through support provided by NASA to Jackson State University through The University of Mississippi under the terms of Grant No. NGT5-40098.
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+ The socio-economic trends and history of Central Mississippi reveal a major rural influence based upon a dependence on agricultural activities as part of the economic engine driving the state’s economy. Yet, in the last several years, the amount of agricultural land in the counties continues to decline. Similar changes in other variables associated with agricultural land use and the continuity of farming in the state have also been changing. Indeed, under the pressure of urban growth, some farmers are forced to use less productive soils or have abandoned the agricultural business. Considering the gravity of the problem and the implications for sustainable development, public concern has increased in the state of Mississippi that urbanization and other factors may be eroding potential farmland. Given the effects of the current trends on the future capacity to produce food items, there are concerns that the growing incidence of farmland loss may also erode the basis for sustainable use of agricultural land, biodiversity and protection of the state’s ecological treasures. Notwithstanding the gravity of these trends, no major effort in the literature has aimed at documenting the incidence of agricultural land loss and the linkages to urbanization in the region of Central Mississippi. What changes have taken place in the size of agricultural land within the counties and what factors are responsible for it? This paper examines the issue of farmland loss in Central Mississippi with a focus at the county level between 1987 and 2002 from a temporal-spatial perspective. In terms of methodology, the paper uses a mixed scale approach based upon the existing literature. Data were drawn from the United States Census databases of Population and Agriculture. This information is analyzed with basic descriptive statistics and GIS with particular attention to the spatial trends at the county level. Results indicate that the counties under consideration have experienced considerable change in the amount of agricultural land and other variables associated with the use of farmland, due to urbanization. With the types of changes occurring, instituting effective policies anchored in sustainability, community participation, and growth management will go a long way in addressing the situation. Other strategies for farmland protection based upon land information inventory and mapping in the region, are also recommended. The paper stands as an update of the existing literature and offers a valuable tool for decision makers within the domain of natural resources management.Agriculture has played a significant role in Mississippi’s history. During the early periods of its settlement, the inhabitants relied upon the abundant natural resources for food and shelter. Even in the modern era, agriculture remains the number one industry, with enormous employment opportunities for the citizenry in 82 counties of the state. In 1999, it was estimated that one in five employees in the state held a job related to agriculture [1]. The sector employs 30 percent of the state’s workforce either directly or indirectly and generates over $5.6 billion in revenues annually. The state’s farmland stretches through an area covering 11 million acres [2–4].Accordingly, the socio-economic trends and history of Central Mississippi reveal a major rural influence due to a dependence on agricultural activities as part of the economic engine driving the state’s economy [5]. Yet, the amount of agricultural land in the counties continues to decline. Similar changes in a host of other variables associated with agricultural land use and the continuity of farming in the state have occurred. In 1959, Mississippi contained 18,600,000 acres of farmland [6]. That number had declined to 10,600,000 by 2000. Similarly, the total number of farms in Mississippi has decreased by 11,000 since 1982. In 1982, 11 counties in the state had at least 751 farms each; by 1997, only two counties showed that many. Also in 1982, eight counties contained 275,000 acres or more, by 1997 only five counties did. The gravity of such losses in 80 of the 82 counties led to a total decrease of over 2 million acres from 1982 to 1997 [7].Central Mississippi and other regions in the state have also experienced rapid population growth and have expanded into rural areas to accommodate this growth. This urbanization of rural areas triggers changes that often alter the environmental amenities that urban dwellers were seeking when they migrated into the countryside. The scenic appeal and quality of natural resources in rural areas have been important factors in bringing population growth to the countryside. As urban growth expands into rural areas, the land base changes. One critical impact on the natural resource base is the conversion of agricultural land to urban uses. This change often engenders the reduction of aesthetic and ecological values of natural areas [8]. The growing incidence of agricultural land loss partly attributable to urbanization also poses an enormous threat to preservation of agricultural land in the state [9]. From 1992 to 1977, urbanized acreage rose by 196,900 acres. Much of this lost acreage came from pasture (37,500 acres) and soybean production (13,100 acres). Cropland under irrigation in 1992 and developed by 1997 totaled 4,500 acres, most of which produced cotton (2000 acres). Given the effects of these trends on the future capacity to produce food items, there are concerns that farmland loss may also erode the basis for sustainable use of agricultural land, biodiversity and the protection of the state’s ecological treasures [10].Notwithstanding the situation in Mississippi, the phenomenon of sprawling urban development stands as one of the key factors driving land use and land cover changes in the United States. The US Department of Agriculture’s Natural Resources Conservation Service estimates that over 12 million acres of land were converted to developed land in the United States during the period between 1982 and 1997. During that period, farmland accounted for over 50 percent of newly developed land, while another third came from forestland [11]. This intensification in urban land development at the expense of open space and natural lands has sparked a growing debate over the problems and benefits of urban development and sprawl.Given that the literature on urban development is so broadly dispersed, it is quite difficult to limit the definition and conceptual analysis to a single domain. As a result, several definitions for sprawl have been coined that describe sprawl as a specific form of urban development with low-density, dispersed, auto-dependent, and environmentally and socially-impacting characteristics [12, 13]. The negative externalities emanating from urban sprawl have been widely documented [14–17]. Other scholars have identified benefits of sprawl-style development [18, 19]. Of particular concern is the extent of land consumption and the inefficient nature of this type of growth and the increasing amount of critical land resources lost in relation to human population growth [20–23]. However, from the standpoint of research and management, there is still a great need to further our understanding of spatial and temporal patterns of urban land use.To keep up with these changes, agencies from all levels of government and the private sector devote substantial resources to obtaining spatial information systems to study the impacts of urban infrastructure on agricultural land [24]. Notwithstanding the gravity of the trends, no major research effort has aimed at documenting the incidence of agricultural land loss due to urban development in Central Mississippi through the use of GIS. The crucial question remains, what changes have taken place in the amount and distribution of agricultural land within individual counties and what factors are responsible for it? If patterns of land loss can be determined through this technique, then future urban land development and conversion could be better predicted and better judgments could be made in developing land use policies and strategies restricting land use [25].This project examines farmland loss in Central Mississippi from a temporal-spatial perspective with a focus at the county level, between 1987 and 2002. Alternate strategies for farmland protection, based upon growth management, land information inventory and mapping, as well as community participation in the region, are also recommended. This paper contains five sections. Section 1 offers a description of the methodology and the study area. Section 2 presents the results and data analysis, while section 3 discusses the findings and their significance to land management. The fourth section offers recommendations for change in land-use policy. The final section summarizes the importance of the study to the future of agricultural productivity in Mississippi and elsewhere. To analyze the trend, the project adopts a time series approach, descriptive statistics, regression analysis and Geographic Information System (GIS) mapping of socioeconomic and land data from the United States Census. This paper has three objectives. The prime objective is to update the existing literature. A second objective is to provide a useful tool for decision makers within the domain of natural resources management. The third objective is to show how the latest advances in GIS can be used to enhance land management at the county level.The study area (Figure 1) consists of the Central Mississippi Planning and Development District (CMPD, hereinafter called “Central Mississippi”). The District contains the Jackson Metropolitan Statistical Area, which encompasses three counties (Hinds, Madison, and Rankin) and four adjoining rural counties with sizeable natural areas (Simpson, Yazoo, Warren, and Copiah). The Central Mississippi Planning and Development District is situated within two major river basins, those of the Pearl and Yazoo rivers. The population of the areas adjacent to the river basins is estimated to be over 1.5 million [26, 27]. The total size of the study area is 5,233 square miles, which encompasses a diverse landscape, a wide range of economic activities, and extensive areas of land suitable for agriculture and forestry [28, 29]. According to the 1997 Census of Agriculture, the study area contained a total of 1,120,307 acres of agricultural land [30].Environmental features within the region include rangelands, sensitive wetlands, and streams that support an abundance of fish and other wildlife. These environmental systems drain into the major river basins. Other notable features of the study area include ground water aquifers serving the needs of the counties as well as downstream communities. Also located in study area is the Ross Barnett Reservoir, an impoundment of some 33,000 acres, located North of Jackson and stretched across a distance of 43 miles. The Jackson Metropolitan Area draws 75 million gallons of groundwater annually. Worries about the impacts of projected growth on the area’s water resources, have led to water management emerging as a high priority among local government officials in the three urban counties (Hinds, Madison and Rankin) [31, 32].The study area has a large concentration of high-quality farmlands that are rated under the prime soil classes due to their capability for agriculture. Since the area was first settled, agriculture has played a vital role in attracting residents and investment. Today, Central Mississippi is home to a diverse range of agricultural operations. The climate supports a variety of agricultural uses, such as production of corn, cotton, soybeans, rice and crops, livestock husbandry and poultry production. Forestry is gradually emerging as a popular land use in Central Mississippi. The study area also contains extensive acreage of forested timberland. Some of the products include pine and hardwood sawlogs and hardwood-pulpwood cords. Rankin and Copiah counties rank among the most heavily forested in state, accounting for a combined total of 730,500 acres of forested land [33, 34]. However, intensive harvesting throughout the seven counties has caused some negative impacts on the forest ecosystem and biodiversity.The well-known impacts of urbanization have been manifest in the area by a rise in population and concomitant increases in building permits issued, housing construction, and other indicators. Proliferation of pollution-intensive activities prompted in part by mining industry and the presence of 1,070 dumpsites, has raised environmental concerns due to the widespread discharge of mercury, pathogens, and PCBs into water systems. Public managers in the adjoining rural counties must also grapple with the impacts of urban sprawl and various sources of pollution. As a result, the CMPD stands as an ecosystem under stress [35, 36].Such a diversified socioeconomic and environmental profile, built around intense land use and the extraction of natural resources, has substantial implications for the stability of area ecosystems and future use of agricultural land. It is clear that conditions in the Central Mississippi Planning and Development District deserve consideration as an ideal place to study GIS applications in land management. The presence of flourishing agricultural operations and the strength of other sectors, combined with important socio-economic indicators, have led and will continue to lead to changes in the area to both agricultural lands and ecosystems.This paper uses a mixed-scale approach based on government databases. The spatial information for the research was obtained from the Mississippi Automated Resource Information System office in Jackson, Mississippi, the American Farmland Trust and United States Census of Agriculture for 1987, 1992, 1997 and 2002. Federal geographic identifier codes for the seven counties (Copiah, Hinds, Rankin, Madison Simpson, Warren, and Yazoo) were used to geo-code the information contained in the data sets. The spatial data came from land-use capability and classification maps for the study area. This information was analyzed with basic descriptive statistics, regression analysis, and GIS, with particular attention to the temporal-spatial trends at the county level. The relevant procedures consisted of two stages, as described below.The initial step in this research involves the identification of the variables required to analyze changes at the county level from 1987 to 2002. The variables consist of socioeconomic and environmental information, including amount of agricultural land, average size of farms, market value of land, value of machinery, amount of cropland, number of housing permits, population and selected indicators on housing (homeownership rate, income and unit structure)(See Tables 1 through 4). Appropriate variables were derived from secondary sources such as government documents, newsletters and previous works. That process was followed by the design of data matrices for socioeconomic and land use (environmental) variables covering the census periods from 1987 to 2002. The design of spatial data for the GIS analysis required the delineation of city boundary lines within the study area as well. Given that the official boundary lines between the seven counties remained the same, a common geographic identifier code was assigned to each of the areal units to ensure analytical coherency.In the second stage, descriptive statistics and regression analysis were employed to transform the original socioeconomic and land-use data into relative measures (percentages, ratios and rates). This process generated the parameters for establishing, the extent of change or land loss for each of the seven counties facilitating gradual measurement and comparison of the trends in the area overtime. This approach allows detection of levels of change, while the graphics highlight the land-loss trends affecting the study area. The remaining steps involve spatial analysis and output (maps-tables-text) covering the study period, using ARCVIEW. The spatial units of analysis consisted of the seven counties (Figure 1). The study area map indicates boundary limits of the county units and their geographic identification codes. Outputs for each county were mapped and compared across time. This process helped show the spatial evolution of farmland loss, as well as changes in other variables.This section presents the results of the data analysis by first providing a brief synthesis of the descriptive statistics and a regression analysis of the trends. Later, it highlights the spatial factors associated with change in agricultural land in the study area.Tables 1a–1c summarize data on loss of acreages from 1987 to 2002. Between 1987 and 1992, the area of farmland in Central Mississippi declined from 1,334,664 acres to 1,247,314 acres. This number fell further between 1992 and 1997 to 1,120,307 (Table 1a). The seven counties posted a combined total loss of 87,350 acres of arable farmland between 1987 and 1992 (Table 1b). This continued with losses 214,357 acres from 1987 to 1997 and 127,007 acres between 1992 and 1997. This loss was followed with minor gain of 3,362 acres by 2002. Table 1c shows that farmland changes stayed negative most of the time. Hinds County alone suffered double-digit declines of 13.1, 26.1, and 14.9 percent during the intercensal periods 1987–1992, 1987–1997, 1992–1997 respectively. The rural counties of Copiah and Yazoo also recorded double-digit percentage losses. Land loss in Yazoo County was also steadily negative during the census periods between 1987 and 1992, 1987–1997 and 1992–1997 and 1987–2000.The study area as a whole also posted similar levels of declines at a rate of 6.54–16.1 and 10.2 percent during the same period. The two other urban counties of Madison and Rankin recorded declining rates of less than −10 percent from 1987 to 1992. Within this period the land in farms in the two counties showed a sizable decline of 10.2 for Rankin and 16.5 for Madison. The rural counties of the study area (Warren and Simpson) recorded some slight gains. Tables 1a and b show gains of 6.69 percent (7,156 acres) for Warren County between 1987 and 1992 and 1.48 percent for Simpson. In that same period, the rural counties of Copiah lost about 5, 932 acres at a rate of 5 percent, while Simpson experienced a decline of 2,825 acres or-3 percent. Each county also witnessed some losses in the average size of farms, the number of farms and acreage of cropland between 1987 and 2002 (Tables 2a–2g).The statistical summary of the regression test is presented in Tables 2a–2g. The technique serves as a predictive tool that enables a numerical description of the way one variable relates to another. The correlation between two variables reflects the degree to which the variables are related. It ranges from +1 to −1. A correlation of +1 means that there is a perfect positive linear relationship between the variables. When computed in a sample, it is designated by the letter r. Among the individual counties, simple positive correlation of great significance was shown to exist between some of the variables. The predictive component of the test shows the 2007 estimates of the counties as the only ones that are part of a trend significantly different from the mere average of the historical data having the “s” suffix. All other estimates are not significantly different from the historical average. The parameter of the regression line is determined by the formula A and B in the expression Y = A + Bx Year, where Y is the variable on which the regression is done. The estimate for the year 2007 is just A + Bx2007. In the last column to the extreme right appear the Rate figures, the annualised percentage increase or decrease in the value of each variable using the 2007 estimate to generate the slope B. The rate figures on average size, cropland acres, and land value all suggest that the counties listed above contain or are close to a sprawling urban development. In nearly all the counties, cropland is declining, land values are increasing, and farm size is declining. The decline in farm size must be explained by some counteracting causes, such as urban sprawl and the tendency for counties to expropriate land for urban development (Tables 2a–2g).The spatial pattern of the land loss identified in the statistical analysis was put into focus by mapping the trends in ARCVIEW. In Figures 2.1–2.3, the spatial patterns of change in agricultural land have been differentiated in red and green, where the red indicates land loss and green land gain. Figures 2.1–2.3 display the spatial distribution of losses and gains for 1987–2002, 1987–1992, and 1987 to 1997 respectively. During the 1987 to 2002 period, agricultural land decline was visible in just two counties in the study area. Between 1987 and 1992 five out of seven counties experienced land loss. In the other periods (1987 to 1997), six counties experiencing land loss were dispersed around part of the study area. A cluster of five counties that accounted for gains were fully concentrated in the South East and South West section of Central Mississippi during the periods of 1987–2002 than the other years. These maps reveal a gradual change in agricultural land use across time and space.Mapping also shows that the area has witnessed some notable growth in population (Figures 3.1. to 3.2). Between 1990 and 2000, the overall population of the area went from 520,327 to 574,990, an increase of 10.5 percent. Of all the counties in the area, the two urban counties in the North Eastern portion of the district, Madison and Rankin, posted the largest population gains in the region. During the period 1990–2000, the population of Madison grew from 53,794 in 1990 to 74,674 or 38.8 percent [37]. In a similar vein, the population of Rankin rose from 87,161 to 115,237, or 32.3 percent. The third urban county, Hinds saw a meagre population decline (1.4 percent) over this period. Trends in the rural counties of Yazoo, Warren, Copiah and Simpson reveal substantial rises in population during the same period. This growth prompted increased housing indicators, such as number of housing units built, building permits, number of households and rate of home ownership with impacts on farmland (Figures 3.1 to 3.2; Table 3) [38].The results indicate that the counties under consideration have experienced considerable changes in the size of agricultural land and host of other variables associated with the use of farmland due to urbanization. The nature and extent of this change transcend all spatial units, regardless of their designation as urban or rural. Another important point to note is Hinds’ County’s status as the county that contains the Metropolitan Statistical Area’s central city of Jackson. To a great extent, the growth in urbanized land in Rankin and Madison is related to this. In spite of some minor gains attributable to best management practices, agricultural land loss has become a major land management challenge for planners in the area [39–41]. The pattern of demographic change reveals that population growth and sprawl are threatening the preservation of agricultural land. Additionally, the rapid growth rates of the area coincide with a growing disturbance in the surrounding natural ecosystems. Growth pressure is also evident in the widespread request for building permits to meet domestic needs for new homes.In light of these findings, it is evident that GIS stands as a valuable tool for decision makers and resource managers in gauging the problems posed by growth and development. Given the negligible effort to document the incidence of land loss due to linkages to urban development in Mississippi, this study not only fills that void, but also it fills an important gap in the literature. The temporal and spatial display of information pertaining to demographic and socioeconomic indicators of growth, and their potential impacts on land use, offers the decision makers the opportunity to craft response mechanisms to dealing with the problems created by urban development [42]. GIS also offers county managers an appropriate tool for tracking the status of lands with high resource values and protecting them from development. Such information is essential in shaping the contours of Smart Growth policies and enabling local governments in Central Mississippi District to prepare plans for effective land uses.Four recommendations for land use policy and growth management are offered below.Given the degree of pressure due to urban sprawl that is being placed on the state’s agricultural lands, more effective land use policies based on best management practices are needed. Local land use policies should flow from the explicit statements of objectives in which the decision makers delineate goals that can guide production of appropriate planning document and serve as guidance for farmers and developers in addition to decision makers. The existing plans should also contain a set of activities to accomplish the objectives in accordance with available resources as well suitable mechanisms for plan implementation and review. The plan, once in place, should then be followed unless circumstances justify changes. Such plans not only have the potential to improve the management practices of land users, but also to contribute to good land management and sustainable production. In addition, the state of Mississippi should designate areas facing severe farmland loss as special districts for farmland protection programs, as incentives to landowners for farmland protection [43, 44].The various counties in the study area should continue to support active community involvement in land use decisions likely to impact on the future of agricultural land. Although the likelihood of disagreement is high due to differences in attitudes, land value assessment approaches and management practices, all stakeholders will benefit immensely from active involvement in the decisions made by planning agencies. Community participation should also serve as a forum for proactive dialogue on conservation among landowners, developers, and government for the purpose of fostering sustainability [45].Both the study area and the state of Mississippi as a whole take pride in being agricultural area where farmland accounts for a sizable proportion of income in dozens of counties. Yet, in the last two decades, the major components of agricultural productivity have come under intense pressure and have suffered from degradation and conversion to urban use. These impacts have eroded the total land area available for farming. The growing incidence of agricultural land loss and the future access to farmland for residents of the state have reached critical proportion that require the incorporation of sustainability principles into the current policy framework for land management. In the absence of such principles, the current pattern of land-loss will live the agricultural sector in the foreseeable future worse off. In light of the stakes involved in ensuring future productivity, the application of sustainability in land use policies is essential [46].The agricultural counties of Central Mississippi lack an integrated growth management strategy capable of providing greater predictability about where, when, and how much development will occur. Growth management should be applied to the high-growth urbanizing counties in order to protect farmland by channeling new development away from important agricultural areas. Growth management seeks to balance the benefits of development with the costs imposed on quality of life and requires up to date information on the environment. Yet little efforts have been made to collect land information or periodically map critical areas in Central Mississippi. Future growth management legislations should require local governments to identify lands with high natural resource, economic, and environmental values and protect them from development. Local governments should also be directed to make decisions in accordance with comprehensive plans that are consistent with protection of adjoining agricultural land areas. This approach would provide managers with valuable tools in addressing the challenges facing the agricultural sector. Such tools are critical to achieving the long-term economic and ecological needs of the population by helping predict the interactions between agricultural land use and development [47, 48].Several conclusions can be drawn from this study. First socioeconomic factors related to urbanization and developments have significantly altered the agricultural land base of the Central Mississippi Planning District. The analysis here demonstrates that notable losses of farmland acreage have occurred in the study area. These trends have persisted over the years with a significant spread across time and space. Moreover, in spite of small gains attributed to best management practices by some operators, it is probable that the amount of farmland will continue to drop in the study area, as well as across the rest of the state.GIS analysis has provided further insights into the spatial evolution of agricultural land use in the region. Mapping succinctly revealed the spatial patterns of declines and gains in land in farms among the seven counties. This spatial and temporal display of information pertaining to variations attributed to agricultural use and the potential impacts of socioeconomic factors on the use offer decision makers an opportunity to devise appropriate response mechanisms. It also enables them to formulate effective strategies for dealing with land loss in those locations deemed most vulnerable to growth pressures. This study demonstrates that using census data and statistical analysis coupled with GIS analysis can provide useful information for land management decision-making. It also updates the existing literature by offering badly needed empirical support for the incorporation of sustainability principles into land development policies and practices promulgated in Mississippi and it offers a viable tool for decision makers within the domain of natural resources management [49, 50].The Study AreaCounties with Increasing and Decreasing Farm Lands from 1987–2002Counties with Increasing and Decreasing Farm Lands from 1987–1992Changes in Farmland Acreage, From 1987–1997.Population by County 1990–2000Population Change by County 1990–2000Agricultural Land Acreage 1987–2002Change in Farm Acreage 1987–2000Percentage of Change in Farm Acreage 1987 2000Copiah County Regression AnalysisHind County Regression AnalysisRankin County Regression AnalysisWarren County Regression AnalysisMadison County Regression AnalysisYazoo County Regression AnalysisSimpson County Regression AnalysisSelected Indicators on Housing 1999–2002
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+ This paper presents an overall view of major sources that may lead to the pollution of the Tigris within Mosul city. A stretch exceeding 20kms in length is selected that represents the “sick” path of the river. Many sites along the studied stretch are likely to affect the river quality in some way or another. Samples from 40 sources sites are taken for quality analyses These sources – as huge as 400000 m3 a day – are characterized as (medium – strong) in composition. Such wastewaters with the pollutants they carry alter the river water quality rendering it unsuitable for beneficial uses. Such alterations – do leave –many negative consequences concerning human beings and aquatic life. It is found that domestic discharges are among the most important sources of pollution. Sanitary wastes are often discharged – untreated -into the Tigris. Other illegal practices such as in-house slaughtering add to the pollution as well. Industrial, tourist and institutional wastes put an additional burden on pollution of the river water quality. These wastes contain lead, chrome, and other heavy metals that may pose health risks. Wastewater treatment plants that exist in some sectors do not perform as they are expected. They need proper evaluation and rehabilitation. Eutrophication - a characteristic problem in lakes - finds an access to occur into the Tigris. This problem results from intensive use of detergents rich in nutrients (P&N compounds). In general, pollutants of different sources heavily affect the river water. Recovery and self purification of the river is estimated to occur at 40 km far from reference point. The paper concludes with the necessity of construction of a central treatment plant(s) or tackling the pollutants at their origin. The paper also stresses on importance of environmental education and awareness in order to combat pollution problems.The Tigris is considered the sole surface water resource in Mosul city/IRAQ. Its water is used for domestic, municipal, industrial, agricultural, and recreational purposes. Besides, the Tigris is thought to be the ultimate sink for all wastewater arising from above activities.The Tigris had been put under monitoring for years. Many papers and theses were published regarding its sources of pollution, pollutants concentration, degradation in water quality, and other aspects [1–5].This paper is another effort added to previous works trying to shed light on evaluating status of the river as it passes through a selected stretch within Mosul city. Such stretch includes the major point and non point sources of pollution that do affect river water quality and quantity.A detailed survey was made to figure out the activities that might contribute to the pollution of the Tigris within Mosul city (fig 1). More than (60) discharge sites were visited. Samples from (36) sites had been taken for quality characterization. Physical, chemical, and biological tests had been conducted on each adopting “the standard methods” (6, 7). Table 1 shows these tests.The Tigris - like any other water resource - is subjected to numerous sources of pollution. Municipal discharges, industrial pollutants, agricultural activities residuals, direct runoff, tourism, illegal practices, atmospheric pollution, and others are few examples. The most important sources covered for the purpose of this paper are listed in table (2).It is estimated that as huge as 400000 m3 of wastewater is daily discharged - untreated - into the river. This is equivalent to 17000 m3 /hr with a peak of 20000m3/hr at day hours. Domestic waste loads, on the one hand, add a great burden on the pollution of the Tigris. These wastes comprise foul wastes of more than 6000 dwellings and apartments lacking waste collection system. Considerable amounts of these wastes are directly or indirectly (via valleys) discharge their loads into the river (table 3)The field survey reveals that some of the pollution sources are direct point sources while some other sources are indirectly affecting the river water quality. As these sources differ, the nature of their pollution loads vary accordingly as shown in tables (4–6). The characteristics of discharged wastewater can be grouped into the medium-strong categories according to guidelines [8]. The tests also verify that a clear deterioration in water quality does occur. Concentration of various contaminants & compounds exceed the limits recommended by local and authorized agencies [9, 10].Odor, foam, color, death and migration of aquatic life, and dominance of anaerobic conditions can easily be detected near sewer outfalls. These adverse consequences as well as the increase of pollutant concentrations have lessen the river aesthetics, increase hardness, salinity, and rendering the water unfit for different beneficial uses.The changes in the physical characteristics of the river water such as temperature, turbidity, and suspended solids are clearly demonstrated in table 7. These changes are detected as the river passes along the stretch of the study area. In the north of the city there exists a large water impoundment. Some quarries and constructional mills are encountered. Animal breeding (buffalo, sheep and cows) are widespread. In the heart of the city there are more than 15 point sources discharging their loads into the river. It is strongly stressed that such activities will continue deteriorating the river water quality.One of the most important sources of pollution is the domestic discharges. Such wastes render the river water unfit for beneficial uses. Previous studies [11] revealed that the river water is no longer valid for swimming. The total bacterial count in the discharged wastewater is amounted as high as 2×104 – 2×107. These amounts exceed the recommended values [12]. This disorder is attributed to some illegal practices such as discharging toilet waste directly into the river or due to in-house slaughtering activities.Eutrophication, a phenomenon that largely takes place at lakes and slow moving water bodies has found an access to occur in the Tigris. This problem arises from the fact of using large amounts of nutrient-rich detergents. Throughout the past years, detergents were distributed to the families as a part of a monthly ration of oil for food program during sanction (1991-present). Such detergents contained high concentrations of phosphorous the main cause of eutrophication, (see table 8).Moreover, eutrophication can be detected by chlorophyll measurement which surpasses the guidelines of 2mg/l [13]. Eutrophication is known of its vast adverse effects, some of which are listed in table (9). The BOD5 values represent the organic pollution of the Tigris. Organic load leads to decline of dissolved Oxygen and release of ammonia and nitrite. This declination may extend for tens of kilometers.Table (10) illustrates BOD variations along the studied stretch. This table indicates that river water can be classified as poor-good in terms of quality as per authorized standards [14]. It does show that the Tigris starts recovering its health after 40km.On the other hand, the survey revealed the huge adverse impacts incurred by industry, tourism and health-care institutions. Most of local industries have no wastewater treatment plants. Wastes are directly discharged untreated into the Tigris. Most of existing wastewater treatment plants do not perform as expected. Moreover, such plants are secondary and incapable of removing nutrients (P&N compounds) as well as they poorly perform at shock loads occasions. Table (11) shows the performance of some plant covered by the survey.Heavy metals such as chrome, copper, and arsenic may have an access to reach the river. Some industries produce these elements in their processes like textiles industries, tanneries, etc. Al–Layla & Al-Rawi confirm this fact upon studying impact of textile wastewater discharges on the Tigris [15]. Lead concentrations may increase in the river water. This element arises from traffic and reaches the river from runoff or via atmosphere. The problem with heavy metal is that they are absorbed by particulates at normal pH levels causing very low dissolved traces and consequently its monitoring becomes very complicated [16].The Tigris river water quality shows a distinct deterioration within the studied stretch. Point and non-point sources of pollution are widespread along the selected stretch.Domestic wastewater discharges and illegal practices severely affect the Tigris water quality. This is reflected on an increase in organic & bacterial loads, and causing health risks.Industrial, tourism, medical institutions and other services add to the pollution of the river impeding self purification and rendering the water unfit for different uses.Reduced performance of waste treatment plants accompanied by the lack of specialized operators.The increased consumption of detergents increased occurrence of eutrophication.The Tigris starts recovering health and resumes an acceptable quality after 40 km from reference point.Absence of awareness of the pollution prevention measures and the non-existing of a real environmental monitoring authority increased pollution of the Tigris.Water and sewerage directorates should play an effective role in preventing unlawful connections to water networks and reducing hydraulic loads received by the river.Environmental authority should be given the power to enforce the law against illegal practices.All types of media should contribute to raise the public awareness of environmental protection as a holy and human task recommended by all religions (Islam, Christianity, etc).Pollution Sources of the Tigris River within Mosul cityThe Studied CharacteristicsMajor Contributor Sites & Characteristics to the Tigris PollutionResidential Units DistributionCharacteristics of Domestic WastesCharacteristics of Industrial Wastes• All units are in mg/l except pH. EC in micromhos/cm, TH, and Alkalinity in mg/l as CaCO3.• Values between brackets represent the mean valueCharacteristics of Valleys WastesMajor Human Activities Affecting the Physical Characteristics of the River+ Slight Increase; + + Clear Increase; + + + Severe Increase.Components of Used Detergents as a Part of Family Ration.Some Problems Caused by EutrophicationBOD5 Variation along the Studied StretchReference PointWastewater Treatment Plants Performance
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+ Land contamination is one of the widely addressed problems, which is gaining importance in many developed and developing countries. International efforts are actively envisaged to remediate contaminated sites as a response to adverse health effects. Popular conventional methodologies only transfer the phase of the contaminant involving cost intensive liabilities besides handling risk of the hazardous waste. Physico-chemical methods are effective for specific wastes, but are technically complex and lack public acceptance for land remediation. “Bioremediation”, is one of the emerging low-cost technologies that offer the possibility to destroy various contaminants using natural biological activities. Resultant non -toxic end products due to the microbial activity and insitu applicability of this technology is gaining huge public acceptance. In the present study, composting is demonstrated as a bioremediation methodology for the stabilization of contaminated lake sediments of Hyderabad, A.P, India. Lake sediment contaminated with organics is collected from two stratums – upper (0.25 m) and lower (0.5m) to set up as Pile I (Upper) and Pile II (Lower) in the laboratory. Lime as a pretreatment to the lake sediments is carried out to ensure metal precipitation. The pretreated sediment is then mixed with organic and inorganic fertilizers like cow dung, poultry manure, urea and super phosphate as initial seeding amendments. Bulking agents like sawdust and other micronutrients are provided. Continuous monitoring of process control parameters like pH, moisture content, electrical conductivity, total volatile solids and various forms of nitrogen were carried out during the entire course of the study. The stability of the compost was evaluated by assessing maturity indices like C/N, Cw (water soluble carbon), CNw (Cw/Nw), nitrification index (NH4/NO−3), Cation Exchange Capacity (CEC), germination index, humification ratio, compost mineralization index (ash content/oxidizable carbon), sorption capacity index (CEC/oxidizable carbon). Enzyme activities of agricultural interest like urease, phosphatase, β-glucosidase, dehydrogenase and BAA-hydrolyzing protease, which are involved in the nitrogen, phosphorus and carbon cycles, were also assessed. Total content of macro and micronutrients in the final compost was also determined to assess the fertilizer value. The studies revealed that composting could be applied as a remediation technology after removing the top sediment. The maturity indices that are evaluated from the present study can be used to validate the success of the remediation technology.Sediments are generally recognized to play a prominent role in ecosystem cycling, as the top layer is intimately linked to surface waters through physical, chemical and biological processes. The continuous discharge of industrial wastes into the water bodies is increasing the risk of their contamination [1]. Sediments are globally considered the ultimate sinks for the particulate accumulation of organics and their environmental transformation in sediments has been well documented [2, 3]. Organic contaminants in sediment systems are mixtures of hundreds of aliphatic, chlorinated aromatic and other organic compounds, the relative proportions of which vary greatly between sources [4]. Inorganic contaminants also have very complex interactions with both anthropogenic and natural components in marine sediments.The persistence of organo xenobiotics in the environment is a matter of significant public, scientific and regulatory concern because of the potential toxicity, mutagenicity, carcinogenicity and ability to bioconcentrate up the trophic ladder. These concerns continue to drive the need for the development and application of remediation techniques [5]. It is now widely recognized that contaminated land is a potential threat to human health, and its continual discovery over recent years has led international efforts to remediate many of these sites, either as a response to the risk of adverse health or environmental effects caused by contamination or to enable the site to be redeveloped for use [6]. A key factor in both the degradability and bioavailabiity of the most recalcitrant fractions of anthropogenic contaminants is the long-term sorption that can occur between organic molecules and clays or other minerals in soils and sediments [7].Hyderabad, the capital city of Andhra Pradesh, South India has almost 80 lakes in and around the city. With the growing industrial activities over the years about 8 large industrial estates have been developed and these industries are unscrupulously dumping their effluents into the nearby lakes thereby depleting the natural flora, fauna and the ecological balance. The pollutants in these lakes tend to bioconcentrate up the trophic ladder and reach humans. So, rejuvenation of these lakes is a task of utmost importance and employing physico-chemical processes only transform the pollutants from one form to another but biological processes transform them into innocuous end products. These concerns continue to drive the need for the development and application of viable and low cost remediation techniques [8]. Bioremediation is one such technology that offers the possibility to destroy or render harmless various contaminants using natural biological activity. Micro organisms have a unique ability to interact both chemically and physically with a huge range of both man-made and naturally occurring compounds leading to a structural change to, or the complete degradation of the target molecule [9].Composting is one of the bioremediation strategies which when carried out under controlled conditions in the presence of oxygen results in the biological decomposition and stabilization of the biodegradable components. The process of composting includes four main phases, which are the initial phase, the thermophilic phase, the mesophilic phase and the maturation phase after which the compost can be used as an organic amendment. For the compost to be used as an organic amendment it has to be assessed for certain parameters like nitrification index, cation exchange capacity, germination index, humification index, water soluble carbon, compost mineralization index, and sorption capacity index etc.Stabilization or maturation also implies the formation of some humic – like substances the degree of organic matter humification is generally accepted as a criterion of maturity [10]. The humification process produces functional groups and so increased oxidation of the organic matter leads to rise in cation exchange capacity. So compost with high cation exchange capacity is regarded as an index of maturity [11]. The degree of maturity can also be revealed by biological methods involving seed germination and root length [12]. Since immature composts may contain phytotoxic substances such as phenolic acids and volatile fatty acids [13].The aim of the present work is to monitor the process of composting for the contaminated lake sediments and observe the changes occurring in the two piles set up by taking sediment from different strata and discusses few of the maturity indices and thus validates their use as matured composts resulting from the biological stabilization.Initially a 100 × 100 m plot on the lakebed was chosen and the top layer was dredged and sent to a treatment storage and disposal facility located near the outskirts of city for land filling thereby getting rid of the uppermost-polluted layer on the lakebed. The next 0.25 m of the lakebed was removed and a portion of the sediment was set up as Pile 1. A further 0.25 m was dredged whose part was set up as Pile 2. The two piles were set up by taking the sediments from two different strata of the lakebed in a view to assess the leaching potential of pollutants into the different strata of the lake sediments and their amenability to composting. The soil type on the lakebed has the following characteristics Clay – 352 gm/Kg; Silt (20–50μm) – 81 gm/Kg; Sand (50–200μm) – 61gm/Kg; Sand (200 – 2000μm) – 32 gm/Kg.Pile 1 – Sediment from upper stratum:Pile 2 – Sediment from lower stratum.The polluted sediments from the two different strata were initially mixed with lime at 1% (w/w, dry weight basis), which raised the pH to 9.2. Liming helped in the stabilization of heavy metals by precipitation as metal hydroxides at higher pH [14]. The organics and heavy metal concentration in the sediments before and after composting are presented in Table 1. The sediments were kept as such for five days and then mixed with organic amendments like manure (Poultry manure and cow dung) and sawdust, which brought down the pH to 6.8. Lab scale experiments for composting were set up by taking sediment, manure and sawdust in the ratio of 2:1:2 making a total of 120kgs on dry weight basis. The lab scale set up for the aerobic composting pile is shown in Fig. 1. The organic manure is high in nitrogen and is used for the adjustment of carbon/nitrogen ratio. Urea and superphosphate were also added to maintain the initial C/N ratio in setting up of the pile. Saw dust is used as a bulking agent to increase the porosity of the mixture. The sediment was mixed with the organic amendment to maintain the total solids content between 35 – 45% and then homogenized. The aeration was given through natural ventilation and by turning over the piles at an interval of 7 days. The piles were protected with a layer of sawdust and straw on the surface to avoid odors and influence from wind.The composting process lasted for a period of 14 weeks, including four stages such as the initial phase, the thermophilic phase, end of thermophilic phase and the mesophilic phase. The sludge – conditioner mixture was placed on a bed of wire mesh, at an altitude of 30 cm. The material was spread on a layer of wood shavings, which was covered with a layer of sawdust and straw. The moisture content was initially adjusted to 55% and later there was no addition of water. Once the pile was set up, the decomposition of the organic matter by thermophillic microorganisms started, which elevated the temperature to 58° C, causing the destruction of the pathogens. The aerobic conditions were assured by aeration through turning. Thermophillic phase remained for a period of 50 days in the pile. Within a period of 14 weeks, complete stability of the compost and the removal of odors were assured. The mixtures were then kept for another 6 weeks for maturation. All the parameters were assessed for this matured compost. Sifting was done to matured compost to separate the conditioning material and to obtain a homogenous product. Samples from composting mixtures were analyzed once every 7 days. A representative sample was taken by picking up material from different points of the two piles.All the parameters such as moisture content, electrical conductivity, pH, organic matter, and nitrate nitrogen were estimated using standard APHA methods [15]. Total Kjeldahl nitrogen and ammonical nitrogen were analysed using Kjeldahl assembly (Kjel Plus DISTILL M KPS 020, India), the oxidizable carbon (Co) was determined by oxidimetric method the ash content by gravimetric method after burning off the dry mass at 550° C [16]. In the water extracts (compost/water ratio of 1:10) the water soluble forms of carbon (Cw) and nitrogen (Nw) were determined. Total phosphorous (ascorbic acid method) and cation exchange capacity (CEC) was determined according to the method described by Gupta [17].Humic acids, Fulvic acids and non humic fraction were estimated by the method as described by Hsu and Lo [18]. Humification index (HI), i.e., the ratio between Humic acids and fulvic acids is deduced from the equation:20 gms of compost was weighed and extracted with 200ml of deionized water by shaking for 24 hrs. The extracts were centrifuged at 10,000 RPM for 25 mins and filtered through 0.45-μm filter membranes (mdi, India). Water extracts are immediately analysed for organic carbon [18].Assays of hydrolases (β-Glucosidase, BAA-Hydrolyzing Protease, Urease and Phosphatase) and dehydrogenase activities were performed as described by Garcia et al., [19, 20]. Dehydrogenase assay was based on the combination of two methods [21,22]. To determine β-Glucosidase activity 0.05 M 4-nitrophenyl-β-D- gluconopyranoside (PNG) was used as substrate [23] while,0.115 M p-nitrophenyl-phosphate (PNPP) was used as substrate to measure the phosphatase activity. The para-nitrophenol (PNP) produced by both hydrolases was extracted and determined spectrophotometrically at 398 nm. [24]. To determine the BAA-hydrolysing Protease and Urease activities, 0.03 M N-α-Benzoyl-L-Argininamide (BAA) and 6.4% urea, respectively were used as substrates. The ammonium released by the two hydrolytic reactions was measured by an ammonium selective electrode (ORION, Model.95–12). To determine dehydrogenase activity, 0.4% 2-p-iodophenyl-3-p-nitrophenyl-5-tetrazolium chloride (INT) was used as substrate. Iodonitrotetrazolium formazan (INTF) produced in the reduction of INT was measured spectrophotometrically at 490 nm.The absence of phyto-inhibitory substances that reflect maturity of compost was tested by seed germination. Germination tests were performed with garden cress (Lepidium sativum L.) Seeds were soaked in compost extracts in water (1: 10 w/v) for 48 h. Lepidium sativum seeds were used because of their rapid germination and sensitivity to phytotoxic compounds [25]. The germination index, inversely related to the presence of phytotoxic substances in compost, was calculated as the percentage of seeds germinated on filter paper in Petri dishes with 10 ml of compost extract multiplied by the average length of roots in mm expressed as percentage of a control with distilled water [26]. The percentages of relative seed germination, relative root elongation and germination index (GI) are calculated by the following formula:The physico chemical analysis of the manure and saw dust and sediments from two different strata used in the present study are presented in Table 2 and Table 3 respectively. A representative sample was taken by picking up material from different points of the two piles.The temperature variation during composting followed a pattern similar to many other composting systems [27, 28]. In Pile 1 the temperature rised only upto 46° because of the presence of highly toxic compounds and thus low microbial activity while in pile 2 there was a steep increase in temperature up to 58° due to the availability of comparatively more degradable organic matter and hence intense microbial activity. The thermophillic phase lasted for nearly 7 weeks. After this period of elevation the temperature gradually decreased to ambient levels and this marked the end of the thermophilic phase of composting. At this stage the decomposition rate stabilized with a consequent decrease in temperature and microbial activities.During composting the moisture content of the piles decreased which is due to the evaporation of water as a consequence of turning the piles and microbial heat generation. The continuous decrease in moisture content during composting is an indication of organic matter decomposition [29]. The moisture content in Pile 1 decreased up to 42% while that in Pile 2 decreased to 30%. The intense microbial activity and organic matter degradation during the first weeks of thermophilic phase led to the formation of ammonia as a consequence of ammonification of organic nitrogen. The solubilzation of ammonia led to the formation of ammonium and an increase in the pH values in the composting mixtures from an initial of 6.8 to 7.6 in pile 1 and from 6.7 to 8.2 in pile 2. Electrical conductivity increased due to the concentration of the salts because of degradation of organic matter [51]. Production of Nitrate-N also explains the increase in conductivity of the composting mixtures. This is important from an agricultural point of view since an increase in electrical conductivity is a direct consequence of the increased concentration of nutrients, such as nitrate.The changes in the C/N ratio reflect organic matter decomposition and stabilization during composting and are represented in Figure 2. In the initial stage of composting intense mineralization processes takes place which is manifested by a considerable decrease in carbon and increase in ash content in both the piles, the C/N ratio decreased due to the mineralization of the organic matter [30]. The total nitrogen content increased during composting from an initial of 1.6% to 1.75% in pile 1 and from 1.65% to 1.76% in Pile 2. These nitrogen increases are probably due to a concentration effect caused by the decrease of the substrate carbon resulting from CO2 lost [31] as a consequence of the degradation of non-nitrogenous organic matter (Carbohydrates etc.).The organic carbon concentration in Pile 2 degraded much more intensely during the thermophillic phase of composting due to greater activity of the micro organisms and the presence of easily degradable substances. After 14 weeks of composting the organic matter in Piles 1 and 2 decreased from 48.9% to 34% and from 48% to 21% respectively. The C/N ratio decreased due to the mineralization of the organic matter. In pile 1 the Carbon/Nitrogen decreased from an initial value of 30.5 to 19.4 at the end of the process while for the same period of time in pile 2 the initial carbon/nitrogen ratio decreased from 29.5 to 11.9. According to [32] a Carbon/Nitrogen in the range of 10–15 of the compost indicates a good degree of maturity.The compost mineralization index is expressed as Ash content/Oxidizable Carbon. The changes in Compost Mineralization Index are represented in Figure 5. In the initial stage of composting (about 7 weeks) the intense mineralization process takes place, which is manifested by a considerable decrease in carbon and increase in ash content. In Pile 1 the ash content increased from 45.1% to 52.8% and the oxidizable carbon decreased from 19.5% to 16.2% the compost mineralization index increased from 1.85 to 3.01. The ash content in Pile 2 increased from 49.4% to 61.9% and the oxidizable carbon decrease from 20.4% to 14.1% and the compost mineralization index increased from 2.5 to 4.65.The nitrification process has been used as maturity index of composting [33]. The changes in Nitrification Index expressed as NO3− -N/NH4+ -N during the composting process is presented in Figure 2. The NH4+ - N content in pile 2 increased from 658 mg/kg to 725 mg/kg during the thermophilic phase. This increase could be due to conversion of organic N to NH4+ -N via the ammonification process and then the NH4+ - N content decreased to 144 mg/kg towards the end of maturation phase. A similar trend was observed by [34, 35]. This decreasing trend guaranteed that ammonification was ending and can be used as a criterion of compost maturity [36]. The ammonia produced during the thermophilic phase is oxidized to NO3− -N and thus the concentration of ammonia decreases with the increase in NO3− -N. In Pile 1 the increase in ammonical nitrogen was not very significant and so was the decrease. The NO2 concentration was negligible in both the piles, indicating that aerobic conditions prevailed during the composting process.Appreciable amounts of NO3− - N could be observed in Pile 2 the values increased from 0.01% to 4.5%, which were reached after maturation. A NH4+/NO3− ratio in favour of the oxidized form is considered desirable for mature compost. In Pile 1 the NH4+/NO3− ratio was 0.28 while in Pile 2 the ratio was 0.03 towards the end of composting. At the end of the process the concentration of nitrates should be higher than that of ammonium indicating that the process has been prepared under adequate conditions of aeration [29]. Due to the presence of highly toxic organics in Pile 1 there was low mineralization of organic nitrogen which in turn resulted in low ammonia evolution whereas in Pile 2 low concentration of organics resulted I higher nitrogen mineralization and hence higher ammonia evolution A high concentration of NH4- N in compost indicates instability and according to Zucconi & Bertoldi [12] it should not exceed 0.04% in mature compost [33] established a limit a 0.16 as a ratio between ammonium nitrogen and nitric nitrogen as an index of maturity in composts.The water-soluble carbon variation with composting time is presented in Figure 3. Water–soluble organic carbon is the most readily biologically active compound in composts applied to soils. Water-soluble organic carbon level in Pile 1 gradually increased from 1.35% to 1.44% in the 7th week and then gradually decreased to 1.29% towards the end. While in Pile 2 the water-soluble organic carbon level gradually increased from 2.02% to 2.41% in the 7th week and then gradually decreased to 1.32% towards the maturation phase. As carbon component that are easily available to microbes organic and amino acids, proteins were degraded during the thermophillic stage of the decomposition, breakdown products were continuously released resulting in an increase in water-soluble carbon [18]. A decline in water-soluble carbon is often used as an indicator of compost maturity [37]. In matured compost most of the soluble organic carbon is present as humic substances, which are resistant to further decomposition, thus explaining its increased stability observed with time during composting. A limit of Cw < 1.7% can be used to reflect a good maturation degree. [33].The water-soluble carbon and nitrogen index is presented in Figure 3. Water-soluble organic carbon level in Pile 1 gradually decreased by about 4.4% while in Pile 2 the decrease was 34.6%. The easily biodegradable carbon components that are highly available to microbes were degraded during the thermophillic stage of the decomposition, and the breakdown products were continuously released resulting in an increase in water-soluble carbon [18]. During composting the total nitrogen content increases with simultaneous decrease of its solubility in water (Nw) and towards the maturation phase the amount of mineral water soluble forms of nitrogen increased. [16]. In Pile 1 the Nw decrease form 13.5% to 7.2% and in Pile 2 the decrease was form 13.5% to 5.0%. The CNw increased from 0.99 to 1,7 in Pile 1 while in Pile 2 the CNw increased from 1.4 to 2.6 at the end of composting period.Humic substances comprise the most important fraction of organic matter because of their unique properties, such as the capacity to interact with metal ions, the ability to buffer pH, and the ability to act as a potential source of nutrients for plants [18]. The relative contents of humic substances and non-humic substances during composting are presented in Figure 4. The quantities of Humic acids, Fulvic acids and non-humic fraction in composting mixture at various stages of the composting process represent the humification process.In Pile 1 total humic substances increased from 15% of organic matter to 25% of organic matter after 7 weeks, stabilizing at this value till the end of the process. The fulvic acid level gradually decreased from 6.1% of organic matter to 5.4% in mature compost. The humic acid level increased from 3.2% to 10.8% in the mature compost.In Pile 2 Total humic substances increased from 31% of organic matter to 52% of organic matter after 7 weeks, stabilizing at this value till the end of the process. The fulvic acid gradually decreased from 8.2% of organic matter to 5.4% in mature compost. The humic acids increased from 5.2 to 21.9% in the mature compost.The increasing level of humic acids during composting process represents the humification and maturity of compost. In general, fresh composts contain low levels of humic acids and higher levels of fulvic acids [10]. During composting humic acids increased, where as fulvic acids slightly decreased. The non humic fraction in Pile 1 increased rapidly from 14% to 17% but then did not decrease further due to low organic matter degradation. in Pile 2 non humic fraction increased rapidly from 18% to 30% of organic matter for the first 7 weeks of composting, and then decreased to 12% in the mature compost may be due to decomposition and humification of the break down products and presence of easily biodegradable organic matter during the maturation stage.The Humification index and Humification ratio of the two piles are presented in Figure 8. Humification index in Pile 1 remained steady at 0.6–0.7 for the first 4 weeks and increased sharply to 1.3 in the seventh week and slowly increased to a final value of 1.8 towards maturity of compost. Humification index in Pile 2 remained steady at 0.6–0.8 for the first 4 weeks and increased sharply to 1.8 in the seventh week and slowly increased to a final value of 4.0 towards maturity of compost. The changes in humification index reveal that Fulvic fraction and non humic fraction extracted from sediment contain relatively high levels of biodegradable organic matter that was mostly decomposed during first 7 weeks of composting. The humification ratio in Pile 1 increased from 0.29 to 0.94 while in Pile 2 it increased from 0.4 to 1.04.One of the variables that are frequently determined to estimate the degree of transformation reached by compost during the process of composting is cation exchange capacity (CEC). The changes in cation exchange capacity are presented in Figure 9. Its determination in an organic amendment is of great value because it allows us to know the stability degree of the amendment. Several studies accomplished with different kinds of compost have demonstrated that CEC increases with the stability degree of the compost. On the other hand, this parameter gives an indication of the amendment’s capacity for catching nutrients and immobilizing phytotoxic substances as well as for buffering unforeseen pH changes. The obtained results showed that the CEC increased from an initial value of 31cmol/kg−1 to final values of 38.9cmol kg−1 in Pile 1 while in Pile 2 the value increased from an initial value of 55.3 to 68.6 cmol kg−1. The value in the Pile 2 was higher than the minimum recommended for mature compost (67 cmol kg−1) thus indicating that the compost is mature [11, 52].The Cation Exchange Capacity/Corganic ratio reflects the degree of maturity of specific humic compounds and according to Inbar et al., [38] it can be connected with the increase of the functional groups during humification process. The cation exchange capacity increases and organic carbon decreases with composting. The changes in Sorption Capacity Index are represented in Figure 5. At the end of the active phase Pile 1 had a sorption capacity index of 1.1 while Pile 2 had a value of 3.2 which was greater than 1.7 the lowest limit for describing well humified manures [11].The GI values increased during the composting process due to decomposition of the phytotoxic organic compounds. The change in germination index is presented in Figure 9. These phytotoxic compounds, which were present in raw waste or produced during the first days of composting as intermediate products of microbial metabolism, were degraded during the process, giving mature composts, which could safely be used with plants. The sample taken from the Piles 1 had GI a of 49 while Pile 2 had a GI of 95 which is greater than 80 and according to Zucconi et al., [25] & Tiquia et al., [39] indicates a phytotoxic-free compost.The degradation of the labile substrates contained in organic matter can be followed by studying specific hydrolases, which are relatively easy to determine, and specific to the substrate. The hydrolases monitored in the present work (BAA-Hydrolysing Protease, Urease, β-glucosidase, Dehydrogenase and Phosphatase could represent a good index of qualitative fluctuations of substrate during composting since they are substrate-inducible enzymes. The changes in enzymatic activities in Pile 1 and 2 are shown in Fig. 6. The high initial activity of these enzymes reflected the high microbial activity. The presence of a high content of degradable compounds in the pile 2 might have stimulated enzyme synthesis. As substrate decreased, the enzyme activity decreased as well [40].The β-glucosidase activity decreased throughout the composting process, in Pile 2 as may be expected, since carbonated structures are degraded as composting proceeded and only the most resistant and those with the smallest number of side chains remain at the end of the process. β-Glucosidase and BAA-hydrolyzing protease which are enzymes involved in C and N cycles, respectively showed a sharp decrease during the first 7 weeks, and then stabilized as a consequence of decrease in available organic substrates [41].Urease activity is closely related with the nitrogen cycle and it is involved in the hydrolysis of proteins to ammonium hydrolyzing urea-type substrates. It is believed that denaturalization of the enzyme during composting due to high temperature does not occur since urease is stable up to 80° C [42]. There is a pronounced difference between the values of the initial and composted samples. This is perhaps because the enzyme depends on microbial biomass, which implies that when the biomass is degraded (due to composting) enzymatic activity decreases. The activity of urease which catalyses the hydrolysis of urea to CO2 and NH4+, increased during the first 4 weeks of experiment probably as a consequence of diminution of high initial concentration of NH4+ in the substrate which may be responsible for the inhibition of this activity [43]. Subsequently urease activity decreased until week 7 and then remained more or less stable till the terminal phase of the composting process.Phosphatase is a key enzyme in the phosphorus cycle, which is induced by the substrate. Its activity is largely dependent on microbial biomass [40]. Phosphatase activity showed a sharp increase during the thermophilic phase followed by a gradual decrease. The increase observed during the thermophillic phase can be associated to high microbial activity, which occurs during this stage. The phosphatase hydrolyses compounds of organic phosphorous and transforms them into different forms of inorganic phosphorous [44]. Thus, the decrease in activity observed might be due to enzyme inhibition by inorganic phosphorous temporarily released from the mineralization of the organic phosphorous [45]. Phosphatases are enzymes with relatively broad specificity capable of hydrolyzing various organic phosphate esters [46]. The high initial activity could be related to the presence of organic phosphate compounds which may act as inducers of enzyme synthesis after a slight decrease during the first week phosphate activity stabilized at about 200 μmol PNP g−1 h−1 in Pile 2 and at 260 μmol PNP g−1 h−1 in Pile 1 [19].Dehydrogenase activity in soils and other biological systems has been used as a measure of the overall microbial activity [47] since it is an intracellular enzyme related to the oxidative phosphorylation process [48]. The initial high dehydrogenase activity recorded might have been the result of high microbial activity due to the high water soluble carbon concentration after two weeks dehydrogenase activity decreased until the end of composting. The dehydrogenase activity is represented in Figure 7.A number of physical, chemical and biological indices have been linked to the maturity of composts [49]. The compost in Pile 2 was alkaline, pleasantly earthy in smell and dark down. Compost characteristics are presented in Table 2. From the results obtained, compost in Pile 2 showed a C/N ratio of 11.9, which is in accordance with the recommended value between 10 and 15 [32]. This indicates that there is a considerable decrease in organic matter, water-soluble carbon and nitrogen. A total N content in composts is higher than the minimum level recommended (0.6% total N) by Zucconi and De Bertoldi [12]. Enzymatic activities observed in the present study are intracellular activities of proliferating microorganisms and not due to the activities of extra cellular enzymes. Hence hydrolytic and dehydrogenase activities are sensitive indicators of the state and evolution of the organic matter and the overall quality of the compost. The macronutrient contents were above the minimum recommended (0.5% P2O5; 0.3% K2O; 0.3% K2O; 0.3% MgO; 2.0% CaO) by Zucconi and De Bertoldi [at high pH conditions [50]. Heavy metals present, in 12]. Metal mobility and availability was reduced by increased pH values since cationic ions are less available all the composts, were below the maximum values permitted for each element. Further studies on the regulatory standards of the composted product can validate the applicability of these stabilized sediments for successful application as biosolids.The tentative cost of the treatment is calculated for the entire lake area of 0.5 Km2 from the studies carried out in lab scale. The lakebed is more polluted at certain quartets and towards the inlet and outlet culverts. From our study only 50% of the lake area needs dredging to a depth of 0.25m and the volume of the total sediment to be dredged will be 31,250 m3. The estimate cost for dredging and transportation of the polluted sediment to TSDF is Rs7, 00,00,000/-(US$16,00,000). The cost for ploughing the entire lake area to a depth of 0.6m is Rs 2,00,00,000/- (US$4,50,000). The total cost of rejuvenation of the lake by liming, composting and further biological activity is calculated to be Rs 1,00,00,000/- (US$2,30,000). Therefore the approximate cost calculated from our studies Rs 10,00,00,000/-(US$ 22,80,000). The cost of the treatment calculated is an extrapolation from our studies carried out under lab condition and is subjected to vary with the nature of the pollutants.Remediation of contaminated lake sediments using bioremediation can be a cost effective solution. The present study envisaged on the applicability of this technology for the entire lake insitu by carrying out lab scale studies of different stratum. The study revealed that the after the elimination of the top layer for physicochemical treatment the lower layers showed decreasing levels of concentration with depth indicating the slow transport rate of the pollutant. Aerobic composting, a bioremediation technology adopted in the present study was found to be very effective for the bottom layers and was significant for the top layers. Hence from our study it is concluded that my mixing the entire lakebed with a conventional farming technique called as “Ploughing”, and pretreatment with lime can render the stabilization of metal pollutants. However, liming may not completely stabilize metals in the case of lake beds contaminated with higher metal concentrations. Thus the stabilization of these lakes has to be addressed on pollutant specific scenarios with integration of physico-chemical and biological stabilization methodologies. The success of the composting of lake sediments was achieved by optimizing the C/N ratio of the piles to 30:1, moisture content of 55%. Further addition of bulking agents and organic amendments in the ratio of 2:1:1 (Sediment: sawdust: manure). The success of the composting maturity was assessed by evaluating indices C/N, Cw (water soluble carbon), CNw (Cw/Nw), nitrification index (NH4/NO−3), Cation Exchange Capacity (CEC), germination index, humification ratio, compost mineralization index (ash content/oxidizable carbon), sorption capacity index (CEC/oxidizable carbon). Enzyme activities of agricultural interest like urease, phosphatase, β-glucosidase, dehydrogenase and BAA-hydrolyzing protease, which are involved in the nitrogen, phosphorus and carbon cycles, were also assessed. Total content of macro and micronutrients in the final compost was also determined to assess the fertilizer value. The indices are compared with the literature values and can be concluded that the maturity indices are useful indicators for the success of the composting technology. The treat ability of the contaminated sediment can be evaluated using these indices and on standardizing these indicators can be successfully applicable for prototype application. The study also made an attempt to work out the cost of this treatment option. Thus it can be concluded that treatment of contaminated lake sediments should be addressed as case-by-case scenario. Bioremediation coupled with other physico-chemical and emerging technologies can be an effective insitu solution in handling the remediation of the natural water bodies and bring them back to life to support other dependent systemsSetup for turning mode of compostingVariation of Nitrification Index and C/N Ratio during CompostingVariation of water soluble carbon and CNw during composting.Variation of humic and non humic substances during composting.Variation of Scorpion Capacity Index and Compost Mineralization Index during composting.Activities of various enzymes during composting.Variation of dehydrogenase activity during composting.Variation of Humification Index & Humification Ratio during composting.Variation of Citation Exchange Capacity and Germination Index during composting.Heavy metal and organics concentrations before and after compostingNote: All the values are expressed as mg/KgCharacteristics of amendments usedCompost Product CharacteristicsAll parameters are expressed as % except pH and C/N
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1
+ Plants are good environmental sensors of the soil conditions in which they are growing. They also respond directly to the state of air. The tops of plants are collectors of air pollutants, and their chemical composition may be a good indicator for contaminated-areas when it is assessed against background values obtained for unpolluted vegetation. Both, aquatic and terrestrial plants are known to bioaccumulate heavy metals and therefore represent a potential source of these contaminants to the human food chain. An evaluation of heavy metals was conducted from vegetation samples collected at the Atlantic Fleet Weapons Training Facilities (AFWTF) in Vieques, Puerto Rico. In order to understand the potential risks associated to heavy metal mobilization through biological systems, it is first necessary to establish background values obtained from reference locations. This information allows a better interpretation of the significance of anthropogenic factors in changing trace elements status in soil and plants. Since Guánica State Forest is located at a similar geoclimatic zone as the AFWTF, samples at this site were used as a standard reference material and as experimental controls. Both sampling and analysis were conducted as previously described in standardized protocols using acid digestion of dry ashes. Then, levels of heavy metals were obtained by air-acetylene flame detection in an atomic absorption spectrophotometer. Our results from the samples taken at the AFWTF indicate mobilization of undesirable trace elements through the marine and terrestrial food web. Since plants naturally remove heavy metals from soils, they could be employed for the restoration of this and similarly contaminated sites.It is estimated that explosives and heavy metals usage have contaminated over a million cubic meters of soil at thousands of former military installations throughout the United States [14]. Major contaminants often include toxic elements such as lead, mercury, uranium and cadmium. Although trace elements occur naturally in soil, pollution with metals in bombing ranges is inevitable. Once in open environments, undesirable concentrations of pollutants could be dispersed through multiple biotic and abiotic processes. Either natural or anthropogenic in origin, soil fragmentation due to explosives will increase the surface/volume ratios of soil and thus geoavailability of metals, dust formation and erosion. The transport, residence time, and fate of pollutants in an ecosystem are serious social concerns.Plants are good environmental quality indicators and respond directly to air, soil or water quality [2, 3]. Since plants can naturally uptake pollutants from their local environment, their chemical composition can indicate degree of disturbance when assessed against background values obtained from unpolluted vegetation. Due to the high degree of plant endemism, the Caribbean has been included as one of the ‘hotspots’ for biodiversity [5]. Over 7,000 species of endemic plants have been described in this region alone, accounting for 2.3% of the total number of plant species on earth.A historically contaminated site with explosive and heavy metal is located at the eastern part of Vieques, Puerto Rico. Various types of explosive and non-explosive ordnance were used at the Atlantic Fleet Weapons Training Facility (AFWTF) until May of 2003 [12, 15]. This 900-acre zone was intensively used for aerial and naval bombardment. After forty years of bombing activity, this location harbors unique diversity of plants specially adapted to tolerate and perhaps attenuate this pollution. This region of Puerto Rico thus provides an excellent location for assessing plant diversity involved in heavy metal restoration.We evaluated the use of plants as natural indicators of the flux of specific trace elements in the marine and terrestrial environment. This knowledge might lead to the identification of critical pathways of pollutant’s transport in the ecosystem as well as the development of more cost effective means for the restoration of this or similarly contaminated sites.Vegetation samples were obtained at the former Atlantic Fleet Weapons Training Facility (AFWTF, 18°08.32N, 65°18.10W) in the island of Vieques, Puerto Rico (Fig. 1). Samples were also taken at Guánica State Forest (17°57.213N, 66°50.971W), a Biosphere Reserve designated by UNESCO in 1981. The analyses were done on Calotropis procera (giant milkweed), Panicum maximum (guinea grass), Ipomoea violacea, Acacia farnesiana, Sporobolus virginicus, S. pyramidatus and Syringodium filiforme (Manatee grass). In order to evaluate temporal variation, samples of S. filiforme were collected in 2001, 2003 and 2004.Samples were collected manually and identified by botanists at the Mayagüez Campus of the University of Puerto Rico. Each location was randomly sampled in a 2–5 m2 circular plot. Sampling sites were established independently for each species. Samples were composed of over 40 leaves picked alternately from upper, middle, and the lower foliar sections from 5–10 plants of each species. When possible, samples of root material, stems and fruits were collected. After collection, samples were placed in large plastic bags and immediately transported to the laboratory. Samples were handled only with plastic, glass, or porcelain objects and refrigerated at about 4°C before the analyses.Analyses of heavy metals followed Montgomery et al. (1977) and Thompson (1969). Samples were rinsed thoroughly with deionized water, shaken to remove most of the water, allowed to air dry, and grounded in a ceramic mortar. Approximately three grams of finely cut material was weighed in a porcelain dish that had been heated at 600°C for 2h. Samples were then dried in an oven at 65°C for 24h, allowed to cool in a desiccator, weighed, and incinerated in a muffle furnace for 2–3 h at 575°C. Ashes were dissolved in 5ml of 20% HCl and filtered. The concentration of acid-extractable elements was determined by air-acetylene flame detection in an atomic absorption spectrophotometer (Perkin Elmer Model AA100). Results were evaluated by the t-test using SYSTAT version 11.The elemental composition of vegetative samples obtained at the Vieques’s bombing range is presented in table 1. In general, distinctive profiles are observed within the studied species thus reflecting differences in their physiological properties. Copper and nickel were most abundant in Sporobolus virginicus and manganese was more abundant in Syringodium filiforme at the Guánica State Forest. In contrast, higher levels of lead were detected at the AFWTF than control populations collected from mainland Puerto Rico (Figure 2). The content of lead in pasture grass (P. maximum) from Vieques was as high as 13 μg/g (dry weight), which is above safety guidelines [1]. Lead is used in ammunition, metal products (targets), batteries, and paints. The vegetation therefore, could be an intermediate reservoir through which trace elements in soil, water, or air move to animals and humans [2, 3]. From 1984 to 2000, the US Navy allowed local farmers to graze cows in the eastern part of Vieques including at the AFWTF. The potential for direct exposure and the impact on human health is exemplified by this pathway. Trace element analysis conducted on the island residents has demonstrated significantly higher concentration of various elements including mercury, lead and cadmium [6, 7].In addition, lead levels up to 33μg/g (dry weight) were detected in fruit material of Ipomoea violacea. Both local and migratory birds such as Zenaida aurita (Zenaida Dove) and Spindalis portoricensis (Puerto Ricen Spindalis) feed on this fruit [11]. Natural lead concentration in plants growing in uncontaminated areas is usually low, ranging from 0.1 to 10μg/g (dry weight) and averaging 2μg/g [3]. Moreover, the concentration of cobalt and manganese found in the vegetation in Vieques was greater than those detected in the control population. Fire, decomposition of dead vegetable tissue, aerial dispersion and consumption by herbivores could all be historic and/or on-going transport pathways out of the former bombing range.At the marine ecosystem, levels of lead detected in Syringodium filiforme from the AFWTF indicate the dispersion of metals throughout the marine food chain (Table 2). In 2001, differences were significantly higher (p<0.05) at the AFWTF for lead, copper, nickel and cobalt. Only aluminum was higher (p<0.05) in plants collected at Guánica. The content of lead in S. filiforme cannot be explained solely as a result of natural processes. The oceanic pH (approximately 8.0 ± 0.5) limits the solubility of many metals, including lead, and metals must be dissolved in order to be available for marine plants to accumulate in their tissues. At the AFWTF, the US Environmental Protection Agency Discharge Monitoring Reports from 1984–1999 identified excessive concentrations of lead with occasional average levels of up to 5 mg/L, as well as fluctuations in pH [13]. These parameters could enhance metal bioavailability, thus increasing uptake of metals by marine life during military maneuvers. Heavy metals were undetectable in seawater when military practices did not take place [10]. In May 2003, military operations ceased at the AFWTF. Samples obtained in 2004 of S. filiforme showed lower concentrations (p<0.05) of cobalt, copper, nickel and lead to those levels observed in 2001. At the AFWTF however, the level of these elements are yet higher (p<0.05) than mainland, Puerto Rico.Syringodium is commonly found in shallow waters in southern Puerto Rico. Tribble (1981) demonstrated than coral reef fish have a preference for Syringodium rather than other marine plants such as Thalassia. Their distribution in coral reefs is usually limited by selective grazing activity [9]. The level of lead and other elements in S. filiforme demonstrate the potential for dispersion and dangerous bioaccumulation along the marine food chain. Fishes, crustaceans, and manatees directly or indirectly consume this marine plant. The US Fish and Wildlife Service reported manatees feeding in Vieques and most intensively near the former AFWTF. Understanding the dynamics of trace elements and other pollutants at this location could help establish management practices intended to prevent further exposure to human and endangered species. In turn, mitigation and better restoration mechanisms might be developed.Sampling locations at the former Atlantic Fleet Weapons Training Facility (Vieques, PR).Average content of lead (Pb) in various plant species from the AFWTF (Vieques) and reference location (Mainland, Puerto Rico).Elemental analysis of terrestrial plant material collected at the AFWTF, Vieques (Puerto Rico).Average (standard deviation; n = 2 to 4);nd, not-detectable.Elemental analysis of Syringodium filiforme collected at the AFWTF and Guánica State ForestAverage (standard deviation; n = 2 to 10);na = not available.This project was funded in part by Casa Pueblo de Adjuntas, the Industrial Biotechnology Program of UPRM and the PR Department of Health. We thank Y. Zenón for his help during sampling activities at the AFWTF, and the UPRM botanist’s group and C. Delannoy for their contribution to this study.
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+ Polycyclic Aromatic Hydrocarbons (PAHs) are a group of compounds that pose many health threats to human and animal life. They occur in nature as a result of incomplete combustion of organic matter, as well as from many anthropogenic sources including cigarette smoke and automobile exhaust. PAHs have been reported to cause liver damage, red blood cell damage and a variety of cancers. Because of this, methods to reduce the amount of PAHs in the environment are continuously being sought. The purpose of this study was to find soil bacteria capable of degrading high molecular weight PAHs, such as pyrene (Pyr) and benzo[a]pyrene (BaP), which contain more than three benzene rings and so persist in the environment. Bacillus subtilis, identified by fatty acid methyl ester (FAME) analysis, was isolated from PAH contaminated soil. Because it grew in the presence of 33μg/ml each of pyrene, 1-AP and 1-HP, its biodegradation capabilities were assessed. It was found that after a four-day incubation period at 30°C in 20μg/ml pyrene or benzo[a]pyrene, B. subtilis was able to transform approximately 40% and 50% pyrene and benzo[a]pyrene, respectively. This is the first report implicating B. subtilis in PAH degradation. Whether or not the intermediates resulting from the transformation are more toxic than their parent compounds, and whether B. subtilis is capable of mineralizing pyrene or benzo[a]pyrene to carbon dioxide and water, remains to be evaluated.Polycyclic Aromatic Hydrocarbons (PAHs) are a class of compounds composed of two or more fused benzene rings [1, 2]. They occur naturally, resulting from the incomplete combustion of organic matter, volcanic eruptions and forest fires. They are also distributed into the environment due to such human activities as cigarette smoking, automobile exhaust, and the processing, production and spillage of petroleum [3, 4]. Many of these compounds have been found to have toxic, mutagenic and/or carcinogenic properties [5–7]. PAHs are photocytotoxic to both plants and animals [8, 9]. They are found in soil and water, and therefore are able to enter vegetation and aquatic organisms that humans will ultimately consume. Their fused benzene ring structure makes PAHs very hydrophobic and therefore, highly stable. These properties make them persist in the environment as recalcitrant pollutants for very long periods of time. For these reasons, the removal of PAHs from the environment is important and necessary.Biological systems have been widely used in various aspects of the remediation of the environment. Plants, fungi and microorganisms have been used as very effective tools in the removal of toxic metals and harmful organic xenobiotics from the environment. Microbial activity is the most significant cause of PAH removal from the environment, and the ability of bacteria to degrade PAHs has been well documented [10]. This documentation, however, is mainly limited to low molecular weight (LMW) PAHs, which contain fewer than four benzene rings. High-molecular weight (HMW) PAHs, containing four or more benzene rings, are more difficult to degrade because they are extremely hydrophobic [11, 12]. Few bacteria strains have been reported with the ability to degrade HMW-PAHs. The focus of this research is finding soil bacteria that are capable of degrading HMW-PAHs, such as Pyrene (Pyr) and Benzo[a]pyrene (BaP). In this study, we hypothesize that soil bacteria capable of degrading HMW-PAH will be isolated from PAH-contaminated soil.Tryptone, bacto-agar, sodium chloride (NaCl), yeast extract, K2HPO4, KH2PO4, (NH4)2SO4, sodium citrate•2H2O, MgSO4•7H2O, Ca(NO3)2, MnCl2, FeSO4, glucose, pyrene (Pyr), 1-aminopyrene (1-AP), 1-hydroxypyrene (1-HP), benzo[a]pyrene (BaP), N, N-dimethylformamide, acetonitrile and HPLC grade hexane were purchased from Sigma-Aldrich Chemical Co., (St. Louis, Missouri). Contaminated soil sample taken from a site in Southern Maryland was provided by Dr. Huey-Min Hwang.Luria Bertani (LB) broth contained 5.0g tryptone, 5.0g NaCl and 2.5g yeast extract per liter of distilled water. Minimum Medium (MM) was prepared by dissolving 13.9g K2HPO4, 6.0g KH2PO4, 2.0g (NH4)SO4, 1.9 g Sodium Citrate • 2H2O, 0.2g MgSO4 • 7H2O, 0.1ml 1M Ca(NO3)2, 1.0ml of 0.1M MnCl2 and 0.1ml of 0.01M FeSO4 in 1L of distilled water. Glucose stock solutions (10%) were prepared. All solid media contained 1.8% agar along with the other appropriate constituents. Stock solutions (1.0mg/ml) of pyrene (Pyr), 1-aminopyrene (1-AP) and 1-hydroxypyrene (1-HP) and benzo[a]pyrene (BaP) were prepared in N, N-dimethylformamide (DMF).One gram of PAH-contaminated soil was re-suspended in 100ml of sterile dH2O with vigorous stirring for 10 minutes. This suspension was then allowed to settle at room temperature for 30 minutes. A 100μl sample of the supernatant was spread onto a LB agar plate and incubated at 30°C. Colonies from the LB plate were then spread onto Minimum Medium plus 1% glucose (MG) plates and incubated at 30°C. These spread and transfer processes were repeated several times on MG, MM, and finally, MM+PAH (1μg/ml each Pyr, 1-AP, 1-HP and BaP). Eleven (11) distinct isolates were obtained. One colony of each of the eleven isolates was used to inoculate 5ml of MG and incubated at 30°C in an orbital shaker at a speed of 150 rpm. The cultures were then stored in 10% glycerol at 4°C and labeled Tgr1-11. A 200μl aliquot of each isolate from frozen stock was used to inoculate 10 ml of MM+PAH and incubated at 30°C in an orbital shaker at a speed of 150 rpm for 24hr. Over a four week period, the PAH concentration was increased from 10, 15, 20, 25 to ~33μg/ml. The isolates that survived were subsequently sent to Dr. Veronica Acosta at the USDA in Lubbock, TX for identification by Fatty Acid Methyl Ester (FAME) analysis.Five milliliters of MM were inoculated with 100μl of Tgr3. Cultures were incubated in an orbital shaker whose cover was wrapped with aluminium foil to exclude light. Temperature was set for 30°C and shaker speed was 150 rpm for 24 hours. From the 24 hour culture, a subculture was grown in MM until it reached midlog phase (12–14 hours). A 500μl sample of the midlog culture was used to inoculate 25ml of MM, MG and MM supplemented with 20μg/ml pyrene or benzo[a]pyrene. Turbidity measurements were taken using a Klett Colorimeter from the time of inoculation (to) until measurements remained constant.Five milliliters of MM were inoculated with 100μl of Tgr3. Cultures were incubated in an orbital shaker whose cover was wrapped with aluminum foil to exclude light. Temperature was set for 30°C and shaker speed was 150 rpm for 24 hours. From the 24 hour culture, a subculture was grown in MM until it reached midlog phase. A 100μl sample of the midlog culture was used to inoculate 5ml of MM, MG and MM supplemented with Pyr or BaP, to a final concentration of 20μg/ml. The cultures were incubated at 30°C for 4 days, with shaking. Every 24 hours, serial dilutions of the cultures were made and spread onto a MM plate. Plates were incubated until colonies appeared (~48 hours) and then counted.Five milliliters of MM were inoculated with 100μl of Tgr3. Cultures were incubated in an orbital shaker whose cover was wrapped with aluminum foil to exclude light. Temperature was set for 30°C and shaker speed was 150 rpm for 24 hours. From the 24 hour culture, a subculture was grown in MM until it reached midlog phase. Half of the midlog culture was autoclaved at 121°C/15psi for 15 minutes. A 100μl sample of the midlog culture or heat-killed culture was used to inoculate 5ml of MM supplemented with pyrene or benzo[a]pyrene, to a final concentration of 20μg/ml. The cultures were incubated at 30°C for 4 days, with shaking. Every 24 hours, the PAH was extracted with 2.0 ml of HPLC grade hexane, 1.0ml at a time, vortexed at maximum speed for 20 seconds, and incubated at −20°C to aid with layer separation.The PAHs extracted from completed transformation experiments were analyzed using Waters HPLC system consisting of dual λ absorbance detector and a Supercosil LC-PAH (5μm, 25.0cm × 4.6mm ID) column. Analysis was performed in gradient mode from 60% water (solvent A) to 100% acetonitrile (solvent B) at a flow rate of 1.5ml/min. After isocratic run for 5 minutes in 40% solvent B, the mobile phase was changed by a linear gradient over 25 minutes to solvent B and elution continued for a further 10 minutes. The UV light absorption was monitored at 2 nm wavelength intervals from 210 to 600nm. The wavelength used to integrate peaks was 254 nm.We spread 100μl aliquots of the supernatant from suspension of PAH-contaminated soil in water on LB plates. Successive spread and transfer of bacteria to different growth media supplemented with selected PAH, allowed us to isolate about a dozen bacteria. Isolates were grown in MM for 24 hours prior to initial characterization using Gram staining according to the Becton-Dickinson Gram stain protocol. As shown in Table 1, 11 distinct isolates were obtained. These isolates were further characterized using fatty acid methyl ester (FAME) analysis. Fatty acids were extracted from pure bacteria isolates using the procedure described for pure culture isolates by the Microbial Identification System (MIS, Microbial ID, Inc., Newark, DE). The fatty acid methyl esters (FAME) were extracted and sent to the USDA Plant Stress and Water Conservation Laboratory in Lubbock, TX where they were analyzed in a 6890 GC Series II (Agilent Technologies). FAMEs were identified, and their relative peak areas were determined by the MIS Aerobe method of the MIDI system (Microbial ID, Inc., Newark, DE). The aerobic library for the TSBA 40 method (Version 4.10) was used to identify the bacteria. Isolates 3, 4, (Tgr3 and Tgr4) were identified as Bacillus subtilis and Burkholderia cepacia, respectively. Both isolates 7 and 9 (Tgr7 and Tgr9) were identified as Pseudomonas cepacia.To assess the potential of B. subtilis as a PAH degrader, several investigations were done. B. subtilis is able to grow in 20μg/ml pyrene and benzo[a]pyrene. Growth physiology results suggest that the growth of B. subtilis was not affected by 20μg/ml pyrene or benzo[a]pyrene because its growth in either PAH was almost identical to that in minimum medium alone (Figure 1). Turbidity measurements taken after every 24 hours of growth also indicated that B. subtilis reached exponential phase within the first 24 hours of growth and began to decline slowly thereafter in each growth medium. Figure 2 shows that B. subtilis achieves maximum viability at 48 hours in all media used. As expected, MG is still the best growth medium, with virtually no difference between MM, MM+Pyr and MM+BaP.PAH transformation experiments revealed that B. subtilis has the ability to transform pyrene and benzo[a]pyrene in its growth medium. PAH transformation occurred within the first 24 hours of growth in 20μg/ml pyrene and benzo[a]pyrene, with approximately 15% and 8% of the pyrene and benzo[a]pyrene being transformed, respectively (Figure 3). B. subtilis reached its peak growth (~ 1.25 × 10 11cells/ml) after 48 hours in pyrene, where it transformed about 55% of the pyrene in the growth medium (Figure 3). After 48 hours of growth, B. subtilis transformed about 50% of the benzo[a]pyrene in its growth medium, where it was also at its peak growth (~ 1.09 × 1011 cells/ml). Even though the number of viable cells significantly decreased after 72 hours of growth in benzo[a]pyrene, B. subtilis continued to transform up to about 65% of the benzo[a]pyrene present. As far as we know, this is the first report suggesting a potential role for B. subtilis in the degradation of PAH.Soil bacteria are frequently examined for their abilities to degrade PAHs because of the prevalence of the compounds in soils. In previous research, bacteria have been isolated from contaminated soils in areas such as coal gasification sites, creosote-contaminated sites and oil fields [1, 13–14]. These sites are usually assumed to be contaminated with PAHs because of these anthropogenic activities. Since bacteria are present virtually everywhere in nature, it is reasonable to expect them to be found in PAH-contaminated sites. Those bacteria that survive in such sites are likely to be able to degrade or metabolize the contaminant. Few bacteria are known that can degrade HMW PAHs. White rot fungi are more often associated with HMW PAH degradation than bacteria. They are believed to produce such enzymes as lignin peroxidase and P450 monooxygenases for use in PAH degradation [15]. As the search for HMW PAH degrading bacterium goes on, more genera are being identified as degraders. The Mycobacterium genus is the family that is most often associated with degradation of HMW PAHs [14, 16, 17]. Mycobacterium austroafricanum GTI-23 can utilize phenanthrene, fluoranthene and pyrene as a sole source of carbon and energy, and mineralize 300μg/ml pyrene in liquid culture [18]. It can also partially degrade 300μg/ml fluorene and benzo[a]pyrene. After 17 days of growth, M. austroafricanum completely degraded [4, 5, 9, 10 14C]Pyrene to 14CO2.Another Mycobacterium, strain RJGII-135, isolated from a coal gasification site, was able to degrade more than 45% [14C ]Pyrene, with metabolites such as methylated -4-phenanthtrene-carboxylic acid, methylated-4,5-Phenanthrene dicarboxylic acid and 4,5-pyrene-dihydrodiol being produced [14]. Although significantly slower, RJGII-135 also degraded [14C]benzo[a]pyrene, producing metabolites cis-4-(7-hydroxypyrene-8-yl)-2-oxobut-3-enoic acid, methylated 4,5-chrysene-dicarboxylic acid, cis-7,8-dihydrodiol-BaP, 7,8-dihydro-pyrene-8-carboxylic acid. This organism was also able to degrade benzo[a]anthracene at a rate similar to that of pyrene. Mycobacterium sp. strain AP1 grew in mineral medium with pyrene as the sole source of carbon [16]. After 6 days of growth, the bacterium decreased the amount of pyrene in the growth medium from 180 g/ml to 50 g/ml while producing, for the first time in literature, metabolite 6,6-dihydroxy-2,2-biphenyl dicarboxylic acid.Pseudomonas stutzeri strain P16, Pseudomonas saccharophila strain P15, Bacillus cereus strain P21 and Sphingomonas yanoikuyae strain R1 all play an active role in pyrene metabolism when grown in minimum salts buffer [19]. Strains P16 and P21 are able to transform pyrene into the first intermediate in degradation by aerobic organisms, cis-4,5-dihydro-4,5-dihydroxypyrene. Strains P15 and R1 were not able to transform pyrene into any intermediates, but they were able to further metabolize cis-4,5-dihydro-4,5-dihydroxypyrene into pyrene-4,5-dione.Isolates 3, 4, (Tgr3 and Tgr4) were identified as Bacillus subtilis and Burkholderia cepacia, respectively. Both isolates 7 and 9 (Tgr7 and Tgr9) were identified as Pseudomonas cepacia by fatty acid methyl ester analysis. Results of transformation experiments clearly suggest that B. subtilis transforms about 40% of pyrene and 50% of benzo[a]pyrene in its growth medium. B. subtilis has previously been used in the bioremediation of selenite, a mutagenic selenium compound that causes base-pair substitution in DNA, because of its ability to convert selenite to elemental selenium [20]. These transformation results are significant because as far as we know, this is the first report suggesting a potential role for B. subtilis in the degradation of PAH. This is even more fascinating considering the fact that Tgr3 was isolated from the same soil sample as Burkholderia cepacia, isolate 4 (Tgr4) and Pseudomonas cepacia, isolates 7 and 9 (Tgr7 and Tgr9), which have previously been reported to degrade HMW-PAH [21–22]. This finding warrants further extensive investigation of Tgr3 isolate. The toxicity of the metabolites produced during transformation must be assessed. Mineralization of pyrene and benzo[a]pyrene by Tgr3 is yet to be determined. If as these results seem to suggest B. subtilis is able to degrade HMW-PAH, its use in the clean-up of PAH-contaminated sites will go a long way in improving the health of the environment for the benefit of humankind.Growth physiology of B. subtilis in MM, MG, and MM+PAH. B. subtilis was grown in each medium for 4 consecutive days. Turbidity measurements were taken every 24 hours using a Klett colorimeter. Results represent the mean +/−SD values of experiment performed in triplicate.Viability of B. subtilis in MM, MG and MM+PAH. B. subtilis grown in each medium was spread onto MM plates every 24 hours. Plates were incubated at 30°C for 48 hours and colonies were counted. Results represent the mean +/− SD values of experiment performed in triplicate.Transformation of benzo[a]pyrene and pyrene by B. subtilis. B. subtilis was grown in MM plus 20μg/ml BaP or Pyr for 4 consecutive days. Every 24 hours, the PAH was extracted with HPLC grade hexane and analyzed by HPLC. Results represent the mean values of experiment performed in triplicate.Characteristics of soil bacterial isolates: Isolates were grown in MM for 24 hours and Gram stained according to the Becton-Dickinson Gram Stain Protocol Growth and Viability of Tgr3This research was supported in part by a grant from the Army Research Office (Grant# DAAD 19-01-1-0733), awarded to Jackson State University, and in part by a grant from the National Institutes of Health Research Centers in Minority Institutions, NIH-RCMI (Grant #1G12RR13459) awarded to Jackson State University. Part of this work was used by Rochelle Hunter to satisfy the Masters Degree thesis requirement at Jackson State University. I thank Dr. Veronica Acosta for her technical assistance in identifying the isolates by Fatty Acid Methyl Ester (FAME) analysis.
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+ The degradation of various formulations of the racemic mixture and the enantiomers (including mefenoxam) of metalaxyl in typical soils from Germany and Cameroon in controlled incubation experiments was studied. The kinetics of the degradation or transformation was determined by means of reversed phase HPLC, while the enantiomeric ratios were measured by HPLC with a chiral Whelk O1 column. The dynamics of the quantitative changes in microbiological properties induced by the addition of these fungicides at their recommended field rates were determined in the soils during a 120-day incubation experiment. The degradation followed first-order kinetics (R2≥0.96). Higher metalaxyl acid metabolite concentrations were found in German than in Cameroonian soils. The enantiomers of the fungicide had different degradation rates in both soils, with half-lives ranging from 17 to 38 days. All forms of metalaxyl had lower degradation rates in the Cameroonian soil than in the German soil. The degradation of the R-enantiomer was much faster than the S-enantiomer in the German soil and slower than the S-enantiomer in the Cameroonian soil, suggesting that different microbial populations, which may be using different enzymes, have different degradation preferences. The type of soil significantly influenced the effect of these fungicides on the soil parameters studied. Incorporation of these fungicides resulted in a change in the ecophysiological status of the soil microbial community as expressed by microbial activities. The activity of phosphatases and β-glucosidase, the mineralization and availability of N and most plant nutrients in soils were stimulated, whereas the activity of dehydrogenase and the availability of NO3−, were generally adversely affected. The soil NH4+, NO3−, and enzymes activities values in general did not correlate with the degradation of metalaxyl in both soils. However, the degradation of formulated and unformulated metalaxyl was positively correlated to the activity of acid phosphatase in the German soil (R2, 0.84 and 0.94 respectively) and in the Cameroonian soil (R2, 0.97 and 0.96 respectively).The phenylamide fungicide, metalaxyl, is a chiral compound (structural formula in Fig. 1), which is marketed in its racemic as well as in its enantiopure R form (trade names, e.g., mefenoxam, ridomil etc.). It is manufactured by Syngeta and is used as a seed treatment for banded or broadcast soil application and as a foliar spray in combination with protectant type fungicides such as copper or folpet. It has activity against fungal pathogens of the order Peronosporales which cause late blight, downy mildew, damping off, and stem and fruit rots of many plants. The compound is taken up by roots, leaves, green stems and shoots and transported acropetally within the plant, and inhibits the fungal protein synthesis [1, 2]. It is stable under a broad range of pH, temperature and light [3].Because of its broad-spectrum activity, metalaxyl is registered for use on a wide range of crops and in several countries in temperate, subtropical and tropical regions. The addition of metalaxyl, while enhancing the plant growth and crop yield [1], can affect the homeostasis of the soil system [4]. Any perturbation is likely to lead to a shift in the equilibrium of the system while directly affecting the structure and function of the soil microbial community. This is evidenced by previous observations [5], showing that applications of metalaxyl on vineyard soils over 3 years, markedly decreased microbial numbers and decreased the activity, and increased the number of micro-organisms involved in the mineralization of organic matter. It was reported that the systemic application of metalaxyl induced a brief stimulation and a subsequent suppression of soil fungi and actinomycetes [6].Metalaxyl has two enantiomers which are expected to be formed in a 1:1 ratio if synthesized from racemic materials [7]; however, the enantiomeric distribution should be checked, as some assumed racemates do not give 1:1 peak ratios. Many reports have documented the microbially-mediated degradation of metalaxyl in soils [8–10], and the faster degradation of the R-metalaxyl in temperate soil [7, 11–13]. These reports have not addressed degradation behaviour of the enantiomers applied in various formulations of metalaxyl in soils of different climatic environments. Information about pesticide dissipation with time is essential in assessing environmental risks [14]. The replacement of racemic metalaxyl by, e.g., mefenoxam [15], means that knowledge of and data on the persistence of this enantiomer are essential for use, management and its registration, especially in the tropical regions of Africa, e.g., Cameroon, where metalaxyl is heavily used in cocoa farming [16]. Additionally, knowledge on the degradation mechanism is a prerequisite for the registration in new fields of application.Mefenoxam (also called R-metalaxyl) is the R-enantiomer of metalaxyl and has been on the market since 1996 under various formulations and trade names including Ridomil gold, Fonganil gold, Apron XL, Subdue, MAXX. It provides the same level of efficacy as metalaxyl but at half the application rate. The introduction of mefenoxam may contribute to risk reduction for metalaxyl [17]. Thus, mefenoxam is to replace technical metalaxyl in parts of the world where the registration of metalaxyl has not been renewed. Many indicators of soil quality and health have been suggested including, potentially mineralizable N and soil enzymes [18–19]. The importance of soil enzymes resides in their relationship to soil biology, their ease of measurement, and their rapid response to changes in soil management. Concern over the effects of these fungicides on soil processes is based on the fact that many of the reactions in nutrient cycling are mediated by microbes [20]; there is also the possibility that these chemicals can enter into the food chain and, thus, affect higher organisms including humans.Application rates of metalaxyl range from 0.151 to 8.970 ai (active ingredient) kg ha−1 for agricultural crops, from 0.154 to 0.700 g ai kg−1 seed for agricultural seed treatment, from 1.00 to 8.07 kg a.i ha−1 for ornamental trees and plants. Multiple applications (depending on plant) are approved. This intensive use of metalaxyl has not been without problems. Reports have shown that strains of fungi resistant to metalaxyl may develop [21]. Metalaxyl has been shown to affect soil biology adversely [6, 22]. Moreover, metalaxyl has been found in the water supply (sea and ground water) [23–24] and in food [25].Mefenoxam is a new product and quantitative studies on its fate and effects are required. Numerous studies have documented changes in the soil ecosystem as a result of pesticide application [26]. Information is scarce concerning the effect of mefenoxam and metalaxyl on soil quality. Since phenylamide fungicides are among the fungicides most frequently employed worldwide, it is important to consider their possible impact on soil response to changes in soil management. Enzyme analyses integrate chemical, physical and biological characteristics and can be used to monitor the effects of soil management, including pesticide use on long-term productivity. Several enzyme activities were measured simultaneously, in order to obtain a more valid estimate of the metabolic response of soil to fungicide stress following a reported suggestion [27].Pesticide degradation studies are essential to evaluate its impact in the environment and on non-target organisms. In this study, a combination of chemical and biological soil properties have been used for the evaluation of soil properties changes resulting from a single application of the maximum commercial recommended field rate of mefenoxam and metalaxyl to tropical and temperate agricultural soils originating from Cameroon and Germany respectively. The specific objectives of this study were (i) to study the enantioselective degradation and persistence of the racemic and enantiopure forms of metalaxyl in tropical and temperate soils using reverse phase and chiral high-performance chromatography (HPLC); (ii) and to determine the effects of mefenoxam and metalaxyl amendments on selected soil property parameters including nitrogen transformation processes, and soil enzyme activities. We hypothesized that in cases where these fungicides are not toxic to microbiological processes, they would serve as carbon and nitrogen inputs to the soil. This could result in significant increases in soil microbial biomass and some of the more labile soil organic matter fraction [28]. These changes could eventually be followed by an increase in soil organic matter and nutrient availability [29]. Our secondary objectives were (i) to gain information on the effect of the formulation on degradation and (ii) to determine which of the soil property parameters tested could be used specifically as early warning indicators of any side-effects for mefenoxam and metalaxyl on soil biological activity.Soils used in this study originated from Cameroon and Germany. One soil was collected from Monheim, Germany (Temperate soil-German soil). It is an agricultural sandy loam soil, which is used regularly by Bayer AG for adsorption and degradation studies. This part of Germany receives an annual average rainfall of 750 mm, with an annual mean temperature of about 10°C [30]. The soil, originating from Cameroon (Tropical soil-Cameroonian soil), was collected from the experimental research farm of the Institute of Agronomic Research for Development (IRAD) in Nkolbisson, near Yaounde. It is a sand-clay loam soil. The red ferralitic soil in this site covers 60% on the national surface area of the country. This site receives an annual rainfall from 1400 to 1600 mm and the annual temperature ranges from 19 to 28°C [31].These two soils, which had not received any pesticide applications for at least 5 years, were taken from the surface layer (0–10cm), air-dried, disaggregated manually, passed through a 2-mm screen sieve, and mixed to achieve homogeneity. The moisture content of the air-dried soil was determined by difference in pre-and post-oven dried weights. The organic matter content of the soils was determined using the “Loss –On Ignition” method [32]. The soil main properties are listed in Table 1.Pure metalaxyl, analytical standard grade (purity, >99%), (P-metalaxyl), was obtained from Riedel-de-Haen, Germany. Formulated metalaxyl (F-metalaxyl) and mefenoxam were emulsifiable concentrate (EC) formulations containing 24% and 48% of metalaxyl and R-metalaxyl respectively. They were obtained from Novartis Agro GmbH, Frankfurt, Germany. The metalaxyl acid metabolite (CGA 62826) analytical grade (99%)) was obtained from Novartis Crop Protection AG, Basel, Switzerland. All other chemicals used in the study were analytical grade from Merck Co. or Aldrich Chemical Co.F-metalaxyl contains metalaxyl as active ingredient, which is a racemic mixture of R- and S-enantiomers, whereas mefenoxam contains only the active R-enantiomer. Both compounds, having the same molecular weight (279.34), empirical formula (C15H21NO4) and structural formulae (Figure 1), have been compared on a μg/g basis.Racemic metalaxyl, in its pure and formulated form, as well as formulated R-metalaxyl were spiked in different experiments to the two soils. The amounts administered correspond to field application rates of roughly 1 kg/ha racemic metalaxyl assuming a penetration depth of 5 cm and a soil density of 1,5 g/cm3 (details in Table 2). To avoid potential effects of solvents upon the microbiological activity of the soil, the volumes of the application solution were limited to 200–900μl and were dispensed onto portions of ~30 g air-dry soil in porcelain dishes, the treated sub samples of soils were thoroughly mixed with a spatula until the solvent was completely evaporated (~10min) and the respective compounds were evenly distributed. The respective sub samples were subsequently added to the total soil mass of the corresponding soils (700–1500g). Subsequently, the soil gross mass for each test chemical was mixed in a tumbling mixer for 1hr. This was followed by the adjustment of the moisture content of the soil to 60% of the maximum water holding capacity, to allow optimal conditions for activity of aerobic soil micro-organisms to occur. Aliquots were then taken. Batches of 100 g of soil each (based on dry weight) were incubated in the dark in an Erlenmeyer flask, under controlled temperature (20 ± 2°C) for 120 days. The moisture content in each flask was checked gravimetrically every 2 weeks and at each sampling period. During incubation, flasks corresponding to the appropriate period were removed for analyses of fungicide residues. Each incubation was carried out in duplicate.P-metalaxyl and F-metalaxyl (144 μg of active ingredient/100 g of soil on a dry weight basis each) and mefenoxam (72 μg of active ingredient/g 100 g of soil on a dry weight basis) respectively were mixed thoroughly and separately into the soils at the recommended commercial application rate for cocoa crop in Cameroon of 1,080 g active ingredient (a.i)/ha for P-metalaxyl, F-metalaxyl and 540 g a.i/ha for mefenoxam (Novartis, 2000, pers. communication). These application rates were based on the maximum single use rate, assuming a depth in the soil of 5 cm and a soil density of 1.5 g/cm3. To avoid the potential effects of solvents upon the microbiological activity of the soil, the calculated volumes of the application solution, 553μl and 550.5μl of the solution of P-metalaxyl in methanol for the German soil and for the Cameroonian soil respectively; 949μl and 814μl of solution of F-metalaxyl in water for the German soil and for the Cameroonian soil, respectively; 237.5μl and 203.5μl of solution of mefenoxam in water for the German soil and for the Cameroonian soil respectively, were dispensed onto portions of ~30g air-dry soil in porcelain dishes. The treated subsamples of soils were thoroughly mixed with a spatula until the solvent had completely evaporated (~20min). The subsamples were subsequently added to the total soil mass of the corresponding field moist soils [1114 and 1146g for German soil and for Cameroonian soil respectively for P-metalaxyl; 1912 and 1694 g for German soil and for Cameroonian soil respectively for F-metalaxyl and mefenoxam. Subsequently, the gross soil mass for each test chemical rate was mixed in a tumbling mixer for 1hr. This was followed by adjustment of the moisture content of the soil to 60% of the maximum water holding capacity, prior to sampling, to allow optimal conditions for activity of aerobic soil micro-organisms. Batches of 100 g of each soil (equivalent dry weight) were incubated in the dark in 750-ml glass jars, under controlled temperature (20 ± 2°C) for 120 days. The jars were kept covered with perforated parafilm. The moisture content in each flask was checked gravimetrically each week and at each sampling period. During the incubation, jars were removed periodically and sampled once for analyses of soil chemical and biochemical parameters. Each experiment was carried out in duplicate.Methanol Suprasolv (Merck, Darmstadt, Germany) (200 mL) was added to the respective Erlenmeyer flask containing incubated soil. The methanolic soil suspension was acidified with formic acid (400μl) (p.a. Merck, Darmstadt, Germany). The flask was subsequently shaken on an overhead shaker (Gelhardt, Rotierapparat RS20, Germany) at speed 6 for 1h. After the extraction the resulting mixture was spiked with 50μl of internal standard solution (metazachlor, 100μg/mL). The suspension was hand-mixed for about 15 seconds and allowed to settle for 15 minutes. The clear supernatant was decanted through a glass fibre filter (Gelman Sciences Type A/E Glass 142 mm) into a 500mL round-bottom flask. The residual soil slurry was re-extracted following the same procedure with 100mL methanol and 100μl formic acid. The filter was rinsed with 6mL methanol. The pooled methanolic extract was evaporated to ~10mL on a rotary evaporator at 300mbar and 60°C. The resulting extract was cleaned up using a preconditioned (6mL methanol followed by 6 mL water) C18 SPE cartridge (500mg, 6mL Baker, Deventer, Netherlands). The analytes were then eluted with 6mL methanol. The eluate was concentrated to ~5mL by means of a rotary evaporator at 150 mbar and 60°C, to yield the final soil extract. Twenty μl aliquots of this solution were analysed by reverse phase high-performance liquid chromatography (HPLC).The degradation of metalaxyl compounds and the formation of their metabolites were monitored by HPLC-DAD using a Gynkotek/Dionex HPLC system consisting of a degassing unit (ERC-3822), a Gina 50 autosampler, a Dionex P 580 pump, a column oven at 20°C and a diode array detector (UVD 340S) operated at 205 nm. The system was operated under control of the Chromeleon 6.0 software package. An Ultrasep PAK (L=250mm i.d 3mm and pore size 6μm) C18RP column was used for non-chiral separation. The mobile phase (0.5 mL/min) consisted of water and acetonitrile, both HPLC grade (Baker, Deventer, Netherlands) with 0.1% formic acid. The gradient was programmed: 0–13 min: 42% acetonitrile; 13–14 min: 42–>50%; 15–18 min: 50–>100%; 18–19 min:100%, 19–20 min: 100->42%; and 20–25 min: 42%. The calibration was performed as multilevel internal standard calibration (IS= metazachlor) by using metalaxyl and its acid metabolite prepared in HPLC-grade methanol. The procedure gave recoveries of 100.0% with 4.1 % RSD for metalaxyl and 96.0% with 7.2 % for its acid metabolite (recovery rates obtained from spiked German soil). The limit of determination in the analysis was 0.05 μg/g soil for both metalaxyl and its acid metabolite. Blanks were determined to be below this limit of determination. The data reported are uncorrected for recoveries.The identity of the compounds found in the samples was confirmed by comparing retention time and UV-spectra to those obtained from standard solutions in the same sequence. Additional confirmation of the identity of the compounds was obtained by HPLC-MS/MS in selected samples. The operating parameters were as follows: HPLC: Phenomex Luna 3 C18 100 A (150mm × 2mm) column, mobile phase; same as described above; flow 0.250mL/min. MS: TSQ 7000 (Finnigan MAT, Bremen, Germany) equipped with an ESI II and APCI source. For ESI, the ionisation voltage was set at 5 kV and transfer capillary temperature at 220°C. For APCI, a vaporizer temperature of 450°C and a transfer capillary temperature of 230°C were used. The ionisation current and detector voltage were set to 5μA and 1.3kV, respectively. For the selected reaction monitoring (SRM) scans, a dwell time of 20 ms and a scanning time of 0.5 s were applied. SRM transitions of 280->220 amu (metalaxyl) and 266->220 (acid metabolite) amu were utilized [9].Aliquots of the extracts that were prepared for RP-HPLC (0.5 mL) were conditioned for chiral HPLC by solvent exchange to n-hexane:2-propanol (80:20) by means of careful evaporation to dryness at room temperature under vacuum using an Eppendorf Concentrator 5301 (Eppendorf, Hamburg, Germany). The residues were redissolved in 0.1 mL n-hexane:2-propanol (80:20). A 5μl aliquot of this solution was then analysed by chiral HPLC using the same HPLC-DAD instrument as before for reverse phase analyses. In particular, a chiral column 250 × 4 mm of (R,R) Whelk-01 (5μm), obtained from Merck, Darmstadt, Germany, equipped with a diol pre-column 4×4 mm (Merck) was used. The mobile phase consisted of n-hexane:2-propanol (73:27) (isocratic) mobile phase. The flow rate was set to 0.9 mL/min. The calibration was performed by using high-purity metalaxyl and enantiopure R-metalaxyl standards prepared in HPLC-grade n-hexane:2-propanol (80:20). The (R, R) Whelk-01 column resolved the enantiomers of rac-metalaxyl (see Figures 2, 3). The elution order was S (10.25 min) before R (11.88 min). The enantiomer ratios in samples were determined using their peak area ratios. As in some cases racemates do not give peak ratios of 1:1 at all concentrations, the chiral separation was tested for concentration dependencies concerning the peak area ratios. Ratios of 1:1 were determined for the racemate for the whole range of concentrations that were analysed.First order rate constants were derived from “ln (Co/C) versus t” plots by linear regression analysis for each experiment (Excel 5.0, Microsoft, Inc.). The half-life (T1/2, days) was estimated from eq I.The enantiomeric composition (EC) was used as measure of the enantioselectivity of the degradation of enantiomers of metalaxyl in soils. The EC was defined by the eq II.where R and S are the concentration of R- and S-enantiomers in %, respectively. The EC values thus defined range from 0 (R=0, S=100%) to >1 (R>S). The EC for racemic metalaxyl is 1 (R=S). The degradation behaviour of R-and S-enantiomers was assessed by plotting values of enantiomer composition (EC) versus time (Figure 4).Available N (NH4+ and NO3−) was extracted into 2M KCl following the procedures described in “Soil analysis handbook reference methods” [32] and quantified using a reported colorimetric procedure [33].Dehydrogenase activity in soil was determined following the method of reduction of 2,3,5-triphenyltrazolium chloride (TTC) [34]. Each soil sample (20g) was thoroughly mixed with CaCO3 (0.2g) and 6 g of this mixture were treated in triplicate with 3% (w/v) 2,3,5-triphenyltrazolium chloride (1ml) and incubated for 24 h at 37 ± 1°C. The triphenyl formazan (TPF) formed was extracted quantitatively from the reaction mixture with methanol and assayed at 485 nm in a Shimadzu UV 1201 UV-VIS spectrophotometer. Acid and alkaline phosphatase activities were determined according to previously described methods [35, 36], with slight modifications. Soil samples (1g) were mixed with the modified universal buffer (MUB) of pH 6.5 and pH 11 for acid and alkaline phosphatase assays respectively and 0.05M p-nitrophenyl phosphate (1 ml) and incubated for 1 h at 37 ± 1°C. Then, 0.5M CaCl2 and 0.5 M NaOH (4ml) were added and the mixture was centrifuged at 3000 rpm for 10 min. The p-nitrophenol (PNP) in the supernatant was determined colorimetrically at 400 nm. Toluene was not included in the procedure because it has been shown to increase the observed activities of both acid and alkaline phosphatases [37] and can be used as source of C by most soil micro-organisms [38].β-Glucosidase activity was measured following a reported method [39]. Four ml of MUB (pH 6.0) and p-nitrophenyl-β-D-glucopyranoside (1ml) were added to soil (1 g) and the reaction mixture was incubated at 37 ± 1°C for 1h. The rest of the method was the same as described above for acid and alkaline phosphatase activity. No toluene was used in this assay. Results of enzyme activities are reported on an oven dry-weight basis, determined by drying the soils for 24 h at 105°C.The results at each sampling period and from over the total incubation period were compared using analysis of variance (ANOVA), with treatment as the independent variable. For the effects studies, results were reported as percentages of the control. When treatment responses differed significantly from controls (p<0.05), multiple comparisons were made using paired-t test procedure [40]. Effect levels for significant responses were based on the nominal test substance level at each sampling period.The presence of metalaxyl and its acid metabolite in the soil samples were verified with HPLC-MS/MS. No other metabolites were detected, though. The concentrations of F-metalaxyl and mefenoxam as well as their metabolites residues remaining in the German and Cameroonian soils are separately presented in Tables 3 and 4. The data for duplicate incubations are summarized in these tables. The degradation kinetics in German and Cameroonian soil experiments are presented in Figure 5. The concentrations of the parent compounds were plotted against time. Metalaxyl was degraded according to first order kinetics for the first 90 days. The linear regression equations were obtained from the ln (Co/C) vs. t plots. The correlation values under different conditions are presented in Table 5. The degradation of all forms of metalaxyl in both soils complied with the first-order reaction kinetics, with correlation coefficients, R2 ranging from 0.96 to 0.98 (Table 5). This compound was degraded in soils to levels of < 3% and <17% of the initial concentration within 90 d of incubation in German and Cameroonian soils respectively (Tables 3, 4) (Figure 5). Data for the formation and subsequent degradation of the primary metabolite, the metalaxyl acid are also shown in Figure 5.The degradation of F-metalaxyl and mefenoxam gave higher amounts of the acid metabolite in the German soil (Figure 5A) as compared to the Cameroonian soil (Figure 5B). As much as 89% of the initial concentration of mefenoxam was converted into its acid metabolite in German soil (Table 3) within 90 days of incubation compared to <1% in Cameroonian soil in the same period of incubation. Similar observation was made for F-metalaxyl, but a lower production of acid metabolite in the German soil (~49% of the initial concentration within 30 days of incubation) was obtained. The degradation profile of P-metalaxyl (data not shown) was similar to that of F-metalaxyl in soils. The maximum concentration of the acid metabolite formed from the P-metalaxyl was observed after 30 days of incubation in the German soil where ~57% of the initial concentration was converted into the acid metabolite. In the Cameroonian soil, as for the other products, the production of acid metabolite was significantly less important. Possibly the metalaxyl was transformed to other metabolites than the acid metabolite or even mineralised in the Cameroonian soil. A maximum concentration of ~4% at the 14d of incubation was observed. The degradation rate constant (k values) of these fungicides are listed in Table 5.In the chiral HPLC chromatograms obtained from the soil incubation of metalaxyl enantiomers, peak area ratios significantly different from the racemic standards were obtained. This indicated enantioselective degradation of metalaxyl in both soils (Figure 2). In the German soil the R-metalaxyl was degraded much faster (k = 0.064 day−1) than S-metalaxyl (k=0.033 day−1) when spiked with formulated racemic metalaxyl (see chromatograms in Figure 2). This is in agreement with previous reports on the behaviour of metalaxyl in temperate sandy loam soil [7, 13]. In the Cameroonian soil, the opposite was observed (see chromatograms in Figure 3). The S-enantiomer was degraded faster (k = 0.0 26 day−1) in relation to R-metalaxyl (k = 0.014 day−1) when spiked with F-metalaxyl. The EC values in this experiment changed from initially 1 to 0 after 75 d of incubation in German soil (evolution towards S), compared to changes from 1 to ~6 in the Cameroonian soil (progression towards R) (see Figure 4), indicating a more selective process by means of enantioselectivity for the degradation of metalaxyl in the German soil.The difference in the degradation behaviour of metalaxyl in soil could be explained by the fact that the different soil types may contain different microbial populations equipped with different enzymes, which are preferential degraders of different enantiomers. This could be one reason for the different ranking order of the two enantiomers of metalaxyl in their degradation rate constants.The changes of NH4+ and NO3− content in soils during the incubation are shown in Figure 6. In the shorter term (3–30 days), F-metalaxyl and P-metalaxyl caused a significant decrease in NH4+ content of both soils. Mefenoxam, while causing a significant decrease in the German soil NH4+ content, stimulated that of the Cameroonian soil as early as 14 days after application. The inhibitory effects of these fungicides on short-term exposure were reversible on long-term incubation (especially after 120 days) as evidenced by the significant increase in NH4+ content in both soils (Figure 6a, b). This increase was more pronounced with mefenoxam followed by F-metalaxyl. Probably mefenoxam is more bioavailable to bacteria than F-metalaxyl and P-metalaxyl. This assessment is supported by previous report indicating that after 21 days of incubation 78% of mefenoxam was degraded by rhizosphere microbial populations [11]. Metalaxyl has been reported to have DT50 values (the time taken for 50% active ingredient to be metabolized) in soil ranging from 3 to 8 weeks [41]. Similar trends were recorded for NO3− content in Cameroonian soil (Figure 6a). This indicated that these compounds stimulated the growth and the activities of ammonifying and nitrifying bacteria, which were mainly responsible for the mineralization of organic N to NH4+ and oxidation of NH4+ to NO3− respectively. The inhibitory effects of all these chemicals on nitrification was more pronounced in German soil, even at the 120th day of incubation (Figure 6b), but generally mefenoxam was comparatively less inhibitory and even showed a significant increase on the 75th day of incubation (Figure 6b). P-metalaxyl and F-metalaxyl effects resulted in delays of 30 days in the recovery of NH4+ and NO3− in the Cameroonian soil (Figure 6a). These effects can be considered normal according to the theoretical framework for testing the side-effects of pesticides [42]. As delays of recovery of this available N were more than 60 days in the German soil (Figure 6b), the effects of P-metalaxyl and F-metalaxyl were considered critical in this soil [42].Pesticides have been shown to have direct and indirect effects on soil enzyme activity [43]. The addition of fungicides, in general, stimulated the activities of phosphatases and β-glucosidase in both soils as shown in Figures 7, 8, 9. Dehydrogenase activity was generally negatively affected, especially in the Cameroonian soil treated with F-metalaxyl and P-metalaxyl (Figures 8, 9). Mefenoxam significantly stimulated the activity of acid and alkaline phosphatases and β-glucosidase in both soils (Figure 7). Dehydrogenase activity was the most affected of all enzyme activities under mefenoxam stress. This inhibitory effect was more pronounced in the German soil. This difference in the dehydrogenase activity in the two soils may be ascribed to the difference in the decomposition rates of fungicides or their transformation to less toxic by-products in both soils as suggested earlier [44]. This adverse effect was not permanent since a significant increase in this enzyme activity was observed on the 90th day of incubation in both soils (Figure 7). A significant increase in the two phosphatase activities and in β-glucosidase activity was observed on the addition of F-metalaxyl and P-metalaxyl to both soils (Figures 8, 9). This increase was more pronounced in the Cameroonian soil. Dehydrogenase activity was stimulated at the 90th and at the 30 to 90th day of incubation by F-metalaxyl and P-metalaxyl respectively in the German soil (Figures 8b, 9b). P-metalaxyl, however, significantly decreased the activity of dehydrogenase in the Cameroonian soil (Figure 9a). As P-metalaxyl and F-metalaxyl caused delays of more than 75 days of recovery of the dehydrogenase activity in the Cameroonian soil (Figures 8a–9a), their effects on the activity of this enzyme were, thus, critical in this soil [42].In general, dehydrogenase activity appeared more sensitive to all fungicides in both soils, though to varying degrees. This is in agreement with many reports on the adverse effects of pesticides including fungicides on the dehydrogenase activity [45]. Dehydrogenase occurs intracellularly in all living microbial cells and it is linked with microbial respiratory processes. Its rapid degradation in soils could be followed by cell death and, thus, it does not accumulate in soils [46]. The dehydrogenase activity has been reported to reflect the microbial activity of soil [27, 47, 48].The fungicides, in general, stimulated the activities of phosphatases and β-glucosidase. Being extracellular enzymes, they are generally protected from degradation by adsorption on clays or humic substances and, thus, may accumulate [27, 49, 50]. Moreover, the extracellular enzymes, immobilized by soil colloids, may not be as sensitive to fungicides as those associated with microbial cells [43].Results suggest that the addition of the fungicides led to a change in the ecophysiological status of soil microbial community as expressed by the availability of NH4+-N and NO3−-N and the enzyme activities in soils. An attempt was made to correlate the concentrations of fungicides after the respective incubation intervals in the German (Table 2) and the Cameroonian (Table 3) soils and the corresponding concentrations of NH4+-N and NO3−-N (Figure 6) and the enzyme activities (Figures. 7–9) in soils. A weak relationship (data not shown) was found when fungicide residual concentrations were plotted against NH4+-N and NO3−-N levels in both German (r2 = 0.00004 to 0.2058) and Cameroonian (r2 = 0.0672–0.5430) soils, suggesting a lack of proportionality between the transformation rates of nitrogen and the degradation of metalaxyl and mefenoxam in these soils. The degradation of fungicides studied did not correlate also with the activity of dehydrogenase, Alkaline phosphatase and β-glucosidase in the German (r2 = 0.0325–0.8529) and the Cameroonian (r2 = 0.0427–0.7519) soils. However, a close positive correlation was found between P-metalaxyl degradation and the activity of Alkaline phosphatase in the German soil (r2 = 0.9294) and between that of F-metalaxyl and the activity of β-glucosidase (r2 = 0.9877) in the Cameroonian soil. A close positive relationship was also found between the activity of acid phosphatase and the degradation of all the fungicides in the Cameroonian soil (r2 = 0.9611–0.9778). In the German soil only P-metalaxyl degradation showed a high correlation (r2 = 0.9417) with the activity of this enzyme.The almost lack of correlation found between some enzyme activities and the degradation of fungicides suggests that dehydrogenase, alkaline phosphatase and β-glucosidase in general may not be implicated in the degradation of the studied fungicides in soils, except for P-metalaxyl whose degradation may involve alkaline phosphatase in the German soil and the degradation of F-metalaxyl in the Cameroonian soil involving also β-glucosidase. The high correlation found for the acid phosphatase activity suggests that this enzyme may be closely involved in the degradation of the fungicides in the Cameroonian soil and to some extend in the degradation of F-metalaxyl and P-metalaxyl in the German soil.Our findings show that the degradation of metalaxyl in soil occurs with some chiral preference. The rate of degradation of enantiomers and thus, the chiral preference depends on the type of soil and is enzymes mediated of which acid phosphatase may play an important role, especially in the tropical soil. These findings may also have some relevance to pesticide registration and approval.The application of the investigated phenylamide fungicides at their maximum recommended field rates had positive and negative effects on soil chemical and biochemical properties. The positive effect on phosphatases and β-glucosidase activities was probably due to the microbial growth stimulated by the addition of these fungicides which served as source of energy. In general, mefenoxam like F-metalaxyl exerted a negative influence on biochemical parameters of soil as manifested by the observed decrease in available nitrogen, especially nitrate in German soil, and altered enzymatic activities, especially that of dehydrogenase. The effects of P-metalaxyl and F-metalaxyl on the content of available N were normal in the Cameroonian soil and critical in the German soil. These chemicals also exerted a critical effect on the activity of dehydrogenase in the Cameroonian soil. The stimulatory effect of mefenoxam on available nitrogen, phosphatases and β-glucosidase activities was, in general, greater than that of F-metalaxyl. F-metalaxyl and P-metalaxyl, in general, exerted similar effects on soil properties.Structures of the two metalaxyl enantiomers.HPLC-DAD chromatograms showing elution of R and S-metalaxyl in the German soil after a) 0; b) 14; and c) 30 days of incubation. Column: chiral Whelk O1; mobile phase: n-hexane:2-propanol (73:27).HPLC-DAD chromatograms showing elution of R and S-metalaxyl in the Cameroonian soil after a) 0; b) 75; and c) 120 days of incubation. Column: chiral Whelk O1; mobile phase: n-hexane:2-propanol (73:27)Profiles of the enantiomeric composition of F-metalaxyl and P-metalaxyl in soils.Profiles of Mefenoxam and F-metalaxyl fungicides degradation in A) German soil and B) Cameroonian soil. Note the concurrent formation and degradation of the respective acid metabolite resulting from Mefenoxam and F-metalaxyl. Concentrations plotted versus incubation time (days).Effect of fungicides on the availability of NH4+-N and NO3−-N in the Cameroonian (a) and German (b) soils. * P<0.05 between fungicide addition and no fungicide addition (paired Student t-test).Effect of mefenoxam on enzyme activities on both soils in the Cameroonian (a) and German (b) soils. * P<0.05 between fungicide addition and no fungicide addition (paired Student t-test).Effect of F-metalaxyl on enzyme activities in the Cameroonian (a) and German (b) soils. * P<0.05 between fungicide addition and no fungicide addition (paired Student t-test).Effect of P-metalaxyl on enzyme activities in the Cameroonian (a) and German (b) soils. * P<0.05 between fungicide addition and no fungicide addition (paired Student t-test).Selected physico-chemical properties of soilsRates of application and measured final concentration of the investigated fungicides in soilsa.Active ingredient (a.i.) refers to concentration of pure metalaxylConcentrations of metalaxyl and its enantiomers as well as the metabolite in the German soil after the respective incubation intervals.Concentrations of metalaxyl and its Enantiomers as well as the metabolite in the Cameroonian soil after the respective incubation intervalsDegradation Rate Constants (K); And Half-Lives (T1/2) as Derived from the Regression Line from the ln (Co/C)/t Plot, as Well as Correlation Coefficients (R2) Values for the Fit for the Degradation of F-metalaxyl and P-metalaxyl and Mefenoxam in German and Cameroonian Soils. Enantioselective Degradation of Racemic Metalaxyl in SoilsWe are thankful to the Alexander von Humboldt Foundation for periods of research spent by the first author in Germany, under a Georg Forster Research Fellowship. We thank Novartis Agro GmbH, (Frankfurt, Germany) for kindly supplying the mefenoxam and metalaxyl employed, and Mr. A. Achermann from Novartis Crop Protection AG (Basel, Switzerland) for providing the R-metalaxyl and metalaxyl acid metabolite standards. We thank Dr. T. Pfeiffer, University of Dortmund, for recording the mass spectral data, Dr. E. Hellpointner from Bayer AG, and Dr. Kai Bester from the University of Dortmund, for their technical suggestions. We acknowledge technical assistance from Michael Schlüsener, Jörn Sickerling, Jürgen Scheen and Jürgen Storp of the University of Dortmund.
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+ Explosive compounds have been released into the environment during manufacturing, handling, and usage procedures. These compounds have been found to persist in the environment and potentially promote detrimental biological effects. The lack of research on bioaccumulation and bioconcentration and especially dietary transfer on aquatic life has resulted in challenges in assessing ecological risks. The objective of this study was to investigate the potential trophic transfer of the explosive compounds 2,4,6-trinitrotoluene (TNT) and hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) using a realistic freshwater prey/predator model and using dichlorodiphenyltrichloroethane (DDT), a highly bioaccumulative compound, to establish relative dietary uptake potential. The oligochaete worm Lumbriculus variegatus was exposed to 14C-labeled TNT, RDX or DDT for 5 hours in water, frozen in meal-size packages and subsequently fed to individual juvenile fathead minnows (Pimephales promelas). Fish were sampled for body residue determination on days 1, 2, 3, 4, 7, and 14 following an 8-hour gut purging period. Extensive metabolism of the parent compound in worms occurred for TNT but not for RDX and DDT. Fish body residue remained relatively unchanged over time for TNT and RDX, but did not approach steady-state concentration for DDT during the exposure period. The bioaccumulation factor (concentration in fish relative to concentration in worms) was 0.018, 0.010, and 0.422 g/g for TNT, RDX and DDT, respectively, confirming the expected relatively low bioaccumulative potential for TNT and RDX through the dietary route. The experimental design was deemed successful in determining the potential for trophic transfer of organic contaminants via a realistic predator/prey exposure scenario.Explosive compounds were released to the environment during the manufacturing, handling, use, and disposal of munitions at military sites in the United States and throughout the world. The result was contamination of ground and superficial waters, and soils and sediments, sometimes at exceedingly high concentrations (e.g., 34 mg/L for TNT in superficial water) [1–2]. Explosives such as 2,4,6-trinitrotoluene (TNT) and hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) and their degradation products typically degrade slowly in many environmental matrices, therefore yielding long-term contamination at the military sites where they were released [3]. The nitroaromatic compound TNT was the most abundantly produced explosive in the world and was released to surface and groundwater mainly from runoff and leaching from storage and disposal areas and from receiving lagoons at munitions production and processing plants [1]. The cyclonitramine compound RDX is one of the most commonly used and most powerful munitions and was released in waste streams generated during manufacturing and processing activities, leaching from storage lagoons and burial areas, and from demilitarization operations [1]. RDX has a lower sorption coefficient in topsoil and is more commonly found in groundwater compared to TNT [4].Challenges in establishing ecological risks and remediation goals at contaminated military sites typically exist because of inadequate knowledge of the environmental fate and effects of explosives in aquatic ecosystems. Explosives and several related compounds are known to cause a variety of adverse effects in animals. Organism-level effects have been reported in a relatively small number of aquatic species (see review [1] and also [5–9]. The fate of explosives and related compounds in fish and aquatic invertebrates is poorly understood. Explosives and related compounds have low potential to bioaccumulate in animals as expected due to their low hydrophobicity [10–16]. Moreover, recent investigations revealed that TNT entering animals undergo extensive chemical transformations and the bioaccumulation of breakdown products typically exceeds the bioaccumulation of the parent compound in animal tissue [13–17]. All investigations of the bioaccumulation of explosive compounds in aquatic organisms used spiked water as the exposure medium.Fish bioaccumulate xenobiotic compounds through direct absorption (mostly through the gills) from contaminated water and through the ingestion of contaminated food [18] and water. Dietary exposure to organic contaminants results in significant bioaccumulation for a variety of compounds [19–20] and sometimes results in detrimental biochemical and physiological effects [21, 22]. While the dietary uptake of explosives in aquatic species has never been investigated, significant bioaccumulation of TNT metabolites in a species of salamander through exposure to contaminated prey has been reported [23]. Therefore, there is potential for dietary uptake of TNT and other explosives in fish.This study investigates the potential for dietary uptake of the explosives, TNT and RDX, to the fathead minnow (Pimephales promelas). The organochlorine dichlorodiphenyltrichloroethane (DDT) was used as a comparative compound because it is substantially more hydrophobic, and hence bioaccumulative, than explosives. Methods were designed to provide realistic exposure estimates for the trophic transfer of explosive compounds in aquatic systems. The source of dietary uptake was the freshwater oligochaete, Lumbriculus variegatus, pre-exposed to contaminants. Previous investigations of dietary uptake of xenobiotics typically used spiked food pellets.Radiolabeled trinitrotoluene (14C-TNT, 40 Ci/mol) was purchased from Chem Service (Westchester, PA). Non-labeled TNT was purchased from Sigma Chemical (St. Louis, MO). Radiolabeled hexahydro-1,3,5-trinitro-1,3,5-triazine (14C-RDX) was purchased from New England Nuclear Research Products (Boston, MA). Non-labeled RDX (>98 percent pure) was obtained from the Naval Surface Warfare Center (Indian Head, MD). Radiolabeled dichlorodiphenyltrichloroethane (14C-DDT) was purchased from Sigma Chemical Co. (St. Louis, MO). Manufacturer reported radiochemical purity and chemical purity were >98 percent for all compounds.Oligochaete worms, Lumbriculus variegatus, were obtained from a commercial vendor (Aquatic Bio Systems Inc., Fort Collins, CO) and maintained under flow-through culture conditions according to standard procedures [24] before use in the experiments. Typical individual worm mass was 6 mg. Laboratory-cultured juvenile fathead minnows (Pimephales promelas) approximately 6 weeks old were purchased from Aquatic Bio Systems Inc. (Fort Collins, CO). Fish were maintained in dechlorinated tap water prior to use in the experiments. Typical fish biomass was 100 mg.Exposure solutions were created by spiking 1 L of water with 2 ml of an acetone solution consisting of radiolabeled and non-radiolabeled compounds (TNT and RDX) or radiolabeled compound only (DDT). The specific activity (disintegrations per minute [dpm]/μmol) of the TNT and RDX exposure water was determined by measuring radioactivity (dpm/ml) via liquid scintillation counting (LSC) and explosive concentrations (μmol/ml) via high performance liquid chromatography (HPLC). Target water concentrations were 5, 8, and 0.02 mg/L for the TNT, RDX and DDT exposures, respectively. Target radioactivity in water was 33, 66, and 3dpm/μL for the TNT, RDX, and DDT exposures, respectively.Approximately 650 worms were exposed to 14C-TNT, 14C-RDX or 14C-DDT for 5 hours in separate 1-L glass beakers. Such short exposure period was selected to maximize the fraction of the total radioactivity in the tissue corresponding to parent compound, as the relative contribution of the breakdown products of TNT increase with exposure duration for L. variegatus[14]. Transfer factors were determined as the ratio between radioactivity in the worm (dpm/mg) and in the exposure water (dpm/μL). Water samples were taken for chemical analysis at the beginning and end of the exposure period. At termination of the exposure period, a subset of worms was sampled for chemical analysis and a subset was processed for use as prey items in the dietary exposure experiment. Groups of two worms were wrapped in aluminum foil packages and frozen (−20°C) until fed to the fathead minnows.Individual fish were placed in 600-ml glass beakers with 500 ml dechlorinated tap water. Fish were fed twice a day at 8:00 AM and 5:00 PM with a meal consisting two frozen worms. Three replicate fish were sampled on days 1, 2, 3, 4, 7, and 14. Fish sampling took place 8 hours after the morning feeding to allow purging of food from the gut.Water (1 ml) from worm and fish exposure beakers were mixed in 12 ml of xylene-based scintillation cocktail (3a70b, Research Product International, Mt. Prospect, IL) and analyzed for radioactivity on a Tri-Carb Liquid Scintillation Analyzer (Model 2500 TR, Packard Instrument, Meridien, CT, USA). Water was also analyzed for TNT, RDX, and the TNT breakdown products aminodinitrotoluenes (ADNTs) and diaminonitrotolunes (DANTs) using the U.S. Environmental Protection Agency method 8330 [25].For radioactivity analysis, two worms in six replicates were placed in scintillation cocktail and analyzed as described above. For TNT and RDX, 20 worms in triplicates were transferred to polypropylene bead-beater vials. Each vial received 100 mg of 1-mm glass beads and 0.5 ml of HPLC-grade acetonitrile. Samples were homogenized using a mini bead-beater (Biospec, Barttlesville, OK) for 100 sec at 4200 oscillations/min and sonicated for 1 hour at 18ºC in a water bath. Samples were centrifuged for 10 min at 7500 g at 4ºC. A fraction of the acetonitrile supernatant (0.05 ml) was assayed for radioactivity as described above and another fraction received 0.5 ml of 1% CaCl2, and was filtered through a PTFE 0.45-μm syringe filter into amber sample vials for HPLC analysis as described above. Laboratory reporting limits for tissue samples were approximately 1.2 mg/kg (6μmol/kg) for all analytes. For DDT, 20 worms in triplicates were transferred to scintillation vials. Each vial received 5 ml of HPLC grade acetonitrile. Samples were homogenized using probe sonication and centrifuged for 10 min at 7500-x g at 4ºC. A fraction of the acetonitrile supernatant (1 ml) was analyzed for radioactivity as described above and another fraction (1 ml) was used for separation of DDT parent compound and breakdown products using the thin layer chromatography analysis as previously described [26].Fish sampled at different time points of the dietary exposure were individually transferred to scintillation vials. Each vial received 1.0 ml of tissue solubilizer (T2, Research Product International, Mt. Prospect, IL). Following complete tissue solubilization (overnight), each vial received 1.0 ml of 1.2 N hydrochloric acid and was assayed for radioactivity as described above.For fish bioaccumulation data, completely randomized one-way analysis of variance (ANOVA) on ranks (Kruskal-Wallis) was used to determine differences between means of the various exposure periods at a 0.05 level of significance. Pairwise comparisons (Fisher LSD method) were used to determine significant differences between fish burden at different exposure periods.Radioactivity was used as a surrogate for expressing the sum concentration of parent compound and all the degradation products for TNT, RDX and DDT in water and whole animal samples. Parent compound and all its degradation products will be collectively referred to as TNT*, RDX*, and DDT*.Aqueous exposures of L. variegatus were conducted to produce prey material for use in the dietary transfer experiments. Mean measured compound concentration in the exposure solution was 4.8 mg/L for TNT and 8.1 mg/L for RDX. The concentration of DDT in the water was not measured using analytical chemistry. Breakdown of parent compound during the 5-h aqueous exposure, determined using HPLC analysis, was minimal (<3%) for TNT and non-detectable for RDX. The relative accumulation of compounds in prey tissue was expressed as water-to-prey transfer factors calculated as the ratio between radioactivity in the worm (dpm/mg) and in the exposure water (dpm/μL). The 5-hour transfer factor for DDT* in L. variegatus was much greater than those of TNT* or RDX*, and the relative bioaccumulation of TNT* was greater than that of RDX*, based on concentration determinations using total radioactivity (Table 1). Using measured tissue concentrations of parent compounds in worms determined using HPLC analysis (7.48 and 2.14 μg/g for TNT and RDX, respectively), 5-h bioconcentration factors (BCFs) were much lower (1.5 and 0.27 μL/g for TNT and RDX, respectively) than the transfer factors determined using radioactivity presented in Table 1. Higher relative bioaccumulation of DDT* in L. variegatus was expected based on the major differences in Kow among the compounds used in this study (log Kow = 6.2, 1.60 and 0.87 for DDT [27], TNT [28], and RDX [29], respectively) and the positive relationship between log BCF and log Kow[30].The relative contribution of parent compound to the overall bioaccumulation of radioactive compounds in L. variegatus varied substantially among exposures (Table 2). While for DDT the parent compound was dominant in worm tissues, TNT and RDX parent compound was present at much lower concentrations compared to its breakdown products. The fraction of total radioactivity corresponding to non-solvent-extractable metabolites also varied greatly and was highest for RDX and lowest for DDT. Extractable breakdown products were identified as 2- and 4-ADNT for the TNT exposure and dichlorodiphenyldichloroethane (DDD) for the DDT exposure (Table 2). Peaks of potential RDX breakdown products were not discernible in HPLC chromatograms, indicating the polar nature of non-identified extractable breakdown products of RDX. The bioaccumulation profile reported for L. variegatus in an aqueous exposure to TNT by Belden et al. [17] was similar to that determined in this study, indicating that TNT is efficiently biotransformed in L. variegatus, as reported for other aquatic invertebrates [31, 13–14], terrestrial invertebrates [17] and fish [15]. Different from the nitroaromatic explosive TNT, the cyclonitramine explosive RDX was less efficiently biotransformed in L. variegatus as all extractable radioactivity corresponded to the parent compound. Biotransformation of RDX has been reported for microorganisms [32] and plants [33] but not for invertebrates or fish. The breakdown of DDT in the tissues of L. variegatus was minimal, as expected based on the reported inefficiency of that invertebrate to metabolize hydrophobic organic compounds [34]. Unextractable radioactivity detected in animal tissue corresponded likely to covalently bound products associated with organic molecules. Such strong binding has been previously reported for TNT in invertebrates [14,16], cell cultures [35] and plants [36]. RDX was found as either parent compound or bound residues in the intracellular compartment of plant roots [37].Detectable concentrations of TNT*, RDX* and DDT* in fish fed contaminant-laden prey worms were determined over the 14-day exposure period (Figure. 1) using radioactivity as a surrogate for the parent compounds and their breakdown products. Tissue concentrations of TNT* and RDX* did not vary significantly over time (Figure. 1). However, an overall trend for increasing DDT* body burden was observed (Figure. 1) and concentrations measured at experiment termination were significantly higher than those determined during the first seven exposure periods. Therefore, the concentration of DDT in fish would likely have increased beyond levels detected at day 14 if the experiment had continued for longer exposure periods. Based on results from preliminary experiments, the whole-body concentrations of TNT and its major breakdown products and RDX in fish fed contaminant-laden worms were too low for detection and quantification using acetonitrile extraction and HPLC analysis. Therefore, the identity of the compounds accumulated in exposed fish from this experiment is unknown.The experimental design employed in this study proved successful for investigating the potential for dietary uptake of organic contaminants from invertebrates to fish. Juvenile fathead minnows consumed contaminant-laden frozen worms quickly and completely. Hansen et al. [38] also successfully used L. variegatus as the contaminant source in an investigation of metal dietary uptake in juvenile rainbow trout.Prey-to-predator transfer factors or bioaccumulation factors (BAFs) were determined as the ratio between radioactivity in the fish (dpm/mg) and in the prey (dpm/mg). The BAF for DDT* (0.422g/g) was substantially higher compared to that for TNT* (0.018g/g) and RDX* (0.010g/g) (Table 2). Comparison of those ratios indicates that while the concentration of DDT* in the fish was approaching the concentration of DDT* in its food source, the concentration of TNT* and RDX* in fish represented only a very small fraction of the concentration in the prey (>2 percent). The higher potential for trophic transfer of DDT and hydrophobic organochlorine compounds relative to less hydrophobic and more readily metabolized compounds has been well documented [39]. Johnson et al. [23] used earthworms exposed to TNT or PCBs (Arochor 1260) as prey and salamanders as predator to investigate the dietary uptake of those compounds in terrestrial systems. In that investigation, dietary bioaccumulation was much greater for the highly hydrophobic PCBs than for TNT and its metabolites, therefore corroborating our finding of much greater trophic transfer of DDT compared to TNT in the aquatic prey-predator system. The bioaccumulation factor for DDT obtained in this study (0.422 g/g) was lower than similar factors (0.89 – 2.80 g/g) reported for hydrophobic organochlorine compounds in rainbow trout [40], likely due to the short exposure duration used in the present study.Fish were held for 10-hours in clean water to allow for the digestion of their last meal and also the egestion of undigested prey from their guts. However, this period may not have been sufficient for complete gut clearance. Therefore, radioactivity associated with undigested prey in the fish gut may have accounted for at least a fraction of the radioactivity determined in whole fish. For TNT*, the radioactivity associated with a meal (two frozen worms), approximately 13,000 dpm, far exceeded the mean radioactivity in fish at exposure termination (3,100 dpm). For RDX, the radioactivity associated with a meal, approximately 3,000 dpm, also far exceeded the mean radioactivity in fish at exposure termination (370 dpm). Therefore, for both explosives, even a small amount of undigested prey could have accounted for the whole body burden measured in the fish. For DDT*, however, the mean radioactivity associated with a meal (4,400 dpm) was substantially lower than the mean concentration in the fish at exposure termination (24,200 dpm), indicating that most of the body burden in fish corresponded to radioactivity present in fish tissues rather than associated with undigested prey.The experimental design employing frozen aquatic worms as prey and juvenile fish as predator proved successful for investigating the potential for dietary uptake of organic contaminants in aquatic systems. This study demonstrated that dietary transfer from invertebrates to fish was negligible for TNT and RDX but relatively high for DDT, a compound substantially more hydrophobic than TNT and RDX.Body burden expressed as radioactivity (disintegrations per minute) per milligram of mass representing the total concentration (as parent compound equivalents) of the parent compound (TNT, RDX or DDT) and all its degradation products at different time points during the 14-day dietary exposure period.Radioactivity in water, prey (Lumbriculus variegatus), and fish (Pimephales promelas) expressed as mean (± 1 standard deviation) disintegrations per minute (dpm) per unit of volume or mass representing the total concentration of the parent compounds (TNT, RDX or DDT) and all their degradation products. Transfer factors represent the ratio between prey body residue (dpm/mg) and water concentration (dpm/L) and between fish body burden (dpm/mg) and prey body burden (dpm/mg).Bioaccumulation factorPercent of total sum-molar concentrations in worm tissues corresponding to unextractable or extractable parent, known or unknown compounds. Numbers (1–4) represent mean ± 1 standard deviation. Unextractable is defined as compounds that are resistent to solvent extraction from tissue. Extractable compounds includes the parent compound, known, or identified, compounds and unknown compounds, which are more polar than the parent compound and were not identified.This study was supported with funds from the Department of the Army, Environmental Quality Technology (EQT) Program (Dr. M. John Cullinane, program manager). Permission was granted by the Chief of Engineers to publish this material. The authors would like to thank Mr. Cory McNemar and Mr. Henry Banks for providing technical assistance and Drs. Jeffery Steevens and Roderic Millward for their insightful comments on the manuscript.
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+ Phytoremediation is a technology that uses vegetation to remediate contaminants from water, soil, and sediments. Unlike traditional remediation techniques such as soil washing or vitrification, phytoremediation offers a technology that is solar-driven, aesthetically pleasing, and cost effective. Recent studies indicate that winter wheat (Triticum aestivum L.) is a potential accumulator for heavy metals such as lead (Pb) and cadmium (Cd) in hydroponic systems. Based on these findings, a laboratory study was conducted with the primary objective of determining the phytoaccumulation capability of this plant species for heavy metals from contaminated dredged materials (DMs) originating from two confined disposal facilities (CDF). The United States Army Corps of Engineers (USACE) manages several hundred million cubic meters of DMs each year, and 5 to 10 % of these DMs require special handling because they are contaminated with hazardous substances that can move from the substrates into food webs causing unacceptable risk outside CDFs. Phytoremediation may offer an alternative to decrease this risk. Chemical analyses by USACE personnel identified 17 metals in various DMs, but in this present study, only zinc (Zn) and Cd were investigated. Pre-germinated seeds of the test plants were planted under laboratory conditions in pots containing the various DMs and reference soil. Four weeks after planting, plants were harvested and separated into roots and shoots for biomass production and tissue metal concentrations analyses. Results showed that T. aestivum plants have the capacity to tolerate and grow in multiple-metal contaminated DMs with the potential of accumulating various amounts of Zn and Cd. Root and shoot biomass of T. aestivum were not significantly affected by the DMs on which the plants were grown suggesting that this plant species can grow just as well on DMs contaminated by various metals as in the reference soil. No significant differences in the Zn tissue concentrations were observed, differences in Cd tissue concentrations were noted. A maximum concentration of 26 mg Cd kg−1 DW was detected in T. aestivum shoots. Although Cd tissue concentrations of T. aestivum plants in this study were below the Cd plant hyperaccumulation criterion of >100 mg kg−1 Cd found in other studies, this plant species however may still have beneficial uses for phytoremediation studies. T. aestivum plants may serve as an indicator plant for environmental assessment and management, in which the concentration of heavy metals (e.g. Cd) mirrors the concentration in the substrate without dying due to phytotoxicity at low metal concentrations.Sediments from waterways, i.e., dredged material (DM), are often contaminated with several pollutants that enter the waterway via point (e.g., spills and industrial discharges) and non-point (e.g., surfaces runoff) sources. Due to these past and present pollutions, an increasing amount of DM is not available for beneficial use such as beach nourishment, habitat creation and restoration, landfill cover, and land site remediation. The United States Army Corps of Engineers (USACE) manages several hundred million cubic meters of DM each year. Five to ten percent of these DMs require special handling because they are contaminated with hazardous substances that can move from the substrates into food webs causing unacceptable risk outside confined disposal facilities (CDFs) [1]. An emerging technology known as phytoremediation uses vegetation to remove pollutants from water, soil, and sediments. Current remediation techniques involve decontaminating the DM by means of mechanical, chemical, thermal, or biological processes, or any combination of them. Phytoremediation holds a better promise than current remediation practices for effective cleanup of hazardous waste sites because it is more cost-efficient and aesthetically pleasing. Phytoremediation may offer an alternative to decrease the environmental risk of contaminated DM.This laboratory study focused on the phytoextraction capability of a potential phytoaccumulator cultivated on DMs from two CDFs located in Bayport, Wisconsin and Monroe, Michigan (United States). Recent studies indicate that winter wheat (Triticum aestivum L.) [2], plants are potential accumulators for toxic heavy metals such as lead (Pb) and cadmium (Cd) in hydroponic systems. The current study was designed such that seed germination, fertilization, and watering required minimal care and maintenance. For phytoextraction to be a viable technology for industry, technology vendors and/or stakeholders, the contaminant must be available for uptake by the plant roots. Also, the root uptake and subsequent translocation of the heavy metal to the shoot is important, in that it eases harvesting and export of the aboveground plant material from the site [3]. Moreover, the success of phytoremediation of metals depends upon a plant’s ability to tolerate and accumulate high concentrations of the metals at stake, while producing a large plant biomass [4].The specific objectives of this study were to determine the following: 1) the total and bioavailable concentrations of the heavy metals of interest in the five DMs and in the reference soil, 2) the dry biomass yield of the test plant grown in the various substrates and, 3) the test plant’s effectiveness for accumulating high levels of heavy metals in various parts of the plant’s biomass (roots and shoots).This study was conducted using a randomized complete block design with six blocks (Bayport 1, Bayport 2, Bayport 3, Monroe 4, Monroe 5, and the reference soil) with four replicates per block of the test species. The study included a total of 24 experimental units.Substrates in this study are defined as the plant growth media. The test substrates were the DMs from two CDFs located in the Great Lakes Area. Baccto® Lite (obtained from Huttos Garden Center, Jackson, MS) was used as the reference soil to validate the performance of the test plant (Table 1). In this experiment, the reference soil was used as a plant control substrate to validate the performance of the test plant. The DMs were collected and shipped by USACE personnel. Best, Tatem, and Winfield [1] reported on the chemical and physical analyses of these DMs (Tables 2 and 3). Dredged materials were shipped to Jackson State University’s Botanical Laboratory and stored in sealed air-tight plastic containers. The DMs were dried at room temperature to reach a 50% moisture level. Moisture level was measured using a soil moisture meter (LI-COR Inc., Lincoln, NE). The DMs were mixed with vermiculite (obtained from Huttos Garden Center, Jackson, MS) to increase soil aeration [6]. The mixture was a 9:1 DM/vermiculite ratio.Seeds of T. aestivum were purchased from the Rainbow Whole Food Store in Jackson, MS. To obtain a sufficient amount of biomass from T. aestivum, ten plants were grown and harvested in the various substrates for four weeks in the laboratory. The laboratory conditions were monitored. The mean temperature was 24.7 ± 0.56°C (mean ± SE), the mean light was level 111.96 ± 2.46 mol photons m−2 s−1 (mean ± SE) for the entire study period, with a photoperiod of 16 hours. Light was measured from the top of the plants’ canopy several times throughout the experiment using a light energy meter (Biospherical Instruments Inc. Model QSL-100). The light source was composed of four Ott-Lite® F96 80 watt fluorescent bulbs.The plants were grown in 400 ml porous bottom pots (7.5 mm × 8 mm) each with its own 10.16 cm planting reservoir trays containing the needed solutions. The prepared DM/vermiculite mixtures were placed in the pots (350 g per pot). T. aestivum seeds were soaked in deionized water for 24 hours at room temperature before planting. After soaking, twenty seeds were initially planted in each pot and covered by a thin soil layer. After germination, T. aestivum was thinned to ten plants per pot on day 17. Thinning was done to ensure that a sufficient amount of dried biomass would be attainable for the acid digestion procedure and to avoid crowding. The plants were watered every other day or as needed with distilled water and once a week with 100 ml of Hoagland’s solution [8]. The plants were harvested after four weeks of cultivation. They were cleaned, washed successively with deionized water and a weak aqueous solution of sulfuric acid (1 ml of concentrated sulfuric acid per liter of deionized water) to remove any external heavy metal residue. The plants were then separated into roots and shoots, and placed in marked brown bags. The bags and contents were dried in a Blue M Electric Company (Blue Island, Illinois) convection laboratory oven at 100°C for 24 hours. Plant parts were removed from bags and weighed using a balance (XE-100, Arvada, CO) to obtain root and shoot dry weights.This method was used to quantify the biologically available fractions of heavy metals in the DMs and reference soil using a mild extractant, DTPA [9]. The DTPA mixture was composed of 0.005 M of DTPA, 0.01 M CaCl2, and 0.1 M triethanolamine (TEA) all buffered at a pH of 7.3 using 0.1 N and 1.0 N HCl. The moist weights of the DM and reference soil were determined before oven drying. Twenty-grams of each oven-dried substrate were placed in 500 ml centrifuge bottles. Then, 50 ml of DTPA mixture was added to the centrifuge bottle and amended with deionized water to bring the volume up to 100 ml. Reagent blanks were prepared to determine if any contamination was detected from the glassware, reagents, or other sources. The mixture was shaken for 24 hours at room temperature and then centrifuged at 9,000 rpm (13,702 g) for 30 minutes, with 5-minute acceleration and 30-minute deceleration. The supernatant was filtered through a Whatman No. 42 filter paper in a Buchner funnel, and then analyzed for heavy metals lead (Pb), zinc (Zn), and cadmium (Cd) using the Atomic Absorption Flame Emission Spectrophotometer (AA Model 6701F, Shimadzu, Japan).After determining the shoot and root dry biomass, shoot and root tissues were pooled for each experimental group and used for tissue metal extraction and analysis. The United States Environmental Protection Agency (USEPA) Method 3050A [10] was used to extract the metals in the plant materials and the substrates. Substrate concentrations obtained from the acid digestion procedure were used for the total substrate metal concentrations. Acid digestion procedures for the plant materials were carried out separately from the substrate digestion procedure. Reagent blanks were used to determine the contamination, if any, from glassware, reagents, or other sources. To perform the plant metal extraction, 0.1 – 0.5 g of each plant subsample was transferred to a 125 ml Erlenmeyer flask. For the substrate metal extraction, 1.0 – 2.0 g of the substrate was transferred to a 125 ml Erlenmeyer flask. The flasks were then amended with 15 ml of 100 % nitric acid (HNO3) and 10 ml of deionized water. The samples were then heated on a hot plate for 45 minutes at medium heat. The samples were allowed to cool and after adding 5 ml of 100% HNO3, the sample was refluxed again for 30 minutes. The last step was repeated to ensure complete oxidation. The sample was then heated, without boiling and evaporated to 5 ml. After this, the samples were allowed to cool again, and 2 ml of deionized water were added along with 3 ml of 30% hydrogen peroxide (H2O2) to each sample. The samples were then heated to start the peroxide reaction. The 30% H2O2 were continually added in 1ml aliquots until the effervescence became minimal. The acid-peroxide digestate was heated for a final time to reduce the volume to 5 ml. After cooling, the samples were diluted to 100 ml with deionized water. The digestate was filtered using a Whatman No. 1 filter to remove any particulates that may have been present in the sample. A different filter and sulfuric acid-cleansed funnel was used for each sample to avoid cross-contamination of the samples. The filtrate was then ready for metal analysis.Metal concentrations were determined using an atomic absorption flame emission spectrophotometer (AA Model 6701F, Shimadzu, Japan). Fresh standards of 1, 5, and 10 parts per million (ppm) Pb; 0.1, 0.2, and 0.4 ppm Zn and 0.2, 0.4, and 0.6 ppm Cd concentrations were prepared for the atomic absorption analysis. The solutions were aspirated and their concentrations were calculated and recorded. The Zn and Cd concentrations in the DM, reference soil, and plant materials were measured.The data were analyzed using analysis of variance [11]. The metal concentrations of the five DMs were compared with the metal concentrations of the reference substrate using Dunnet’s multiple comparison tests. The Duncan’s Multiple Range Test was used as a mean separating procedure for the total and bioavailable metal concentrations. All total and bioavailable metal concentrations were based on the measurements derived from the substrate analysis protocol in this study. The block effects were examined to determine if there were significant variations in responses due to situation of the planted pots within the blocks. The criteria for statistical differences were determined at both the 5% (p-value ≤ 0.05) and the 10 % significance level (p-value ≤ 0.1).The relationship between plant responses (tissue metal concentration and biomass production) and substrate metal concentrations (total and bioavailable) were derived by linear regression. The p-value in the regression model was set at a 5 % significance level (p-value ≤ 0.05). The R2 value of the regression model was used to indicate the explained variance in the model, whereby R2 values of ≥ 0.50 (explaining at least 50 percent of the variance in the data set) was considered meaningful. Linear regression equations were used to predict the tissue metal concentrations resulting from the substrate metal concentrations. The biota to soil accumulation factor (BAF), which is the ratio between the shoot metal concentration in the various plant species and the substrate metal concentration, was also calculated as an estimate for the potential trophic transfer of the metal of interest from the DM into the plants [7].No symptoms of phytotoxicity, such as chlorosis or necrosis, were noticeable from the plant species under investigation. The level of Pb in the plant biomass was below detection. This may be due to the binding of Pb by the substrate in the aged DM, since in the preliminary experiment the test plant species accumulated detectable levels of Pb in both root and shoot biomass (data not shown). However, the levels of Zn and Cd were measurable.Results of T. aestivum biomass are shown in Figure 1. Root biomass ranged from 0.21 ± 0.02 to 0.27 ± 0.01 g DW plant−1. No significant differences were observed in the root biomass of T. aestivum plants grown in the various DMs when compared to the plants grown in the reference soil. The shoot biomass of plants grown on Bayport 1 (0.62 ± 0.03 g DW plant−1) and Bayport 3 (0.50 ± 0.02 g DW plant−1) DMs were significantly lower than those plants grown on the reference soil (0.97 ± 0.06 g DW plant−1) (p ≤ 0.05). The shoot biomass of plants cultivated on Monroe 4 DM (0.67 ± 0.10 g DW plant−1) was significantly lower than that of the reference plants (p ≤ 0.1). Statistically, the total plant mass behaved identical to shoot mass at the 5 % significance level.Zinc accumulation in the root ranged from 69.56 ± 11.01mg kg−1 DW in T. aestivum plants grown on Bayport 1 DM to 94.01 ± 5.34 mg kg−1 DW in plants grown on Monroe 5 DM (Figure 2). Zn accumulation in the shoot ranged from 39.86 ± 3.59 mg kg−1 DW in plants grown on Monroe 4 DM to 59.39 ± 4.87 mg kg−1 DW in plants grown on Bayport 1 DM. However, no significant differences were observed in the root or shoot Zn accumulation by T. aestivum plants grown in the various DMs compared to reference plants. Cadmium accumulation in T. aestivum plants is summarized in Figure 3. Cadmium accumulation in the roots was significantly higher in plants grown in Bayports 1 and 2 DMs than in reference plants. Shoot Cd accumulations were significantly higher in plants grown on Bayport 1 DM, but significantly lower in plants grown on Bayport 2, Bayport 3, and Monroe 4 DMs than in reference plants. Shoot Cd concentrations ranged from 2.46 ± 0.30 mg kg−1 DW in plants grown on Bayport 2 DM to 26.34 ± 0.45mg kg−1 DW in plants grown in Bayport 1 DM. At a lower significance level, (i.e. 10%), significant differences in tissue Cd concentrations were found (Figure 3). No significant relationship between the shoot metal concentrations (Zn and Cd) and substrate metal concentrations were found using linear regression techniques (data not shown). Consequently, BAFs were not calculated. Statistical analysis of the relationship between the total plant biomass production and substrate metal concentrations showed no significant correlation.The total Zn concentrations were significantly higher in Bayport 1, Bayport 2, and Monroe 4 DMs than in the reference soil (p ≤ 0.05) (Table 4). At the 10 % significance level, the total substrate Zn concentration was also significantly higher for the Bayport 3 DM (Table 4). Concentrations of the bioavailable Zn in the substrates in this study ranged from 2.53 ± 0.00 mg kg−1 DW in the reference soil to 6.60 ± 0.05 mg kg−1 DW in Monroe 4 DM. The bioavailable Zn concentrations in all DMs were significantly higher than in the reference soil.Total Cd concentrations in the various substrates ranged from 0.48 ± 0.09 mg kg−1 DW in the reference soil to 1.34 ± 0.04 mg kg−1 DW in Bayport 1 DM (Table 5). No significant differences were found in the total Cd concentrations in the DMs, except for Bayport 1 DM, compared to the reference soil (p ≤ 0.05). Bioavailable Cd concentrations in all the DMs were significantly higher than in the reference soil.All plants can take up metals from the substrate in which they are cultivated at varying degrees. Baker [12] stated that there are two ways in which higher plants can tolerate the presence of metals in their environment:Exclusion, which occurs when the transportation of metals is restricted and low, relatively constant, metal concentrations are maintained in the shoot over a wide range of soil concentrations.Accumulation, which occurs when metals are taken up in a nontoxic form in the shoot at both high and low soil concentrations. He suggested that accumulators can be characterized by a shoot:root metal concentration ratio of >1 due to the tendency to translocate metals from the root to the shoot, whereas excluders are characterized by a ratio of < 1. Baker [12] further stated that an intermediate response of an indicator plant is also likely, whereby the shoot metal concentrations reflect those in the substrate.No significant differences in the Zn tissue concentrations were observed, differences in Cd tissue concentrations were noted in T. aestivum plants. A maximum concentration of 26 mg Cd kg−1 DW was detected in T. aestivum shoots. Although Cd tissue concentrations of T. aestivum plants in this study were below the Cd plant hyperaccumulation criterion of >100 mg kg−1 Cd found in other studies [13], this plant species however may still have beneficial uses for phytoremediation studies. Zaman and Zereen [14] and McGrath et al. [15] concluded that although not capable of hyperaccumulation, T. aestivum plants may pose as an indicator plant, in which the concentration of heavy metals (e.g. Cd) mirrors the concentration in the substrate without dying due to phytotoxicity at low metal concentrations.Substrate-metal binding has been documented in numerous soil chemistry studies [16]. This may explain why the plant tissue Pb accumulation was below detection in the laboratory experiment, although it was demonstrated in the preliminary hydroponic experiment. Studies have shown that Pb can be strongly retained in many soils, thereby hindering Pb mobility into plant tissue [17, 18, 19]. McBride [20] stated that the preeminent way to evaluate soil-metal mobility was to understand the soil properties and conditions that affect the long- and short-term fate of the metals in soils. Ross [21] stated that the exclusion of the large number of influential factors (i.e. bulk density, temperature, aeration, redox potential, pH, and organic matter quantity and quality) and soil interaction was a main problem in McBride’s [20] approach to understanding metal processes in soil. Results from the current study concur with those reported by Best et al. [7] who also failed to detect Pb, but did recover Zn and Ni in plants grown in Monroe CDF-DM.Biomass of Triticum aestivum (mean and standard error; N=4) grown in substrates for 4 weeks. aStatistically significant from the reference soil at the 5 percent significance level; bat the 10 percent significance level according to Dunnet test.Mean concentrations of zinc in Triticum aestivum (mean and standard error; N=4). No significant differences observed when compared to plants grown in the reference soil according to Dunnet test.Mean concentrations of cadmium in Triticum aestivum (mean and standard error; N=4). aStatistically significant from the reference soil at the 5 percent significance level; bat the 10 percent significance level according to Dunnet test.Reference soil profile [7].Bayport dredged material profile [1].Abbreviations: NA, not analyzedPlant available phosphorus-fractionMonroe dredged material profile [1].Abbreviations: NA, not analyzedPlant available phosphorus-fractionTotal and bioavailable zinc concentrations (mean values and standard error, N=3).Significantly different from the reference soil at the 5 percent significance level;significantly different from the reference soil at the 10 percent significance level, according to Dunnet test.Total and bioavailable cadmium concentrations (mean values and standard error, N=3).Significantly different from the reference soil at the 5 percent significance level;significantly different from the reference soil at the 10 percent significance level, according to Dunnet test.This research was financially supported by a grant from the U.S. Department of Education (Grant No. P031B99000601) through the Title III Graduate Education Program at Jackson State University, Jackson, MS. Special appreciation is extended to John Young and Robert Hughes, Trent Lott Geospatial and Visualization Research Center, of JSU for their assistance with graphical presentation.
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+ Lead (Pb) is one of the most toxic metals in the environment and may cause drastic morphological and physiological deformities in Ipomoea lacunosa. The goal of this research was to evaluate some morphological and physiological responses of morning glory grown on a Pb- and chelate-amended soil. Soil samples were analyzed, at Mississippi State University Soil Laboratory, for physico-chemical parameters, such as soil texture (73% sand, 23% silt, 4.4% clay), organic matter (6.24 ± 0.60%), and pH (7.95 ± 0.03), to establish soil conditions at the beginning of the experiments. Five EDTA (ethylenediaminetetraacetic acid) concentrations (0, 0.1, 0.5, 1, 5mM) and four lead (0, 500, 1000, 2000mg/L) treatments were arranged in factorial in a Randomized Complete Block (RCB) design with five replications. Duncan’s multiple comparison range test showed that the mean difference values of stomatal density were significant between 500 and 1000mg/L Pb and between 1000 and 2000mg/L Pb. Two way ANOVA (at 1% level) indicated that interaction between Pb and EDTA had a significant effect on the stomatal density and photosynthetic rates, and at 5% level Pb had a significant effect on chlorophyll concentrations. Lowest concentrations of chlorophyll were recorded at 2000mg/L Pb and 5mM EDTA and exhibited a decreasing trend specifically in the ranges of 1000 and 2000mg/L Pb and 1.0 and 5.0mM EDTA. Duncan’s multiple comparison range test confirmed that mean differences between the control treatment vs. 2000mg/L Pb, and 500mg/L vs. 2000mg/L Pb were significantly different at p>0.05. There was a decrease in leaf net photosynthetic rate with increasing concentrations of Pb from 0 to 2000mg/L. In conclusion, I. lacunosa L. plants were grown to maturity in all treatments with no significant and/or apparent morphological disorders, which indicated that this species might be highly tolerant even at 2000mg/L Pb concentrations in the soil.Lead (Pb) has been used in various forms by ancient civilizations; hence most of its contamination to our environment is anthropogenic. Due to its history as an air emission pollutant, lead has been fairly mobile and is particularly soluble in acid environments [1]. Different plants have different tolerance limits for a variety of elements (including the trace elements). The toxic effect of heavy metals such as Cu, Pb, Cd, Ni on photosynthesis and plant metabolism is well documented [2]. The effects of Pb treatment on thylakoid development in Populus alba (poplar) and Cucumis sativus (cucumber) plants and growth responses in radish (Raphanus sativus) plants are documented in the scientific literature [3, 4]. Decreased levels of chlorophyll were observed in Brassica oleracea (cabbage) exposed to <2.5mg/L of copper [5]. For zinc and copper, most vascular plant species exhibited toxicity symptoms at low concentrations ranging from 0.1 to 2.0 mg/L zinc and from 0.02 to 0.06 mg/L copper [6]. Significant reduction in chlorophyll concentrations were reported in most metal (cadmium and lead) treated radish plants [4]. Duckweed plant exhibited some symptoms of toxicity such as reduced growth and chlorosis at higher levels of metal supply, except for Cr [7].The specific objectives of this study were to evaluate some morphological (necrosis, chlorosis, curling of leaves, stunted growth) and physiological (chlorophyll concentrations and photosynthetic rates) responses of morning glory grown on a Pb- and chelate-amended soil. Data generated will be utilized to further assess the suitability of this species for phytoextraction of Pb-contaminated soils.Seeds of morning glory (Ipomoea lacunosa L.) were obtained from Azlin Seeds, Leland, MS. Plants were grown in the greenhouse at Jackson State University. Several bags (~18.5kg/bag) of potting mix and delta topsoil were purchased from Hutto’s Garden, Jackson, MS and air-dried for 3–4 days under greenhouse conditions. Based from a preliminary experiment, germination and growth of plants were highly responsive to a growth medium consisting of 2/3 potting mix and 1/3 delta topsoil. Plastic pots (1.9L) were filled with approximately two kg of growth medium. A 7-inch plastic saucer was placed beneath each pot to prevent cross contamination among treatments. Any leachate collected in each saucer was poured back into its corresponding pot. Periodically, these saucers were also rinsed with deionized distilled water and were poured back into the respective pots.Soil samples were analyzed, at Mississippi State University Soil Laboratory, for physico-chemical parameters, such as soil texture (73% sand, 23% silt, 4.4% clay), organic matter (6.24 ± 0.60%), and pH (7.95± 0.03) to establish the soil conditions at the beginning of the experiments. Five EDTA (ethylenediaminetetraacetic acid) concentrations (0, 0.1, 0.5, 1.0, 5mM) and four lead (0, 500, 1000, 2000mg/L) treatments were arranged in factorial in a Randomized Complete Block (RCB) design with five replications. For each pot, 250mL of the respective concentrations of lead and EDTA were added as aqueous solutions prior to sowing. Five germinated (~5mm radicle) seeds of I. lacunosa L. were planted in each pot and were thinned to two plants per pot one week after planting. Plants were periodically watered with 100mL of deionized distilled water for the first two weeks. Beginning at third week of growth, all plants were watered twice a week with 100mL of modified Hoagland’s nutrient solution [8]. When the plants were approximately 4–5 weeks old, 250mL of nutrient solution were added 3–4 times a week until the time of harvest. Plants were harvested at 7 weeks after emergence (i.e., at the time of flowering and fruiting).Leaf samples were collected randomly from each treatment and immersed in test tubes filled with water. Experimental peels were obtained from the lower surface of the leaves, prepared as wet mounts and were viewed through an Olympus AH-2 VANOX microscope under 40*10X magnification to obtain stomatal density. Pictures of the stomatal density in each leaf material under different treatments per focal area (frame) were taken. Leaf samples of L. lacunosa were collected from different treatments using #4 (7mm diameter) cork borer. A total of ten discs from each treatment and two discs from each leaf were collected for chlorophyll analysis. Leaf discs were obtained from the 4th node, 10th node, and from the uppermost node of the plant. Two leaves from each node and one leaf from the top, uppermost node of the plant were selected. Chlorophyll from leaves was extracted following a standard method [9]. Net photosynthetic rates (μmolCO2 m−2s−1) of selected leaves of L. lacunosa were measured using a LiCOR-6200 portable photosynthesis system (LI-COR, Inc., Lincoln, NE). Leaves of similar morphology and position on the stem were used in each measurement. Photosynthetic measurements were made between 10:00 to 13:00 hours (CDT) when the prevailing light intensity was >1600 [μmol photons m−2s−1 (PAR, 400–700nm)].Data were analyzed using ProStat software (version 1.04 for windows 95, 32 bit edition, 1996) to determine means, standard deviations (SD), standard errors (SE), two-way ANOVA with replications, multiple comparison tests (Duncan’s Test), and correlations.Stomatal densities on leaf surfaces of I. lacunosa L. were significantly affected by different Pb and EDTA treatments. Results on stomatal density are summarized in Table 1 and in Fig. 1A. Minimum number of stomata was observed under 1000mg/L Pb and 5mM EDTA. Duncan’s multiple comparison range test showed that the mean difference values of stomatal density were significant between the 500 and 1000mg/L and 1000 and 2000mg/L Pb treatments in 1 and 5mM EDTA amended soils. The two way ANOVA test also confirmed the differences between Pb treatments, and the interaction between Pb and EDTA for stomatal density was significant at the 1% level.The effects of Pb and EDTA concentrations on total chlorophyll contents of L. lacunosa leaves are presented in Fig. 1B and Table 2. The two way ANOVA test indicated that Pb had a significant effect on chlorophyll concentrations at the 5% level of significance. Lowest concentrations of chlorophyll were recorded at 2000mg/L Pb and 5mM EDTA and exhibited a decreasing trend specifically between 1000 and 2000mg/L Pb and 1.0 and 5.0mM EDTA. The Duncan’s multiple comparison range test confirmed that mean differences between the control treatment vs. 2000mg/L Pb, and 500mg/L vs. 2000mg/L Pb were significantly different at p<0.05.Results indicated that there was a decrease in leaf net photosynthetic rate with increasing concentrations of Pb from 0 to 2000mg/L (Fig. 1C, Table 3). Two-way ANOVA results indicated that the effect of EDTA, Pb concentrations, and the interaction between Pb and EDTA on net photosynthesis were highly significant at the 1% level. Results also showed that within each Pb treatment, there was a decreasing photosynthetic rate as EDTA concentration increased. Levitt (1980) described metals in the environment as stress factors in that they cause physiological reaction changes which may result in the total inhibition of plant growth. He also indicated that resistance to heavy metals can be achieved by tolerance mechanisms which allow a plant to survive the effects of internal stresses associated with such heavy metals.Tolerance is conferred by the possession of special physiological mechanisms, which collectively enable a plant to function normally even in the presence of high concentrations of potentially toxic elements [10]. Ipomoea lacunosa L. may be one of these metal-tolerant plant species, which in our experiments survived and continued normal growth and development even at 2000mg/L Pb and 5mM EDTA amendments to the soil. Pb-chelates have been shown to accumulate in interveinal areas of dicot leaves specifically in the anticlinal walls of the leaf epidermis [11]. It is possible that this Pb-chelate deposition may have direct toxic effects on stomatal ontogeny, thereby reducing stomatal density. A decrease in stomatal density can lead to reduced stomatal conductance. Reduced stomatal conductance has been shown in most cases to reduce photosynthetic rate because of a reduction in the supply of CO2 for the rubisco enzyme [12, 13]. Duckweed has been shown to exhibit some deleterious consequences of toxicity such as chlorosis at higher levels of Pb, Cu, Cd, and Ni [7]. There were no visible signs of metal-induced toxicity (chlorosis) in the present study even though there was a decrease in leaf total chlorophyll contents at 1000 and 2000mg/L Pb treatments in combination with 1 and 5mM EDTA (Figure 1B). In a similar study, it was found that there was no significant reduction in total chlorophyll content in Pb-treated (500–1000ppm Pb) radish plants [4].The effects of toxic metals on chlorophyllase (an enzyme which decomposes chlorophyll) activity is significant. Some metals such as Cd, Pb, Mn, Co, Ni and their mixture enhanced the activity of chlorophyllase on chlorophyll-a decomposition more than that on chlorophyll-b [14]. Other studies have also reported Pb poisoning of leaves leading to chlorosis in plants growing on metalliferous mine sites [15, 16]. The direct effects of Pb or oxidative stress due to the increased iron concentration may be responsible for changes in thylakoids of cucumber, while the Pb (at higher concentrations)-induced Mn deficiency may be connected with stress-induced changes in poplar plants. If stomatal density and chlorophyll contents are decreased at high Pb levels, there is reduced stomatal CO2 conductance and PAR absorbance by leaves [3]. Reduced stomatal conductance has been shown in most cases to reduce photosynthetic rate because of less CO2 for the rubisco enzyme [12, 13, 17]. Carbon dioxide is one of the major factors that dictates photosynthetic rate in plants, and directly depends upon the stomatal aperture and density. In the present study, stomatal density was reduced at higher concentrations of Pb and EDTA, which might have had an impact on the availability of CO2 and the associated depression in leaf net photosynthetic rates at these levels of Pb and EDTA. Results of the present study show the link between stomatal density, total chlorophyll, and leaf net photosynthetic rate (Fig. 1A–C), all of were low at the higher concentrations of Pb (1000, 2000mg/L) and EDTA (1.0, 5.0mM). This indicates that Pb had either a direct or indirect toxic effect resulting in the reduction of number of stomata, leaf chlorophyll content, and leaf net photosynthesis, by itself or in association with other elements (synergistic effect) in the nutrient medium, which could not be determined based on current research results. Metals such as Cu, Pb, Cd, and Ni have an effect on metal-sensitive processes such as photosynthesis [2]. A previous study also demonstrated that Pb depressed photosynthesis in corn and soybean but did not indicate clearly the mechanism of Pb toxicity on photosynthesis [18].Ipomoea lacunosa L. plants were able to grow to maturity in all treatments with no significant and/or apparent morphological disorders. This indicated that this species is highly tolerant to lead even at 2000mg/L Pb concentrations in soil. With its demonstrated tolerance to moderate levels of Pb, this species can be further evaluated for its suitability as a potential species for phytoextraction of Pb-contaminated soils.Effects of various concentrations of Pb and EDTA on (A) stomatal density, (B) total chlorophyll, and (C) the net photosynthetic rate of Ipomoea lacunosa (L.). Vertical bars indicate standard error of the mean of 4 replications.Effects of Pb and EDTA concentrations on stomatal density (no./mm2) of I. lacunosa L. leaves; two-way ANOVA with replication.significant at 1% level.Effects of Pb and EDTA concentrations on total chlorophyll (μg/mg dry weight of leaf) of I. lacunosa L.; two-way ANOVA with replication.significant at 5% level.Effects of Pb and EDTA concentrations on leaf net photosynthetic rate (μmol CO2m−2sec−1) of I. lacunosa L. leaves (two-way ANOVA with replication).significant at 1% level.This research was supported in part by the Jackson State University and Title III Program at Southern University at New Orleans.
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+ DNTs are considered possibly carcinogenic to humans (Group 2B) because there is inadequate evidence in humans for carcinogenicity though there is sufficient evidence in experimental animals. In this study, MCF-7 (breast) and MRC-5 (lung) cells were exposed to a serial dilution of 2,4 and 2,6 DNTs (control, 1–500 ppm) in 96 well tissue culture plates. After various time intervals (24, 48, 72 and 96 hrs) the plates were washed, and 100 μl fluorescein diacetate solution (10 μg/ml in PBS) was added column wise to each well, and incubated at 37°C for 30–60 min before reading the fluorescence with a spectrofluorometer at excitation and emission wavelengths of 485 and 538 nm respectively. Spectrofluorometeric readings were converted to percentages of cell survival. Regression analysis was conducted to determine the relationship between cell survival and exposed concentration. Linear equations derived from the regression analysis were used to calculate the LC 50 values. Results indicated that 2,6 DNT as more toxic to breast cells; LC50 values were 445 and 292 ppm at 24 and 48 hours respectively compared to 2,4 DNT showing LC50 values of 570 and 407 ppm at 24 and 48 hours, respectively. No significant differences in toxicity existed between the two chemicals with regard to lung cells. Contrary to the above observation, 2,4 DNT was more toxic to breast cells; LC50 values were 407 and 238 ppm at 24 and 48 hours respectively compared to lung cells showing LC50 values of 527 and 402 ppm at 24 and 48 hours respectively. No significant difference existed for 2,6 DNT between the two cell lines. Lungs cells were more resistant to the two chemicals.DNTs (2,4 and 2,6) are produced through dinitration of toluene with nitric acid in the presence of concentrated sulfuric acid [1]. They are used in munitions as smokeless propellant powders, as gelatinizing and plasticizing agents in commercial and military explosive compositions and in the manufacture of dyes [2]. During the production of Trinitrotoulene (TNT) small amounts of DNT isomers also occur as byproducts [3; 4]. Leaching of wastewaters from disposal sites that contain significant amount of DNT arising from their preparation and production at Army ammunition plants have been identified in surface water, groundwater and in soils [5].Major exposure routes of DNTs (via ingestion, inhalation, and dermal contact) have been associated with a significant number of health effects [6]. Exposure to nitroaromatic compounds intially may result in mild irritation of respiratory passages producing nasal discomfort, sneezing, epistaxis, and rhinitis, as well as irritation of the skin producing erythema and papular eruptions progressing to desquamation and exfoliation [1]. Absorption of sufficient amounts of DNTs through the skin or lungs has been published else where in literature. These exposure routes may produce signs of cyanosis due to methemoglobin formation, toxic jaundice, and hepatitis due to severe liver damage, aplastic anemia due to damage to the erythropoietic system, eye damage (cataract), menstrual disorders (hypo- or hypermenorrhea), neurological manifestations (neurasthenia, nystagmus, and irregularities in tendon reflexes), and kidney damage [3; 6, 7]. On the average 500 workers in the USA are exposed to DNTs when they are used as intermediates in the production of munitions and explosives [1].DNTs are possibly carcinogenic to humans because there is inadequate evidence in humans for carcinogenicity though there is sufficient evidence in experimental animals [8]. It has been reported that, chronic exposure to 2,4-DNT induces significant number of cancers in laboratory animals including adenoma, fibroma, hepatocellular carcinoma, and mammary tumors in rats [9]. In the reversion assays with Salmonella typhimurium, with or without metabolic activation, TNT was found to be strongly mutagenic, while 2,4-DNT and 2,6-DNT were weakly mutagenic. [10]. Based on these laboratory and other studies, the U.S. Environmental Protection Agency considers DNTs as probable human carcinogens-Group B2 [11].In this work we exposed DNTs to breast and lung cells to determine the cytotoxic effects of DNTs. These cells were chosen due to their differences in morphology, age, ploidy and characteristics [12].MCF-7 (breast) and MRC-5 (lung) cells were grown with DMEM supplemented with 10% fetal bovine serum (FBS) and 1% streptomycin and penicillin at 37°C in a 5% CO2 incubator to 90–100% confluence. The old medium was removed and cells washed with phosphate buffer saline (PBS), trypsinized with about 4 mL of 0.25% (w/v) trypsin-0.03% w/v) EDTA. Fresh medium was added and centrifuged in 50 mL conical tubes at 3000 RPM for 10 minutes. Cells were incubated for 24, 48, 72 and 96 hours at ± 37°C in 5% CO2. Plates were washed once with 200 μl PBS/well. Hundred micro-liters (μl) of Fluorescien Diacetate solution (10 μg/ml in PBS) was added columnwise to each well and incubated at 37°C for 30–60 minutes before reading the fluorescence with a Spectrofluorometer at excitation and emission wavelength of 485 and 538 nm respectively. Spectrofluorometer readings were converted to cell survival. Regression analysis was conducted between cell survival and exposed concentrations. Linear equations derived from the regression analysis were used to calculate the LC 50’s.Figure 1 is a regression analysis conducted on the breast cells exposed to different concentrations of 2, 4 DNT. The graph shows an inverse correlation between 2, 4 DNT concentration and percent cell survival. It was observed that below 100ppm of 2, 4 DNT exposure, almost all the cells survived and at 500ppm of 2,4 DNT exposure almost 100 % cell death occurred. Similar observation were made when the breast cells were exposed to 2,6 DNT. With the lung cells exposure to both 2,4 and 2,6 DNT, the regression analysis yielded comparable results to the ones obtained for the breast cells.Figure 2 depicts the lethal concentration (LC50) for breast cells exposed to DNT (2,4 and 2,6). The figure shows the LC50 were time dependent. The highest LC50 was observed at 96 hours (138.4 ppm for 2,4 DNT and 121.1 ppm for 2,6 DNT) and the lowest at 24 hours (569.9 ppm for 2,4 DNT and 445.3 ppm for 2,6 DNT).Figure 3 compares the LC50 for lung cells when exposed to 2,4 DNT and 2,6 DNT. At 24 hours the LC50 were 622 and 663 ppm for 2,4 DNT and 2,6 DNT respectively. At 96 hours, the LC50 were 265 and 192 ppm for 2,4 DNT and 2,6 DNT.The LC50 for both lung and breast cells were compared to see if there was any difference between the two cells when exposed to 2,4 DNT. Figure 4 shows that there was a significant difference between lung and breast cells (p>0.05): 2,4 DNT was more toxic to breast cells than to lung cells.Similar comparison was made between lung and breast cells when exposed to 2,6 DNT (figure 5). Over all the result yielded no significant difference between the two cells though the LC50 at 24 hours was higher for breast cells than for lung cells.It has been reported by Tchounwou et al[13] that 2,4-DNT was more toxic than 2,6-DNT when exposed to Hep G2 cells. This observation was contrary to what was observed in this study. We observed in this study that 2,6 DNT was more toxic to breast cells compared to 2,4 DNT and with the lung cells no significant differences in toxicity existed between the two chemicals. The possible explanation between these results might be due to the apparent biological differences among the cells. We also observed that 2,4 DNT was more toxic to breast cells than to lung cells and no significant difference existed for 2,6 DNT between the two cells in this study. In summary, lungs cells were more resistant to the two chemicals. Again, the possible explanation for the observed differences between lungs and breast cells might be due to the apparent biological differences between the two cells. There were major differences between the two cells in terms of morphology, age, ploidy and characteristics. The lung cell is a fibroblast, embryonic and diploid while the breast cell is epithelial, from adult donor and aneuploid. The breast cell is estrogen receptor positive while the lung cell is not. The breast cell is continuous and from neoplastic tissue while the lung cell is finite and from normal tissue [12].Much of the acute toxicity test with DNT had been done on rodents. It has been reported by Vernot et al.[14] that the LD50 for 2,6-DNT and 2,4-DNT were 1,000mg/kg and 1,630mg/kg, respectively for CF-1 mice. They also reported that the LD50 for male Sprague Dawley rats for 2,6-DNT and 2,4-DNT as 180 mg/kg and 270 mg/kg, respectively.Levine et al. [15], reported the LD50 for 2,4-DNT exposures to male and female Swiss mice, as 1,954 mg/kg and 1,340 mg/kg, respectively. They also reported the LD50 for male and female CD rats exposed to 2,4-DNT as 568 mg/kg and 650 mg/kg, respectively. With 2,6-DNT, the LD50 for male and female Swiss mice were 621 mg/kg and 807 mg/kg respectively; for male and female CD rats the LD50 were 535 mg/kg and 795mg/kg respectively. From the above published results, 2,4 DNT was more toxic than 2,6 DNT.Our results indicated that 2,6 DNT was more toxic to breast cells compared to 2,4 DNT. LC50 values were 445 and 292 ppm at 24 and 48 hours respectively for 2,6 DNT, and 570 and 407 ppm at 24 and 48 hours respectively for 2,4 DNT. No significant differences in toxicity existed between the two chemicals with regard to lung cells. Contrary to the above observation, 2,4 DNT was more toxic to breast cells; LC50 values were 407 and 238 ppm at 24 and 48 hours respectively compared to lung cells showing LC50 values of 527 and 402 ppm at 24 and 48 hours respectively. No significant difference existed for 2,6 DNT between the two cell lines. Lung cells were more resistant to the two chemicals. The observed toxicity differences between the lungs and breast cells were contrary to our expectation since it was expected that the breast cells would be more resilient to the chemicals than the lung cells.Regression analysis for breast cells exposed to different concentrations of 2,4 DNT.LC50 values of 2,4 and 2,6 DNT for breast cells exposed at various times.LC50 values of 2,4 and 2,6 DNT for lung cells exposed at various times.LC50 values of 2,4 DNT for breast and lung cells exposed at various times.LC50 values of 2,6 DNT for breast and lung cells exposed at various times.This research was done through the MACERAC Research and Education Program financially supported by a grant from the U.S. Department of the Army (Grant #DACA-42-02-C-0057). Frederick Ballard and Christine Timmons were financially supported by MESIP. Many thanks to Dr. Kelly Mack and Dr. Joseph Okoh (Chair), Department of Natural Sciences, UMES for their support and advice on this project.
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+ Soil metal surveys were conducted in Baltimore, MD (1976–1979), Minnesota (1981–1988) and most recently, New Orleans, LA (1989-present). The unique characteristic of New Orleans is that it has two surveys; Survey I was completed in 1992 and Survey II was completed in 2000. This paper seeks to determine if there is a perceptible change in the amount of metals during less than a decade that separated these surveys. The Survey I collection was 4,026 samples stratified by 283 census tracts. All samples were collected in residential neighborhoods at least one block from a busy street. The Survey II collection was 5,467 samples stratified by 286 census tracts (plus City Park). The Survey II collection included busy streets as a category of samples. For comparison, the busy street category of 1,078 samples was excluded from Survey II for a total of 4,388 samples. The extraction methods of the two surveys used the same protocol for strength of acid (1 M HNO3), shaker-time (2 hours), and room temperature (~22ºC). However, Survey II differed in amount of sample used in extraction. For Surveys I and II, 4.0g and 0.4g were used respectively. The same ICP-AES was used to measure 8 metals in both surveys. To evaluate the analytical results of the two methods, reference soil samples (n=36) from the Wageningen Evaluating Programs for Analytical Laboratories, International Soil-analytical Exchange (WEPAL; ISE) were used. The relationship between the 4.0 and 0.4 g results were linear and the Survey I results were adjusted for sample:acid ratio. Further evaluation was done by creating interpolated Multiple Metal Accumulation (MMA) maps based on the median MMA for each census tract. A new map was created by dividing Survey II MMA by Survey I MMA. The ratio indicates increases of soil metals in the inner city and decreases of soil metals in the outlying areas of Metropolitan New Orleans. Comparing fresh parent alluvium from the Mississippi River with urban soil metal quantities demonstrates that the soils of New Orleans have undergone a massive accumulation of metals. The preliminary results provide ideas about methods needed to further evaluate the changes between these surveys.The New Orleans Soil Surveys were a natural extension of previous urban soil studies first reported for Baltimore, MD [1]. The amount of lead and other metals were measured in 454 garden soils from the entire metropolitan area. Significant differences that could not be explained by chance (p-values as low as 10−23) were identified between soils in the inner city compared with those collected from outlying areas of Baltimore [1]. The unpainted brick buildings of Baltimore did not support the hypothesis that lead-based paint alone explained the observed pattern. We hypothesized that the use of leaded gasoline was a major contributor to the accumulation of lead in Baltimore. Further research was conducted in Minnesota and the pattern first observed for Baltimore was also found in the Twin Cities [2]. Another major empirical finding in Minnesota was that the amount of soil lead is directly related to size of the city [3, 4]. The larger the size of the city the larger the amount of soil lead observed in the city. In Louisiana the studies conducted in Baltimore and Minnesota were repeated. Survey I of New Orleans was conducted between 1989 and 1992 and was the basis for several papers [5–13]. A major finding was that soil lead was a better predictor of the prevalence of childhood lead poisoning than age of housing [9, 12]. A soil lead and blood lead survey conducted in Syracuse, New York revealed a similar finding [14]. Survey II was started in 1998 and completed in 2000 and is the basis for papers on multiple metal mixtures [15–18] as well as mixtures of metals and organic hydrocarbons in the urban environment [16, 19, 20]. The purpose of this paper is to evaluate the differences between the New Orleans surveys that were separated by less than a decade. The main question is whether or not the two surveys show a perceptible change in the amount of metals accumulated in New Orleans.All soil samples were collected from the top 2.5 cm of the soil. Most aspects of the soil surveys were the same. The differences in treatment are described below:The 1980 Census Tracts were used to stratify the samples (N = 283). Maps were prepared by hand from a New Orleans street map marking the boundaries of each census tract prior to conducting the field collection. In the field, wherever possible 15 samples per census tract were collected according to the following scheme. All samples were collected at least one block from a busy street and away from street corners. Ten samples were collected within 1 m of the street (street side), 3 samples, each matched with street side samples, were collected within 1 m of house sides (foundation), and 2 samples were collected from vacant land or parks as far as possible from streets and house sides (open space samples). The total number of samples collected for Survey I was 4,026. Soil extractions were prepared using 4.0 g of soil with 20 ml trace metal grade 1M HNO3.The soil samples were stratified by 1990 Census Tracts (N = 286) plus City Park. Wherever possible, 19 samples per census tract were collected. Sampling was conducted by using U.S. Census Bureau TIGER 95 maps as a guide. Ten samples were collected within 1 m of the street (street side), 4 samples collected within 1 m of busy streets (busy street side), 3 samples, each matched with street side samples, were collected within 1 m of house sides (foundation), and 2 samples were collected from vacant land or parks as far as possible from streets and house sides (open space). The total number of samples collected for Survey II was 5,467. To compare the two surveys the busy street samples collected in Survey II were excluded as a sample category. The busy street category of 1,072 samples was removed and the total number of 4,388 soil samples was used in this study. Soil extractions were prepared with 0.4 g of soil and 20 ml trace metal grade 1M HNO3.The extraction methods of the two surveys followed the same protocol for acid pH (1 M HNO3) shaker time (2 hours) and room temperature (~ 22 ºC). The two surveys differed in amount of sample, 4.0 g (Survey I) compared with 0.4 g (Survey II). ICP-AES techniques were used to measure 8 metals. The soil and acid were shaken for 2 hours in polypropylene cups and then filtered using Fisherbrand P4 paper. The extract was stored in 20 ml polypropylene scintillation vials. The ICP-AES was calibrated with NIST traceable standards and a laboratory reference, run at a rate of 1 per 15 samples. Internal laboratory references included one soil sample from City Park and one high metal sample from the junction of Elysian Fields and Highway 10. Duplicate extractions were included for every 15 samples. Further QA/QC was achieved by participation in the Analytical Products Group, Inc. laboratory proficiency environmental testing program. The same Spectro Analytical Instruments ICP-AES was used in both surveys to analyze for metals.The change in extraction occurred after practical experience that indicated that student skills with pipettes were sometimes creating uneven measurements and dilutions. The dilution step was so common with urban soils that we sought to reduce this source of error by changing the amount of soil being extracted. Changing to a 1 to 10 ratio served the purpose of simplifying the extraction by eliminating the dilution step and improving laboratory productivity. For comparison of the relationship between 0.4 and 4.0 g, a group of reference samples were used. The effect of sample preparation methods was investigated with milled and thoroughly mixed soil samples purchased as part of the Wageningen Evaluating Programmes for Analytical Laboratories International Soil-analytical Exchange (WEPAL ISE).A CIROS CCD ICP-AES (Spectro Analytical Instruments) was used for the metal analysis. The results yielded a linear relationship between the two sample extraction methods. Because of the linearity it was possible to adjust for differences for the method of Survey I to Survey II results. Table 1 shows the results of the WEPAL data that were used to adjust the results of Survey I to match the results of Survey II. The multiple metal accumulation (MMA) refers to the sum of the concentrations of metals in each sample.The following procedures were used to make MMA maps for each survey: First, the median soil MMA was assigned to the centroid of each census tract for both surveys. Second, the centroids with their corresponding MMA value were used to create a floating-point grid by kriging interpolation in Surfer 6.0. Third, the resulting Surfer grid was imported into ArcView GIS 3.3 using the Grid Machine extension. Finally, dividing the grid of Survey II MMA by grid of Survey I MMA created a ratio MMA grid reflecting the differences between the surveys [21, 22, 23].Table 2 shows the results of 8 metals measured and transformed for Survey I, metal results obtained directly from Survey II, and metal results measured in fresh parent alluvium of the Mississippi River. Comparing Pb in the two surveys, note that the smallest amount of Pb appears below the median in Survey II and that the largest amount of Pb appears in the samples above the median of Survey II. This means that the differences between the lowest and highest amounts of Pb are larger in Survey II. The results for Survey I appear to be attenuated for lower and upper quartiles compared to the results for Survey II. This trend shows up with most of metals in the MMA map presented in Figure 1. Note especially that, with the exception of Mn, fresh alluvium from the Bonnet Carré spillway contains only a fraction of the metals that are found in the city.The current soil metals found in Mississippi River Alluvium provides the basis for comparing the amount of metals in parent materials of the Mississippi River Delta and the degree of the metal contamination of the alluvial soils of New Orleans [15]. As shown in Table 2, the Mississippi River Alluvium compared with Survey I and II soils shows a difference of median MMA by a factor 2.6–2.9. Using the Bonnet Carré Spillway samples as a baseline, the New Orleans soils exhibited the largest amounts of contamination with factors of 26 and 20 for Surveys 1 and 2 for Pb respectively, 13 for Zn, a factor of about 5 for Cu, and a factor of < 5 for Cd, Cr, and Ni. An interesting exception to the pattern is Mn which shows about the same abundance in the fresh alluvium as found in the urban alluvial soils of New Orleans. Reduced Mn is water soluble and flows with ground water to the river where it discharges and oxidizes into an insoluble compound. A biogeochemical mechanism results in the accumulation of Mn in alluvium along the banks of the river [17].The preliminary results of the two surveys suggest that overall, during the past decade there has been a modest decrease of metals in the soils of New Orleans. The decrease is not uniformly distributed. The major decrease of metals appears in the outlying area of the city while the inner city exhibited a prominent increase of metals. The most substantial increases in the inner city are found in soil Pb and Zn.Possible reasons for the significant increases in Pb and Zn include: New Orleans is a historic city with old wood homes dating to the 1880’s through the 1950’s when use of Pb-based paint was prevalent [24]. Zn is also a major constituent of paint [24]. During the 1990’s, New Orleans underwent an economic revival accompanied by housing renovation where power sanding was common [24]. Automobile engine wear and tire wear also releases Zn and other metals. The annual average daily traffic (AADT) increased by 4% on U.S. Interstate 10 into New Orleans from 126,500 in 1992 to 131,123 in 2001. Thus, in 2001 each km of the highway results in over 130,000 km vehicle travel each day.This will result in traffic flow of 47,450,000 km (29,466,450 miles) of vehicle travel per km (mile) per year. The material consequences of this quantity of traffic would be very large in terms of the metal released by engine and tire wear per distance travelled. The suburban decreases are not easily understood. There are several possibilities: The suburbs have relatively little traffic flow. The suburbs are relatively new, many of the homes are brick, and they do not have significant surface areas with lead-based paint compared with the old wood buildings of the inner city. Perhaps clean dust from the outlying areas is being deposited into the suburbs whereas contaminated dust originating from the suburbs is being pulled and deposited by the urban heat island into the inner city. This hypothesis is supported by the observation that in three cities, Indianapolis, Syracuse, and New Orleans, children’s blood lead is inversely related to seasonal changes of soil moisture indicating that the soil Pb is both a sink and a source of exposure and continuously undergoing deflation and deposition, depending on weather conditions (25).Perhaps the comparison is biased by the method of analysis. Further study is being conducted to approach the problem with alternative methods for comparing the surveys. First, instead WEPAL ISE samples, analysis of 4.0g and 0.4g extracts is being done on archived soil samples from ten Census Tracts. Increasing the number of samples used to establish the relationship between the two extraction methods may improve the results [22]. In particular, WEPAL ISE soil samples fail to cover the upper range of contamination that is such a prominent characteristic of the soil metal results observed in New Orleans. Second, the analysis is being conducted using the Spectro FME ICP-AES instead of the Spectro CIROS ICP with axial orientation. The older Spectro FME ICP with radial orientation may have different analytical characteristics whereby it attenuates both lower and higher measurements. This characteristic was noted when we switched analysis from AAS to ICP-AES. Improved methods for statistical analysis may also be available to refine the evaluation between the two surveys [26]. Further study using alternative approaches is being undertaken to evaluate the differences between the soil metal surveys of New Orleans.In perspective, the issue of metal accumulation in cities is a major concern because exposure to metals may pose an enormous set of chronic problems, especially to children, which ultimately affect the wellbeing of the larger society [18, 27].Comparing the fresh alluvial materials of the Mississippi River Delta with the current urban soils shows that the city communities of New Orleans have undergone a massive accumulation of soil metals since the city was first settled. In the 1990’s, preliminary results indicate that New Orleans soil Pb and Zn increased significantly in the inner city communities and decreased significantly in the outlying communities. These results are reflected by the MMA map. The preliminary conclusions are under review and other techniques are being used for evaluating the changes in extraction techniques between the surveys, changes in analytical instrumentation in comparing method, and the choice of GIS techniques.Preliminary MMA maps of New Orleans. Eight metals were measured in soils collected from all residential communities of New Orleans. The sum of the metals in soils is referred to as Multiple Metal Accumulation (MMA). The figure has three MMA maps. The MMA map for Survey I (1992) is shown on the upper left, the MMA map for Survey II (2000) is displayed on the upper right, and the MMA map of the ratio of Survey II/Survey I is shown below. These preliminary maps indicate that there has been an increase in the sum of metals by census tract in the inner city and a decrease in the sum of metals by census tract in the outlying areas of the city. See the text for further discussion.The following are the WEPAL derived estimated values of parameters A (intercept) and B (slope) in Y = A + BX where Y is predicted value of Survey II sample preparation data and X is observed value of Survey I data for each metal:Results of transformed metals of Survey I, soil metal data for Survey II, and Alluvial Parent Material of the Mississippi River Delta. All metal results are given in mg/kg.Support for Survey I was from a combination of Louisiana Education Quality and Science Fund (LEQSF) funding and ATSDR/MHPF for multi-metal analysis of the soil samples. Survey II was conducted through funding by the ATSDR/MHPF Cooperative Agreement to Xavier University of LA.