(8.3)where DT,R is the absorbed dose averaged over tissue or organ T, due to radiation R, wR is the radiation weighting factor and wT is the tissue weighting factor for tissue or organ T (Table 8.18).Table 8.18
Tissue weighting factor (wT) values (reproduced with permission from [34])

Tissue

wT

ΣwT

Bone-marrow (red), colon, lung, stomach, breast, remainder tissuesa (nominal wT applied to the average dose to 14 tissues)

0.12

0.72

Gonads

0.08

0.08

Bladder, esophagus, liver, thyroid

0.04

0.16

Bone surface, brain, salivary glands, skin

0.01

0.04

a Remainder tissues (14 in total): adrenals, extrathoracic (ET) region, gall bladder, heart, kidneys, lymphatic nodes, muscle, oral mucosa, pancreas, prostate, small intestine, spleen, thymus, uterus/cervix


The effects of ionizing radiation, mostly as a health hazard, have been studied for several decades. In this regard, the term nominal cancer risk coefficients has been introduced. These coefficients define the incidence probability of stochastic effects per radiation dose. The nominal risk coefficients depend on age, sex, averaged lifetime risk, among other radiobiological factors. In the twentieth century the nominal risk coefficient for cancer risk after exposure to ionizing radiation was estimated in 5.5% and 4.1% per Sievert (Sv) for the general population and for adult workers, respectively. The nominal risk for heredity damage was estimated in 0.2% and 0.1% per Sievert (Sv) for the general population and for adult workers, respectively (Table 8.19).Table 8.19
Nominal risk coefficients for cancer and hereditary effects (10−2 Sv−1) (reproduced with permission from [34])

Exposed population

Cancer

Hereditary effects

Total

General
Workers

5.5
4.1

0.2
0.1

5.7
4.2


As previously mentioned, people are exposed to ionizing radiation from natural and artificial sources throughout their life. Which is why dose limits were implemented, seeking to prevent deterministic effects or to reduce the risk of stochastic effects. These dose limits are applicable only for situations of planned exposure and doses above the normal natural background radiation. However, these dose limits are not applied in the medical field so as to not hamper the effectiveness of diagnosis or treatment. Dose limits are applied into two main groups of exposed individuals: (1) occupationally exposed and (2) public (Table 8.20) [85]. Dose limits are strongly regulated to ensure that no one is exposed to an excessive amount of radiation in either normal or planned situations.Table 8.20
Recommended dose limits in planned exposure situations (reproduced with permission from [34])

Type of limit

Occupational

Public

Effective dose

20 mSv/yeara,b

1 mSv in a yearc

Annual equivalent dose in

Lens of the eye

20 mSv/yeara,d

15 mSv

Skin (averaged over 1 cm2 of skin)

500 mSv

50 mSv

Hands and feet

500 mSv

–

a 100 mSv in 5 years (averaged over 5 consecutive years), with no single year exceeding 50 mSv
b After a worker declares a pregnancy, the dose to the embryo/fetus should not exceed about 1 mSv during the remainder of the pregnancy
c Value applies to the average value over a period of 5 years
d Revised value by ICRP Statement on tissue reactions in ICRP publication 118 [32]


Q1.
When radon is inhaled the largest dose is found at the level of the(a)
Mouth
 (b)
Bronchial wall
 (c)
Bifurcation of the trachea
 (d)
Alveoli

 Q2.
When is the initial triage used?
 Q3.
How many categories of priority are used in initial triage?
 Q4.
What methods of treatment are used for internal contamination with radionuclides?
 Q5.
In hematopoietic syndrome the death is mainly due to(a)
Anemia
 (b)
Dyspnea
 (c)
Hemorrhage and infection
 (d)
Thrombopenia

 Q6.
In gastrointestinal syndrome the death is mainly due to(a)
Anemia
 (b)
Fever
 (c)
Vomiting
 (d)
Circulatory collapse

 Q7.
The LD50 for humans is about without medial support measures is(a)
1 Gy
 (b)
4 Gy
 (c)
10 Gy
 (d)
50 Gy

 Q8.
Discuss the most appropriate biodosimetric assay(s) for use in a suspected cases of radiation exposure which occurred approximately?
 (a)
12 h ago
 (b)
1 month ago
 (c)
1 year ago


When radon is inhaled the largest dose is found at the level of the(a)
Mouth
 (b)
Bronchial wall
 (c)
Bifurcation of the trachea
 (d)
Alveoli


What methods of treatment are used for internal contamination with radionuclides?


In hematopoietic syndrome the death is mainly due to(a)
Anemia
 (b)
Dyspnea
 (c)
Hemorrhage and infection
 (d)
Thrombopenia


In gastrointestinal syndrome the death is mainly due to(a)
Anemia
 (b)
Fever
 (c)
Vomiting
 (d)
Circulatory collapse


The LD50 for humans is about without medial support measures is(a)
1 Gy
 (b)
4 Gy
 (c)
10 Gy
 (d)
50 Gy


Discuss the most appropriate biodosimetric assay(s) for use in a suspected cases of radiation exposure which occurred approximately?


SQ1.
c
 SQ2.
Initial triage is used to screen the patients with severe injuries after a mass biological, chemical, radiological or nuclear event.
 SQ3.
Four categories of priority.
 SQ4.
Isotope blocking, dilution, or displacement, and the use of ion exchange resins, and ion mobilization or chelation.
 SQ5.
c
 SQ6.
d
 SQ7.
b
 SQ8.
The data in Sect. 8.6 should be referred to in order to formulate a full answer, based on the scenario of exposure. However, the short answers are: (a) gamma-H2AX and dicentric assays; (b) dicentric or CBMN assays; (c) FISH translocation assay.


Initial triage is used to screen the patients with severe injuries after a mass biological, chemical, radiological or nuclear event.


Isotope blocking, dilution, or displacement, and the use of ion exchange resins, and ion mobilization or chelation.


The data in Sect. 8.6 should be referred to in order to formulate a full answer, based on the scenario of exposure. However, the short answers are: (a) gamma-H2AX and dicentric assays; (b) dicentric or CBMN assays; (c) FISH translocation assay.


Environmental radiobiology refers to the study of the effects of radiation on ecosystems and species that are part of various habitats, collectively known as “the environment.” The discipline is part of Radioecology which is a broad area of research, covering the transfer, uptake and effects of radionuclides in the environment. Radioecology includes, for example, the speciation of radionuclides in environmental media, the transfer of radionuclides through the different environmental compartments and exposure of wildlife to ionizing radiation and its consequences. While this chapter focuses predominantly on the biological and ecological impacts of radiation on non-human species—since transfer is a key aspect of wildlife dosimetry—the environmental behavior of key radionuclides is briefly covered in Sect. 9.2.


It is important to understand that the basic mechanisms that lead to effects in humans, discussed in earlier chapters, also occur in non-human biota, but the effects of concern lie at higher levels of organization, such as the population or ecosystem. For example, a harmful mutations induced by radiation exposure may lead to cancer on humans, but in the environment, where the sustainability of the population is a critical endpoint, low levels of carcinogenic mutations are unlikely to impact the overall population. This means that the tools and techniques needed to document and evaluate radiobiological effects in natural populations, and ultimately in ecosystems, are much more complex to those used in human radiobiology.


A key issue is the importance and the difficulty of conducting good experiments in field situations, particularly at environmentally relevant concentrations and with proper controls. Single species studies in the laboratory have an important role in determining high and low dose effects, understanding mechanisms and testing resistance. But results can be misleading if they are extrapolated to environmental conditions, with lower doses, chronic exposures, and a variety of confounding factors such as genetics, age, life stage, predation, availability of resources, as well as the interaction with other stressors and difficulties to make a proper dosimetry [1].


Another important issue is how to measure impacts on ecosystems. Several robust biomarkers are available to determine impacts at the level of the gene, cell, tissue, organ, and organism. These are discussed in Sects. 9.3 and 9.4 of this chapter. Population level markers are also available including population numbers, mortality and morbidity, fecundity and population growth rate, but at the level of the ecosystem, the complexity makes it very difficult to assess ecosystem health following radiation exposure, including effects on functions and services. The importance of legacy sites is discussed in Sect. 9.4, as natural labs like, for example, “Radioecological observatories” (https://radioecology-exchange.org/content/radioecological-observatories) where all the mechanisms of effect from populations to ecosystems can be deeply studied. Other approaches include measurements of biodiversity index and the use of drone technologies to monitor ecosystem change at the gross level, for example, forest cover and diversity, lake eutrophication, or extreme habitat change.


Transfer of anthropogenic radionuclides through food chains has been studied since the time of atmospheric weapons testing and has been supported by data from nuclear power generation and accidents, as well as studies of the behavior of naturally occurring radionuclides (NORs). While there is a wealth of data on the transfer of radionuclides through human food chains, there has been less focus on wildlife and especially organisms that are not common sources of food for humans such as insects and invertebrates. While much of the focus in studying the environmental impacts of radiation has been on the uncertainties in effects measurement, it is important to stress that there are also uncertainties in dosimetry, and especially from internal radionuclides. Hence, knowledge of the factors influencing the behavior of radionuclides in the environment will be fundamental to support dosimetry and exposure assessments. This includes information on the behavior of naturally occurring radionuclides, which is needed both to calculate background doses to organisms, and thus put anthropocentric exposures into perspective, as well as to assess doses in areas with enhanced levels of natural radioactivity.


Naturally occurring radionuclides (NORs) include the radionuclides 14C, 3H, and 40K and also radionuclides that arise from three decay chains: the uranium (238U), the thorium (232Th), and the actinium (235U) decay chains [2] (Figs. 9.1 and 9.2). When they are contained in or released from processing materials they are defined as NORM [3]. Uranium and thorium are both metals belonging to the heavy actinide series, giving rise to long and complex decay chains that contain important radionuclides in the context of environmental radiation exposure (Fig. 9.1). Key radionuclides include isotopes of radon (222Rn with a half-life of 3.8 days; 220Rn with a half-life of 55 s), radium (226Ra half-life of 1602 years, 223Ra half-live of 11.43 days; 228Ra with a half-life of 5.7 days), and polonium (210Po with a half-live of 138 days, 216Po with a half-life of 0.145 s, and 212Po with a half-life of 299 ns). Compared to typical exposures from accidents such as Chernobyl and Fukushima, which are predominantly beta and gamma-emitting radionuclides, NORM exposures are often characterized by high levels of alpha emitters.

3 decay chains with alpha and beta decays of Uranium, U 238, Actinium, U 235, and Thorium, T h 232.

Fig. 9.1
Uranium (including uranium 238U and actinium 235U) and thorium decay chains

An illustration of the natural radionuclides on building materials, groundwater and surface water, soil, and rock.

Fig. 9.2
Natural radionuclides distribution in different environmental compartments


3 decay chains with alpha and beta decays of Uranium, U 238, Actinium, U 235, and Thorium, T h 232.


An illustration of the natural radionuclides on building materials, groundwater and surface water, soil, and rock.


Radionuclides in the environment can be distributed through the Earth’s atmosphere, hydrosphere, and lithosphere (Fig. 9.2). The behavior and fate of radionuclides in the environment depend on physical and chemical properties of radionuclides, the location and the type of emission source, and the environmental conditions [4]. Radionuclides undergo chemical reactions that affect their distribution and retention time. Organisms interact with the nonliving environment and can be exposed to the radionuclides. In order to estimate the doses received by an organism, the activity concentration of radionuclides in the organism’s habitat is calculated.


The natural environment is a highly complex system in which elements flow and circulate through the spheres of the Earth. To simplify the study of radionuclides, the environment can be divided compartments such as air, surface and groundwater, sediment, soil, and biota. Compartments are usually chosen so that they are distinguishable by spatial boundaries [5]. In each compartment, there are certain processes that have the greatest influence on behavior, so simplifications are made by only taking into account the key interactions that are important to consider for the radionuclide in question. As such, an environmental compartment can be chosen so that it is a volume of medium within which it is assumed that system parameters are constant and chemical concentrations do not vary spatially [6]. For example, in the air compartment, the aerosol formation and particle deposition process of emitted radioactive iodine (e.g., 131I) are key processes to consider, while in the soil compartment, the association with organic matter has been considered the process that determines the largest share of the fate of iodine. Assumptions can be made so that only the key reactions and dynamics are taken into account.


In general, the first step in studying the behavior of the radionuclide in the environment is to obtain knowledge of the location and properties of the emission source. Knowing where the radionuclides come from and in what form they occur can already reveal much information about where the radionuclides will be transported to. For example, the radioactive uranium released from nuclear explosions may end up in very different locations than uranium in nuclear waste dumped into the sea or uranium brought to the surface during the mining of uranium-bearing ores [7]. In addition to the location, the type of emission source should be considered. Anthropogenic emissions of radionuclides result from human activities. These radionuclides are released into the environment at a certain point in time. Unlike anthropogenic emission sources, natural emission sources from the subsurface have been present since the creation of the Earth. Uranium and thorium ores, for example, can be considered as diffuse sources of radionuclides in the Earth’s crust. If groundwater near a uranium deposit flows in a particular direction toward areas where drinking water is extracted, it may behave as a point source. Anthropogenic radionuclide sources, such as nuclear weapon tests and nuclear power plant accidents, release radionuclides at high temperatures and pressures in a certain area over a relatively short period of time and can therefore, be considered a point source. Depending on the weather conditions, the radionuclides can be further dispersed as clouds, with the emission spreading diffusely rather than being a point source. Other point sources, such as the emission of nuclear waste dumped in the ocean, release radionuclides diffusely over a large waterbody. Radionuclides that are dispersed without a specific point of discharge and over a long period of time may be considered as a diffuse source. Agricultural practices, for example, often require high levels of fertilizers, which end up in water bodies through various diffuse processes. Phosphate rock in fertilizers can contain small amounts of naturally occurring radionuclides such as uranium, thorium, and radium. Human activities can enhance the release of radionuclides.


The study of the fate of radionuclides is complicated by the property of radioactive decay. Radioactive decay changes the type of radionuclide, thereby altering its physicochemical properties and potentially altering the fate of the entity. That is, when a radionuclide decays, the daughter element often has very different chemical properties than the parent element [8]. If the parent element is a solid and its daughter is a gas, the parent may partition into other compartments, such as air or water. For example, in the natural uranium (238U) decay series, radon (222Rn) is formed after the decay of radium (226Ra). Radium is an alkaline metal that can be present in a mineral structure within the parent rock or in the pore water as an ionic salt, while radon is an inert gas. If the released radon is captured in a closed space such as the basement of a building or a cave, it can be inhaled by an organism. The gaseous 222Rn decays further releasing alpha and beta particles and eventually decays into stable solid 206Pb. The latter is a metal chemically toxic for organisms. When radionuclides are the stressors of concern, both chemical- and radiation-induced effects on organisms are expected.


Once the radionuclide is emitted, its chemical speciation determines how the radionuclide reacts with components in the environment. It is important to keep in mind that radionuclides are not only physical entities, but also have chemical characteristics [9]. For a more detailed discussion of the importance of the chemical characteristics of radionuclides, the reader is referred to the text by Whicker and Schultz [10]. Radionuclides can occur in various chemical forms or species that have different mobility. The following examples of species are for thorium (Th). Radionuclides such as Th can occur in elemental form (e.g., Th0), but these are very rare in the environment. They can be present as free ions in water (e.g., Th4+). However, dissolved Th is almost always complexed in natural water [11]. Free ions can be bound to inorganic or organic molecules in either the solid or dissolved phases, such as thorium hydroxyl complexes Th(OH)40, Th(OH)3+, Th(OH)22+, ThOH3+, Th(SO4)2+, Th(HPO4)32−, Th-oxalate and Th-EDTA complexes. Radionuclides can also be components of a mineral, such as thorianite (ThO2), and thorite (ThSiO4). The thermodynamic properties of various species can be used to compute liquid-solid equilibria relations. These theoretical calculations reveal much about the possible conditions for and the extent of mobility of radionuclides [11]. The thermochemical data and adsorption results from laboratory experiments help to explain the behavior of radionuclides, such as Th in natural waters, sediments, and wastes.


In general, the total sum of chemical species can be expressed as [9]:

 (9.1)where (MS) is the total sum of species present; (M)n+/− the element present as positively or negatively charged free ion (n+/−); (MmLm)n+/− an element complexed by any kind of ligand, L, such as an oxide, organic, or any other form, negatively or positively charged; (MmA) an element adsorbed onto a surface or trapped in a crystal lattice, or in an amorphous structure, A; m is the number of M or L molecules in the complex; and n+/− is the number of charges.


The fraction of the different chemical species in this formula, that are present in the environment, will depend on the source of the radionuclide and the physicochemical conditions of its surroundings. Parameters such as pH, redox state, ionic strength and the presence of complexing ligands will influence the proportions of each chemical species present.


Some chemical species of radionuclides undergo chemical reactions that influence their mobility or retention. The main chemical reactions determining speciation are adsorption and desorption processes, ion exchange and dissolution reactions, precipitation and co-precipitation, complexation to inorganic and organic ligands [12] and redox reactions. For a detailed explanation of the mechanisms of these reactions, please refer to a course on aquatic chemistry such as Langmuir [8] or Sparks [13].


Of particular interest when studying the behavior of radionuclides are the chemical reactions at the solid–water interface, such as complexation with ligands and adsorption to mineral surfaces. These reactions will largely determine whether the radionuclide is mobile and potentially available for the biota to take up. A dissolved species can associate with an ion or molecule ligand and form a complex [8]. For example, Th is a complex-forming actinide metal for which the chemical speciation of the cation changes with the pH. The multivalent Th cations tend to form strong hydroxyl (OH) complexes. Only in acid waters, the OH concentration is low enough so that competition with ligands is minimal. In these conditions, it is easier for ligands to displace OH and complex it. Complexation with carbonates, humic materials, or other ligands increases the solubility of the Th species and thus the mobility in the environment. An adsorbed species can associate with charged surfaces or broken bonds of minerals. For example, Th adsorbs onto clays, oxides and organic matter in soils and sediments. The adsorption of Th increases if the pH increases from acid to neutral conditions [11]. Sorption processes increase the retardation of Th and thus decrease its mobility in the environment. In general, Th in the soil compartment will remain strongly adsorbed onto soil constituents so that contamination of groundwater through the transport of Th from soil to groundwater will not occur in most soils [14]. Certain microorganisms (Pseudomonas aeruginosa) present in soils may enhance the dissolution of Th by producing chelating agents that can form soluble complexes with this radionuclide [15]. This is not the only way for microorganisms to influence the speciation and mobility of radionuclides. They can also, for example, change their redox state, immobilize them by processes such as biosorption, biomineralization, and precipitation [16]. In the water compartment, soluble Th ions will hydrolyze at a neutral pH forming complexes with OH. The Th-hydroxyl complexes can in turn be absorbed on suspended particles in the water. Although dissolved Th tends to form strong complexes, facilitating its transport, Th concentrations in natural waters—with pH between 5 and 9—remain limited by the scarcity of the element, small solution rates and insolubility of Th-bearing minerals [11]. In groundwaters at mining facilities, Th concentrations may be higher due to the more acidic conditions which cause the leaching of Th.


A common approach to quantify the mobility and availability of radionuclides in the environment is to estimate the ratio between the activity concentrations of the radionuclide in two chosen compartments or trophic levels [9, 17]. The radionuclide retention on the solid phase is estimated by determining a partitioning coefficient. The coefficient describes the partitioning of a radionuclide between the solid and aqueous phases and takes no explicit account of sorption mechanisms [18]. It is assumed that an equilibrium exists between the dissolved and sorbed amount of radionuclides and that exchange is reversible [19]. This simplification relates the concentration of a radionuclide in water to the amount of radionuclide adsorbed:

where Maq and Mads are the aqueous and adsorbed species, respectively.


A solid-liquid distribution coefficient (Kd) is derived from the ratio of radionuclide concentrations in the solid phase to that in solution and is calculated as:

 (9.2)where Aint is the initial radionuclide activity (Bq), Aeq is the equilibrated radionuclide activity (Bq) in the aqueous phase, V is the volume of the liquid phase (L), and m is the mass of solid phase (kg).


The adsorption of radionuclides onto soil particles is often expressed as a Kd value. The Kd is determined by adding a known amount of sorbent (i.e., clay, oxide, soil) to a solution with an initial radionuclide concentration, and after equilibration and phase separation (e.g., by ultracentrifugation or a dialysis membrane), radionuclide concentration in the aqueous phase at equilibrium is measured.


In case of radiocesium (e.g., 134Cs and 137Cs), for example, the CsKd value is obtained by the ratio of the total radiocesium activity concentration in the solid phase and in liquid phase after a chosen time of contact between the two phases. The experimental design must be carefully thought out, as parameters such as contact time, radionuclide concentration, solid to liquid volume, and the ion composition of the aquatic phase affect the Kd value. Radiocesium dissolves well in water, so that radiocesium exists in the aqueous phase only as a free ionic species. Only one metal species of Cs should be considered, which simplifies the study of adsorption equilibria. Moreover, radiocesium cations can be directly adsorbed from solution by an organism, because the cations have no tendency to form soluble complexed species [20]. Thus, the CsKd value can be determined in a relatively simple manner and it can provide useful information about the radiocesium accessible to the organism for uptake [18].


However, caution must be taken in interpreting a Kd value, as it may change over time [18]. On the one hand, the Kd changes in a short term, because an equilibrium is not always reached instantaneously, as for example for radioactive isotopes of iron [21]. On the other hand, the Kd changes in long term, because adsorbed radionuclides, such as 137Cs, can migrate deeper into structures of minerals so that it is no longer available and becomes fixed. Kd values are often determined by short-term laboratory experiments lasting several hours or days. However, Kd values can also be determined in the field, where the results depend on the time elapsed since the contamination occurred and this gives a more reliable picture of the long-term fate of the radionuclides. The time effect was studied in a laboratory study [22] with soils showing that CsKd values of mineral soils with 5% clay minerals can increase from 30 to 1000 L/kg in 40 days and 200 to 5000 L/kg in 415 days for peaty soils with 10% clay minerals. In this example, the CsKd of the mineral soil increases by a factor of 30 over a relatively short period of time, and the CsKd of an organic soil increases accordingly but over a much longer period of time. Laboratory results of CsKd values can only partly explain the reduction in Cs soil-to-plant transfer in the field. A study after the Chernobyl accident [23] shows that 137Cs soil-to-plant concentration ratios, that were initially elevated, were reduced by more than 50 times in the following years. This trend was explained by an initial step of radionuclide release from fuel particles into soil aqueous phase, followed by a reduced transfer attributed to the progressive fixation of 137Cs by soil minerals, referred to as “aging effect” that makes 137Cs gradually less available for uptake by the plant.


In many cases, the factors that influence the transfer of radionuclides to biota are similar for humans and include soil and water chemistry, speciation of radionuclides, as well as biokinetics (biological and ecological half-lives) and interactions between radionuclides and stable elements. For example, the soil-to-plant transfer of 137Cs, is influenced by clay content and K levels in the soil, and radiostrontium (90Sr) by Ca levels. Another example, is the uptake of U to fish and other aquatic organisms, that is are dependent on pH and carbonate concentrations, which change the availability and complexation of this element [24]. In contrast to Cs, radionuclides such as U exist as several species in the environment. The bioavailability of different U species in soil to ryegrass was studied in a laboratory pot experiment [25], which showed that speciation has an important influence on the uptake of U by grass. From the results, it was concluded that the uranyl-cation (UO22+) and uranyl-carbonate complexes (e.g., UO2CO3(aq), UO2(CO3)34– and (UO2)2CO3(OH)3–) together with uranyl-phosphate (UO2PO4–) are the forms that are most readily taken up by ryegrass and thus are more bioavailable compared to other uranyl-phosphate complexes (e.g., UO2HPO4) and the hydroxy- (e.g., UO2(OH)2(aq) and UO2OH+) and sulfate-complexes (e.g., UO2SO4(aq) and UO2(SO4)2–). As demonstrated in the previous examples, some species are not available for uptake by biota. Hence, a value other than the total concentration in the compartment should be used to estimate the bioavailability of a given radionuclide and, the exposure of biota through ingestion of radionuclides should only be estimated from the activity concentrations of the bioavailable species [17].


Internal exposure and toxic effects of radionuclides require that an organism takes up the radionuclide, and for chemically available species to be taken up by biota, the radionuclide must be able to cross cell membranes [26]. To investigate whether this exposure will occur through ingestion, it is important to know whether this contaminant is a source for ingestion by biota. A radionuclide’s potential for biota uptake in soil and sediments is defined by its bioavailability or bioaccessibility. There is a slight difference between the bioavailability and bioaccessibility of pollutants in sediment and soil. This difference has implications for the design of experimental set-ups, but also for the interpretation of results. The bioaccessible fraction is the species in the environment, which are available to cross an organism’s membrane if the organism has access to the radionuclide in the longer term [26]. The bioavailable fraction is freely available to cross an organism’s membrane from the medium the organism inhabits at a given time. For example, technetium (Tc) may be highly mobile in aqueous solution at oxidation state +7 (i.e., Tc(VII)), but strongly absorbed and retarded in the subsurface at oxidation state +4 (i.e., Tc(IV)) [27]. Technetium is used in nuclear medicine for diagnosis and is emitted in the environment from the nuclear fuel cycle. Technetium exists primarily in two stable oxidation states as Tc(VII) or as Tc(IV), and the two species can have a different fate when released to the environment. While TcO4− in solution is bioavailable, TcO2·nH2O is expected to be adsorbed at low concentrations and precipitated at high concentrations. The species TcO2·nH2O can become available for uptake when oxidized by air and is thus bioaccessible.


Besides the speciation of radionuclides, the extent to which radionuclides can be transferred to different compartments is influenced by competition between ions. On the one hand, stable isotopes of the radionuclides may compete for adsorption to the solid phase or uptake by biota. For example, radionuclides such as 3H, 40K, 48Ca, 54Mn, 60Co, 65Zn, and 131I, are isotopes of essential biological nutrients [10]. Therefore, their uptake and retention characteristics are largely controlled by the flux of these essential nutrients through biological processes. On the other hand, elements that are chemically similar to the radionuclides may compete. Certain radionuclides behave in the environment in a similar way to essential elements for biota, due to their chemical properties. For example, 137Cs and 90Sr have similar chemical properties and follow the same transfer and cycling processes in the environment as the macronutrients potassium (K) and calcium (Ca), respectively. The tendency of these radionuclides to accumulate in the biota is reduced if there is an abundance of the analogous element in the environment [10]. Conversely, the accumulation of the radionuclide in the biota increases when there is a scarcity of the analogue element. For example, low concentrations of K and Ca in the soil can result in increased uptake of radionuclides by plants, as they find it more difficult to discriminate between nutrients and radionuclides under these stressful conditions [20]. As mentioned earlier, the long-term bioavailability of 137Cs and many other radionuclides depend heavily upon ecosystem characteristics, and in particular, soil properties [10]. Soils and sediments of high clay content can effectively immobilize 137Cs by chemical binding. In such systems, the soil acts like a sink for 137Cs and in time very little of the nuclide is available for biological incorporation. Other systems have sandy soils with a low cation exchange capacity, and larger quantities of 137Cs can be recycled through the biota of such systems for long periods of time [9].


In summary, depending on their speciation, radionuclides can be transferred in the biosphere from the emission source to different compartments until they reach an equilibrium or final sink, or they can be recycled within the environment.


Information on the uptake of radionuclides to biota is vital for calculating dose to the organisms, since both external and internal irradiation contributes to exposure. Soil and sediment dwelling organisms often have high external dose rates by virtue of their habitat, but also internal exposure from ingested radionuclides. Many field studies on radiation effects in wildlife are flawed due to underestimation of the internal dose, reporting only ambient air dose rates [28]. This is particularly important for 〈- (e.g., Ra) and ®-emitting (e.g., Sr) radionuclides, for which internal exposure is the greatest contributor to dose, but also internal contributions from radiocesium or radium, for example, can make a significant contribution to the overall dose.


There are a number of programs available for estimating the dose to biota. These are usually based on rather simplistic geometry and homogeneous internal distribution, but the basic principles are similar to those used for human dosimetry. They can also be adapted to give organ specific doses. For example, the ERICA Assessment Tool can calculate doses to a wide range of reference animals and plants, as well as user constructed organisms (see Box 9.1).


The ERICA Assessment Tool is a free to download, computer software system for assessing the risks of ionizing radiation to terrestrial, freshwater and marine biota (https://erica-tool.com/). The system is based on the three tier ERICA Integrated Approach that was originally developed as part of the ERICA EURATOM project [29] (see also https://wiki.ceh.ac.uk/display/rpemain/ERICA).


The ERICA Tool includes various components, all of which are linked to internationally recognized programs and databases. These include
Modelling transfer of radionuclides through the environment: links to IAEA Wildlife Transfer Database (WTD) and IAEA handbooks [30]; https://www.wildlifetransferdatabase.org/.

Methodology for estimating dose rates to biota from internal and external distributions of radionuclides: ICRP biota DC software version 1.5.1 for the calculation of dose conversion coefficients (DCC) [31].

Risk characterization in order to evaluate the significance of the dose rates received by organisms, including comparison with background radiation doses, screening values [32], Environmental Media Concentration Limits (EMCL) [33], derived consideration reference levels (DCRL) and biological effects (FREDERICA database, https://www.frederica-online.org/mainpage.asp).


Modelling transfer of radionuclides through the environment: links to IAEA Wildlife Transfer Database (WTD) and IAEA handbooks [30]; https://www.wildlifetransferdatabase.org/.


Methodology for estimating dose rates to biota from internal and external distributions of radionuclides: ICRP biota DC software version 1.5.1 for the calculation of dose conversion coefficients (DCC) [31].


Risk characterization in order to evaluate the significance of the dose rates received by organisms, including comparison with background radiation doses, screening values [32], Environmental Media Concentration Limits (EMCL) [33], derived consideration reference levels (DCRL) and biological effects (FREDERICA database, https://www.frederica-online.org/mainpage.asp).


The tool contains data on concentration ratios and DCC for all radionuclides in publication 107 [34], and in addition to a selection of pre-created reference organisms, allows users to create their own assessment organism.


The ERICA tool has been updated since its original release, and the current version, ERICA Tool 2.0 (beta version released in November 2021—https://erica-tool.com/the-erica-assessment-tool-has-been-updated-to-version-2-0/) includes updates on concentration ratios, as well as new approaches for calculation of dose contribution from short-lived progeny, noble gases radon and thoron [35–37].


Internal and external exposures are determined from specific dose conversion factors (DCC) combined with using field measurements of concentration activities or default concentration ratios (CR). The CR represents the activity concentration of radionuclides in biota (fresh and dry weight in animals and plants, respectively) and the activity concentration in soil (dry weight, upper 10 cm), water, or air for a given radionuclide [38]. The tool also allows the calculated exposures to be compared to background radiation or screening values.


The calculation of external dose rates takes account of the occupancy of the organism (i.e., percentage of time spent in, on, or above soil, sediment, or water) and is determined by:


DR—dose rate (Gy/unit of time)

DCC—dose conversion coefficient

Cmedia (Bq/kg or Bq/L)


Internal doses

DR—dose rate (Gy/unit of time)

DCC—dose conversion coefficient

Corganism (Bq/kg)


There are several other simplifications to the approach, including assumptions on habitat ranges and feeding habits of biota [38]. CR are lacking for many organisms and radionuclides; however, the tool provides default CR based on available data and assumptions (e.g., similar taxonomy or chemical behavior to other organisms or radionuclides).


Uncertainties in dose estimates can be reduced if field measurements are available, but determination of internal concentrations of radionuclides can also be challenging, as organisms may be too small for direct radiochemical analyses, or it can be difficult to distinguish between radionuclides internalized in animal tissues, from those adsorbed to the body segment or cuticle. Efforts have been made to compare ERICA default CRs with field measurements at Chernobyl, showing a relatively good agreement between the CR values calculated for many organisms [39]. However, it was concluded that such similarity may have resulted from the broad range of estimated CR values available [40].


In soil, Beaugelin-Seiller [41] concluded that DCC values are highly dependent on factors such as the porosity and soil water content, the body size of the organisms within other factors. For ®-emitters, the difference in DCC values recorded reached a factor of 3, between dry and saturated soil conditions. The calculation of doses in organisms under exposures to NORM is also highly dependent on assumptions of equilibrium that must be made for several radionuclides from the 238U decay series [42]. Usually a 100% equilibrium is assumed, although different equilibrium percentages are also accepted for radon, as it can escape to the atmosphere.


The positioning of organisms in the trophic chains and the composition of their diets may be determinant for the magnitude of exposures. In a coastal sand dune system, under a long-term contamination through atmospheric deposition and sea-to-land transfer of radionuclides at Sellafield nuclear reprocessing site (West Cumbria, England), Wood and collaborators [43] recorded high activity concentrations of 137Cs, 238Pu, 239+240Pu, and 241Am in soil detritivorous (e.g., Collembola and Isopoda) when compared with predators (e.g., Coleoptera larvae). Within the same trophic level, these authors also found significant differences in the whole-body activity concentrations of different invertebrate groups. Size also influences the internal doses to organisms. Dose calculations for two benthic invertebrates, the larval midge Chironomus tetans and the amphipod Hyalella azteca, based on estimations from NORM activity concentrations in sediments impacted by uranium mining demonstrated that the smaller amphipod, received a greater dose of alpha irradiation. This reflected the high content of ingested radionuclides within the gastrointestinal tract and that as diameter of the gastrointestinal tube decreases, the assessment factor (AF) for ingested alpha-emitters increases, as more alpha-particles are expected to reach the tissues of the organisms [42]. Therefore, it was suggested that the contribution of sediment within the gastrointestinal tract for the calculation of internal doses must be considered, and not only the activity concentrations of radionuclides recorded in external sediments.


In the case of accidents, there is also a need to account for historical dose and radionuclide decay, since observed effects may be a legacy of high levels of exposure after the accident. These high exposures can also be a source of confounding factors, since the initial damage may lead to indirect ecosystem changes (such as the replacement of pine trees by less sensitive species) [44]. While much of the focus in studying the environmental impacts of radiation has been on the uncertainties in effects measurement, it is important to stress that there are also uncertainties in dosimetry.


Following the discovery of X-rays by Wilhelm Roentgen in 1895 and of radioactivity by Henri Becquerel in 1896, studies on its effects started immediately. The detonation of the atomic bombs over Hiroshima and Nagasaki in 1945 raised the concern about the health impacts of radioactive contamination and the behavior of radionuclides in the environment [45]. Therefore, a great number of studies using a variety of plants and animals have been performed since then.


The first harmful effects caused by the exposure to ionizing radiation occur at the molecular and cellular level. If these effects are severe enough, they can impact tissues, organs, individual organisms, populations, and entire communities. However, even though an individual organism may suffer from severe damage at the molecular and cellular level, it does not necessarily mean that entire populations and communities will be affected [46]. It seems that individual organisms are able to sustain a certain level of effects before they are reflected at a population level [46]. However, when an effect is seen at the population level or at higher levels of organization (i.e., communities or ecosystems), it means that effects at individual organisms are expected to be occurring (Fig. 9.3) [45].

An infographic of environmental ionizing radiation exposure and effects. It includes the list of most frequent radiation types and exposure for gamma and particulate radiations of beta and alpha particles, and lethal, immune, molecular, reproduction, and genetic effects.

Fig. 9.3
Exposure and effects of different radiation types on organisms


An infographic of environmental ionizing radiation exposure and effects. It includes the list of most frequent radiation types and exposure for gamma and particulate radiations of beta and alpha particles, and lethal, immune, molecular, reproduction, and genetic effects.


There can be two types of effects caused by ionizing radiation. They can be stochastic or non-stochastic (deterministic). Stochastic effects are effects that occur by chance and the higher the dose the higher the probability of its occurrence. However, the severity of those effects is not dependent on radiation dose. The main stochastic effects related to ionizing radiation exposure are cancer and genetic damage/alterations (i.e., mutations) [47]. For non-human biota, stochastic effects that occur at germinal cells will be the ones that will have a higher impact, as they will have a higher probability of being inherited and, therefore, of affecting the next generations, impacting populations and communities [47]. Deterministic effects depend on time of exposure, doses and type of radiation. They are adverse tissue reactions that result from the damage or killing of many cells in an organ or tissue. The severity of these effects increases with dose when radiation levels reach a threshold, below which harmful effects to tissues/organs do not occur. The deterministic effects that are most important at a population level are mortality (which affects density, age distribution, and death rate), fertility (birth rate) and fecundity (which affects birth rate, age distribution, size of the population) [45] (Fig. 9.3). As for other stressors (i.e., chemicals), exposure to ionizing radiation can be acute or chronic. Acute exposures are short-term exposures to relatively high doses of radiation that usually last minutes or hours. Chronic exposures are long-term exposures or lifetime exposures to usually low doses of ionizing radiation. Doses in acute exposures are often reported as total absorbed doses, whereas for chronic exposures doses are often reported as dose rates (i.e., mGy/day, Gy/year, or mGy/h) [46, 48]. For a given dose of ionizing radiation, acute exposure induces higher injury than chronic exposure [46]. The higher the dose the lower the ability of cells to correctly and rapidly repair the damage and also the lower the ability of healthy cells to divide and regenerate the damaged tissue [46]. Depending on the dose received by cells or organisms, several types of effects can occur, namely genetic damage, DNA lesions that can induce teratogenic effects (malformations) on embryos when occurring in germinal cells (i.e., gametes), cell transformation in somatic cells and cell death (Fig. 9.3). In some cases, DNA damage can be so severe that it becomes incompatible with the survival of the cell or of the entire organism. Depending on the kind of cells that are affected (germ cell or somatic cells), there can be different consequences. Severe damage (i.e., DNA double strand breaks, gross mutation like duplications, deletions, translocations, and chromosome gain or loss) will cause cell death potentially leading to the death of the organism or, for example, to its sterility if it occurs in germ cells (Fig. 9.3). If the damage is not enough to cause cell death, it can cause cell transformation and cancer in somatic cells or it can affect the fitness of the organisms and entire populations if it affects germ cells. Mutations can cause a reduction in the production of viable embryos or viable gametes and also, they can be passed and accumulated throughout generations reducing the population’s fitness. Therefore, DNA alterations can have an important impact on fertility and fecundity and consequently in reproduction [46].


Also, there can be effects on the homeostasis of organisms (Fig. 9.3), namely depression of the immune system, alterations in normal metabolism, oxidative stress, and disturbances in the endocrine system [49]. The majority of the studies performed so far are focused on the determination of the acute effects of high doses of radiation, and only few studies are focused on chronic exposures to low doses of ionizing radiation.


The younger the organisms (namely fetuses and embryos) the more sensitive they are to the deleterious effects of radiation exposure. This is due to the higher sensitivity of cells that frequently undergo mitosis (which occurs frequently in young organisms for each tissue/organ as it is part of the growing process). Also, tissues/organs that have the ability to regenerate or that are constantly producing new cells like the hepatic tissue, the skin, the bone marrow, germinal cells, and gut lining are more sensitive to radiation (Fig. 9.3). The higher the cell division rate in an organism the more sensitive it will be to radiation’s harmful effects.


Regarding the sensitivity of parameters like mortality and reproduction, in general the reproductive capacity is a more sensitive parameter to the effects of radiation exposure both for terrestrial and aquatic invertebrates and vertebrates, than life expectancy (mortality) [45]. Negative effects on reproduction rate can occur at less than 10% of the radiation dose required to induce direct mortality in mammals [45].


All organisms evolved in the presence of radiation, being cosmic radiation or natural radiation emitted by NORs present in the earth crust [50]. The studies performed so far, on the effects of ionizing radiation, showed that there is a considerable variation in the response of organisms from the same or different species, due to intra- and interspecies variability in sensitivity. In general, it is widely accepted that mammals are the most sensitive organisms, followed by birds, fish, and reptiles and that invertebrates and other less complex organisms have the highest radiation resistance (Fig. 9.4) [46, 50]. However, it has to be noted that most of the knowledge gathered so far comes from laboratory exposures of specific strains of these organisms and that results may differ significantly from what happens to their wild counterparts.

A phylogeny chart of the approximate acute lethal dose ranges varying between 10 superscript 0 and 10 superscript 4 for the corresponding taxonomic groups of viruses, mollusks, protozoa, bacteria, moss, insects, crustaceans, reptiles, amphibians, fish, higher plants, birds, and mammals.

Fig. 9.4
Schematic representation of overall sensitivities of different taxa to acute gamma radiation exposure. (Reproduced with permission of UNSCEAR, adapted from UNSCEAR 2008 report, Annex E)


A phylogeny chart of the approximate acute lethal dose ranges varying between 10 superscript 0 and 10 superscript 4 for the corresponding taxonomic groups of viruses, mollusks, protozoa, bacteria, moss, insects, crustaceans, reptiles, amphibians, fish, higher plants, birds, and mammals.


The majority of the existing studies on the effects of ionizing radiation in cells are focused on DNA as the main target, making it clear that there is a cause-effect relationship between DNA damage with cytotoxicity and mutagenicity associated with ionizing radiation exposure. However, the cascade of molecular effects that lead to the induction of biological effects in exposed organisms is complex and involves, firstly, the interaction of radiation with water molecules and structural and functional biological molecules inside the cells. This interaction will induce the formation of ions, radical species, and excited molecules that will move from the site where they were formed to other cell compartments, causing damage to other biological molecules. This will trigger several signaling cascades, activating cell responses that will change the normal metabolic state of the cell, including changes in gene expression, enzyme recruitment and activities, DNA methylation patterns, and other stress-induced signaling events. When DNA is damaged, the cell cycle is interrupted allowing for DNA integrity check. DNA can be damaged directly through direct ionization or indirectly through the attack of free radicals that are formed when radiation interacts with water molecules of the cell [51]. Given the high content of water in cells, IR interacts with water in a process called radiolysis, generating free radicals as H∙ or OH∙, which trigger a cascade of events giving rise to other ROS as hydrogen peroxide and the superoxide anion [52] and references quoted. If not neutralized these products may diffuse within cells, as well as between cells, affecting other biomolecules such as DNA, proteins, and lipids, both in target and non-target cells (i.e., cells not directly irradiated) [53, 54]. Regarding DNA, ROS may oxidize bases or cause single and double strand breaks (SSB and DSB) [55]. Also, post-irradiation DNA lesions can be formed as a consequence of the attempt of the cell to repair sugar and base residues, which can be converted to SSBs (Single Strand Breaks) and DSBs (Double Strand Breaks) [51]. If DNA is correctly repaired, the cell will continue its cycle normally, if not, the cell can undergo transformation as mutations and chromosome aberrations may occur or if the damage is too severe, programmed cell death (apoptosis) will occur. The repairability of the damage and the repair accuracy will depend on damage severity and complexity. Low LET (beta particles, gamma and X-rays) and high LET (alpha particles and neutrons) radiation exposure can cause several types of DNA damage that are usually repairable, like SSBs, abasic and apurinic and apyrimidinic sites and DSBs (Fig. 9.5). However, the fraction of irreparable DNA damage depends strongly on LET. High and low LET radiation exposure can cause complex DNA damage, but this type of damage is more frequently associated with high LET radiation. Complex DNA damage is composed by closely spaced DNA lesions that form clusters [51]. Clusters contain two or more DNA lesions of the same or different origins, close to each other and on opposite strands (bistranded lesions). These lesions can be DSBs or non-DSBs oxidative clustered DNA lesions like SSBs, oxidized base lesions, and oxidized apurinic/apyrimidinic sites (AP sites) [51] (Fig. 9.5). These clustered lesions have a high mutagenic and carcinogenic potential since they are considered repair-resistant or even unrepairable due to the relative inefficiency of DNA repair systems to process such closely spaced and complex lesions. As there are several DNA repair systems in the cells and each of them is specialized in the processing of specific lesions, when several types of lesions are closely spaced in the DNA molecule, the different repair systems cannot act properly, retarding the repair and often generating other lesions. High LET radiation is mostly associated with the generation of DSB’s clustered DNA lesions and low LET radiation to non-DSB’s oxidative clustered DNA lesions [51], but this is not completely clear and needs further studies. High LET radiation is also associated with increased frequency of chromosome aberrations, and also to a high frequency of unrejoined DSBs and consequently with a higher cell killing efficiency, as unrejoined DSBs are a cause of cell death.

An illustration of the comparison of low and high L E T radiation D N A damage distributions where photons or beta particles are random in low radiation and alpha particles are concentrated in high radiation.

Fig. 9.5
High and low LET radiation DNA damage effects


An illustration of the comparison of low and high L E T radiation D N A damage distributions where photons or beta particles are random in low radiation and alpha particles are concentrated in high radiation.


Microorganisms, including fungi, can be seen as good indicators of the ecosystem’s “health.” They include ubiquitous and taxonomically diverse microorganisms that play important key roles on diverse ecosystems’ function. Specifically, with regard to radiation, microorganisms play a very important role in the health of these systems and in their cleaning and decontamination.


Microorganisms play a key role in the biogeochemical cycle of elements. In soils, they are important for organic matter turnover and maintenance of soil structure and fertility. As such, changes in the structure of microbial communities, by either metals or radionuclides, can have indirect effects on the above processes. Prokaryotes (bacteria and Archaea) have dominated a large part of the history of our planet, occupying virtually every “inhabitable” niche on earth. To be able to do that they have adapted to withstand large ranges in: (1) temperature, e.g., the hot temperatures found in hot springs and fumaroles, and the contrasting cold temperatures found on sea ice and polar regions, (2) pressure, e.g., deep sea, (3) salinity, e.g., hypersaline lakes, (4) pH, e.g., acid mine drainage sites, and (5) radiation, e.g., naturally occurring (deserts and high mountains, mining sites) and from nuclear contaminated sites [56]. Microorganisms that have adapted to such environments are referred to as extremophiles or polyextremophile (the latter being capable of withstanding different extreme conditions simultaneously), and these conditions are a requirement for their normal metabolic and biochemical operation. Most of these microorganisms belong to the domains Bacteria and Archaea although some fungal species have also been described. To survive these harsh conditions, extremophiles produce various primary and secondary metabolites, such as extremolytes, enzymes, and pigments [57]. Extremolytes, for example, are known to protect extremophiles cell structures and macromolecules from their harsh environments by forming protective water layers (e.g., ectoine), which is a co-solvent that shields proteins and cell membranes from UV light, heat, and dryness [58] around them or acting as chemical scavengers (e.g., carotenoids), protecting cells and their structures from UV radiation and oxidative stress [58]. Ultimately, the exceptional properties of these biomolecules find possible applications in various industrial sectors, in human healthcare, and well-being [59].


With regard to radioactively contaminated sites, microorganisms play an essential role on the mobility, toxicity, and distribution of radionuclides, through processes that include reduction, uptake, and accumulation by the cells, biosorption, and biomineralization with phosphates and carbonates [16].


Culture dependent and culture-independent approaches have shown the effects of long-term exposure to metals or radionuclides on individual species and on microbial communities. In addition, they have allowed those specific genes and cell functions mostly affected by radiation and metals to be identified, thus contributing to a better understanding of the molecular mechanisms behind microbial metal/radioresistance. Furthermore, the acquisition of genetic determinants by horizontal gene transfer contributes to shape microorganisms and microbial communities occupying these sites. More recently, refined metagenomic approaches focusing on prokaryotic communities have been employed and are expected to shed more light on the cells’ strategies to overcome radiation stress to remain operational.


The following section addresses in more detail some of the mechanisms that contribute to the survival and maintenance of microorganisms in these environments. We will end by referring to the impact of more recent methodologies, such as metagenomics and other omics technologies, and their contribution to clarify aspects such as the impact of these contaminants on the microorganisms and communities that exist in these sites.


It has been reported that when radiosensitive microorganisms are subjected to multiple high IR exposures, their resistance increases [60]. This was recently demonstrated by experimental evolution, where populations of Escherichia coli very resistant to IR were generated in the laboratory, after 100 selection cycles, and to which the dose needed to kill 99% of the population increased from 750 Gy to about 3000 Gy [61]. Likewise, radioresistant species can become even more resistant with repeated exposure [62]. This “memory” adaptation is associated with smooth genetic alterations that affect DNA repair and metabolic functions. During this process of adaptation, other physiological characteristics of the microorganisms are profoundly affected as, for example, growth which is slowed down, because the microorganism must direct its energies to other processes, such as effectively repairing the damaged DNA.


The association between genome size and radiosensitivity between taxa has long been suggested. For instance, for the same chronic exposure to IR, fungi, for which genome sizes range between 12 and 20 Mbp, suffer more DSBs per unit time than bacteria with their smaller genomes (3–6 Mbp). However, this is not true for Shewanella oneidensis and Deinococcus radiodurans whose genomes are practically the same size, but while the former is killed after exposure to a radiation dose causing one DSB, the latter manages to recover from hundreds of DSBs. This is probably due to the fact that D. radiodurans has up to ten identical copies of its genome per cell and uses this genetic information to repair its DNA. In addition, there is also evidence for the interference of non-enzymatic antioxidants such as manganese complexes, which protect proteins from IR-induced oxidation, facilitating the maintenance of cell homeostasis and DNA repair. Although in many radioresistant bacteria and yeasts, the most common DNA DSB repair pathway is similar to homologous recombination (HR),1 in fungi, non-homologous end joining (NHEJ)2 is the preferred, as in other eukaryotes, despite being error-inducing. Melanin pigments also seem to be involved in protection against multiple stressors, including IR as it can act as an oxygen radical scavenger [62].


Radioresistant microbial extremophiles have developed strategies to survive and withstand dose rates that to the majority of organisms, including humans, would result in acute health effects [63]. It is believed that radioresistant microorganisms possess highly efficient processes to repair DNA damage. However, it has recently been demonstrated that the repair mechanisms and the proteins involved are common to those found in radiation sensitive microorganisms [64].


The genus Deinococcus is probably the most well studied and characterized and there is a great deal of information, to what its radioresistance is concerned. Metabolically active Deinococci vegetative cells can tolerate chronic radiation levels of more than 100 Gy/h, whereas other bacteria, Archaea, and fungi can be resistant to several kGy of acute IR. D. radiodurans exhibits resistance to acute IR up to 15 kGy, to 60 Gy/h of chronic radiation, and also to high levels of resistance to UV-C irradiation (100–295 nm), desiccation and oxidative stress. Thus, regarding the example of Deinococcus radiodurans, it can be argued that it efficiently and rapidly repairs DNA damage caused by IR. A number of genes have been identified whose expression is activated after irradiation, namely those encoding proteins associated with (1) efficient DNA repair, (2) protection against oxidation and (3) DNA supercoiling, which helps to maintain DNA integrity after irradiation [65]. More recently, it was demonstrated that in this organism, the adaptation to dryness and desiccation is at the basis of its radioresistance [64].


Nonetheless, it has been reported that Deinococcus’ ability to repair DNA damage results from a selective pressure other than ionizing radiation, because there are no terrestrial environments subjected to the levels of radiation it tolerates. Still, the information gathered, albeit with some degree of uncertainty, has contributed to a better understanding of the mechanisms of radioresistance in other organisms, making this an excellent model organism to unravel these mechanisms [66].


Studies have shown that the DNA repair systems used by D. radiodurans are less complex than those of radiation sensitive bacteria, namely Bacillus subtilis, a spore-former species and Escherichia coli. Transcriptomics studies revealed that in response to γ-radiation, specific genes involved in damage response are activated (ddrA, ddrB, and irrE (pprI)). PprI, for instance, regulates the expression of the recombinase recA and pprA, which is a protein involved in DNA ligation that is essential for the radiation resistance exhibited by D. radiodurans. Strains lacking pprI show impaired genome recovery [66]. Another important DNA repair system involves the synthesis of long and single-stranded overhangs, a process referred to as “Extended Synthesis-Dependent Strand Annealing” (ESDSA).3 The process allows the reconstruction of a functional genome from the chromosome fragments produced by the exposure to radiation. Accordingly, the process is used by the RecFOR pathway to repair DNA double strand breaks. To support these observations, strains mutated in the genes involved in the RecFOR pathway are susceptible to γ-radiation [67].


Laboratory experiments with Escherichia coli, and other mesophilic bacteria, have shown that these may become resistant to the chronic exposure to IR just by adding Mn2+ and orthophosphate to its growth medium, which spontaneously form potent Mn-antioxidant complexes. Another important factor associated with radioresistance is cell density. For example, in D. radiodurans high cell concentrations seem to exert a protective effect against a radiation dose of 67 Gy/h [60]. Still, further and more complete studies are required until we know all the phenomena that contribute to the radioresistance exhibited by microorganisms. One thing is certain, it results from the interplay of several factors.


Undoubtedly, multi-omics approaches (genomics, transcriptomics, proteomics, and metabolomics) will shed light and will further contribute to our understanding of the mechanisms involved in microbial radioresistance and detoxification. In order to contribute to a better understanding of the mechanisms involved in uranium resistance/tolerance, a recent high-throughput proteogenomic study was applied to bacteria of the genus Microbacterium, isolated from Chernobyl U contaminated soils and from natural U rich soils. The approach allowed the identification of proteins involved in membrane transport (e.g., ABC transporters and efflux pumps), phosphate (e.g., phosphatases involved in biomineralization) and iron metabolism (e.g., siderophores), in addition to a large percentage of proteins of unknown function, which reveals the complexity of this mechanism [68]. Still, in another study carried out with a member of the genus Geobacter exposed to 100 μM U, proteins involved in DNA protection, in efflux pumps of the RND family and in oxidative stress responses (e.g., SOD and superoxide reductase), were also identified. Exploring these recent approaches will certainly allow us to gain knowledge that will contribute to clarify this complex intricate process. Furthermore, they will allow the selection for the best microorganism(s) with the potential to clean-up these contaminated sites by more eco-friendly processes. So far, in addition to the above study, genomic approaches proved useful in the identification of key genes and their respective products, encoded in the genomes of microorganisms resistant/tolerant to radionuclides/metals and which are, therefore, involved in the detoxification of this contaminants. With this approach, U-resistant bacteria of the genus Burkholderia and fungi of the genus Penicillium have been identified. Transcriptomics studies, by giving access to the analysis of gene expression and regulation, have gained relevance in the area of bioremediation. The information gathered from this comprehensive analysis, and also from future studies employing these methodologies, will surely shed light on the mechanisms of microbial resistance/tolerance to radionuclides/metals, while helping in the identification and selection of microorganisms that can be employed for bioremediation purposes of radionuclide/metals contaminated sites [69].


Unlike most laboratory studies, environmental exposure to radionuclides, (e.g., NORM sites and nuclear power plant accident sites), includes different radiation types (α and β, as well as γ) combined with many other stressors (e.g., temperature, nutrients, toxic chemicals like metals, etc.) over long periods. Thus, in polyextremophiles, the response to the adaptation/resistance should be broader and involve an intricate crosstalk between the different cellular processes [70].


Culture-independent field studies have shown that radionuclide contaminated environments host a wide diversity of bacteria and that radionuclides strongly impact community function and structure. Recently, a metagenomics approach carried out in surface soil samples from Chernobyl and Fukushima, over a gradient of radionuclide concentrations (137Cs 1680—0.4 and 90Sr 209.1—1.9 kBq/kg), revealed that samples clustered according to the level of radiological contamination, irrespective of the collection site [71]. Nonetheless, a lower microbiota diversity was found in Chernobyl samples, which was expected as Chernobyl soils are more contaminated. The following were reported to be the most common phyla: Proteobacteria, Acidobacteria and Actinobacteria. Furthermore, as expected, the functions encoded by the genes identified seem to be related with stress, metal and radiation tolerance. For instance, genes involved in decontamination, DNA repair, information storage and processing, cellular processes and signaling and metabolism. A comprehensive listing of the function of the genes responsive to this type of contaminants has been recently reviewed by Hoyos-Hernandez and co-workers [71].


A similar approach was employed in a study performed by Theodorakopoulos and colleagues [72], in Chernobyl, which demonstrated the high diversity of bacteria in those contaminated sites. The same authors isolated cultivable bacteria of the genus Microbacterium that were employed in laboratory exposure studies, contributing to a better understanding of the mechanisms of tolerance to radionuclides/metals in those bacteria. The identified mechanisms involve biosorption, efflux and biomineralization [68].


Although further studies are required to better understand how radiological contamination exerts a selective pressure and how it shapes the structure of the microbial community, the sensitivity of the various organisms to radioactive contamination under environmental conditions generally exceeds the sensitivity of the same organisms to experimental laboratory exposures [62]. It is clear though that communities from soils of these contaminated sites have functional profiles that allow them to deal with this type of radiological and chemical contamination. Furthermore, these environments constitute a genetic pool from which the phylogenetic affiliation of cultivable and non-cultivable microorganisms can be determined, thus allowing the identification of new genes involved in the resistance to these contaminants, in addition to further contributing to clarify those mechanisms.


Plants are sessile organisms that cannot leave the surrounding environment if the ecological factors are not suitable for their growth. Thus, under unfavorable circumstances, plants have only the choice to perish or adapt to changing environments. The extreme physiological plasticity of plants allowed their diffusion in all ecosystems of the Earth and today we may have a comprehensive vision of the multitude of adaptations carried out by these organisms in diverse places. Indeed, plants such as other living organisms can adapt to cyclical natural disturbances over time, developing the capacity for endurance (resistance) and self-repair (resilience) in different ecosystems.


Laboratory and field studies showed that ionizing radiation may exert different effects on plant metabolism, growth and reproduction, depending on plant developmental stage at the time of exposure, plant physiological and morphological traits, as well as genetic characteristics [73, 74]. Moreover, depending on the dose or radiation type (low or high-LET), ionizing radiation induces detrimental outcomes at high doses, harmful consequences at intermediate levels and stimulatory effects at low doses.


In some cases, ionizing radiation exposure increases embryo lethality, induces dwarf architecture and modifies floral elements [74] and literature herein. Other studies indicated that some irradiated crops showed a taller architecture, increased yields and reproductive success and the ability to endure water shortage [75, 76]. As for many other organisms, within plant cells, the nucleus is considered the primary site of injury by ionizing radiation, which is responsible for random DNA damage and generates different kinds of mutations, such as deletions, base substitutions and chromosomal alteration [74, 77]. There is a direct relationship between the radiosensitivity of a plant and the average volume occupied by a chromosome in the cell nucleus. If the chromosome volume is large, the plant will be more sensitive and, therefore, the dose of ionizing radiation causing severe damages is less. Hence, polyploid species exhibit a minor sensitivity to radiation damage because gene redundancy protects polyploidy from the deleterious effect of mutations [78]. Besides plant cells, it is noteworthy that ionizing radiation may have different impacts on organs and tissues. Generally, more complex tissue architecture is less sensitive to damage; thus, young tissues are more vulnerable than old [73, 79]. At functional level, many studies have evidenced that radiation is dangerous for the photosynthetic apparatus. Generally, a decline of photosynthesis often implicates damage to photosystem II (PSII) and in particular to D1 protein, implicated in the right functioning of photosynthetic electron transport. Together with the impairment of PSII, a significant decrease of photosynthetic pigments and enzymes of the carbon assimilation cycle was also detected [73].


The majority of information on the impacts of radioactivity on plants comes from studies carried out by scientists after the nuclear disasters of Chernobyl (Ukraine) in 1986 and Fukushima (Japan) in 2011 [80].


Since 1986 the Chernobyl red forest has represented a living laboratory for biologists to study for long-lasting plant behavior in response to acute and chronic radioactive contamination. The name “Red Forest” comes from the ginger-brown color of the pine trees as a result of the high radiation levels immediately after the explosion of the nuclear plant. Studies continued in the post-accident period and enlarged the knowledge on the effects of acute and chronic radiation on plants [81]. Generally different plant species show diverse sensitivity to radiation, being shrubs more resilient than conifers. The sensitivity of the pine compared to other tree species was most apparent in the Chernobyl exclusion zone and trees showed dramatic alterations in the morphology of trunks and branches, indicating damage at meristems level [82]. Following the Fukushima accident, despite the much lower exposure levels, Japanese red pine (Pinus densiflora Siebold & Zucc.) and Japanese fir (Abies firma Siebold & Zucc.) species showed developmental anomalies similar to those observed in Chernobyl [83, 84]. However, it is uncertain if the aberrations observed in Chernobyl are due to direct effects of radiation on the trees or multiple stresses due to biotic and other abiotic factors.


It is noteworthy that the quantity of radionuclides absorbed by plants depends on their phenological stage and growth status which, in turn, varies with the pedo-climatic conditions and cultivation factors. Once deposited on the vegetation and in particular on the leaf surface, the radioisotopes are absorbed through stomata and then transported to the other organs including fruits, thus possibly entering the food chain through edible leaves and fruits [85].


Today the Red Forest remains one of the most contaminated sites globally, and the surrounding forest area also represents an area of active research and scientific interest because of the return of wildlife in the exclusion zone. Here, the understory vegetation and deciduous (silver birch) trees have reappeared, but radioactive dust still remains stored in plant biomass and soil, for the very slow matter cycle.


The occurrence of revegetation has proven to be remarkably resilient to the intense radiation around the nuclear disaster zone. The exclusion zone is now dominated by grasslands and shrublands, while the most representative trees are Scots pine and silver birch Betula pendula [74] and literature herein.


Recent studies suggest that plants subjected to not-lethal doses of ionizing radiation show an increased resistance to other environmental stresses. Two strategies have been hypothesized, namely the production of ROS-mediated cell signaling and/or a boost of secondary metabolites [86].


The resilience to radiation in plants of the Chernobyl exclusion zone and from most contaminated sites at Fukushima is due to different mechanisms to protect the genetic material, improving the plant radioresistance [80]. Generally, plants are more radioresistant than animals because they present integrated adaptation mechanisms at genetic, anatomical, and physiological levels.


At genetic level, mechanisms include the regulation of expression of some genes encoding for radical scavenging and DNA-repair enzymes, homologous and non-homologous recombination, and the activation of scavengers. The higher stability induced by polyploidy, typical among plant kingdom, enhances radioresistance thanks to the presence of several copies of the same genes, which may serve as additional wild type copies in the case of radiation-induced injuries [87]. At the structural and metabolism level, plant cells present some traits such as thickened cell walls, cuticles, pubescence, increased deposition of phenolic compounds around membranes [88, 89]. At the anatomical level, complex tissue organization is associated with high resistance to mutagenic effects and the capability to adopt repair mechanisms.


Non-lethal doses of ionizing radiation may also induce hormesis improving plant defense against stressors, through the stimulation of the production of antioxidant enzymes (SOD, CAT, APX) or morpho-anatomical and photosynthetic changes that favor plant growth and metabolism [74, 90, 91].


Radiation-induced hormesis is still an unclear phenomenon in plants because it strongly depends on species intrinsic characteristics. At present, further studies are in progress to understand if it is a sort of compensation to irradiation damage or a transitory change, not enough to induce permanent injuries.


Invertebrates have been considered a relevant group of organisms for studying the effects of ionizing radiation, both focusing on mechanisms of action and on previewing impacts in natural communities. Several reasons can be enumerated for choosing aquatic and terrestrial invertebrates as model organisms for IR studies, namely:1.
They have long served for providing insights into fundamental mechanisms of development, biomedical research (e.g., neurobiology, basic physiology, genetics, immunology, cancer biology), species diversification and genome evolution (e.g., Drosophila melanogaster, Caenorhabditis elegans; planarians and crustaceans) [92–95]; for studying the effects of ionizing radiation in neuronal function [96] and as model organisms in radiation hormesis studies [97].
 2.
Due to their important role in food webs, transferring carbon from producers to higher trophic levels (i.e., cladocerans, copepods), as detritivores contributing for degradation of organic matter through comminution (e.g., oligochaetes) and turnover of microbial communities (i.e., bacterivorous nematodes).
 3.
The role of some species as ecosystem engineers dynamically working the structure of soils and sediments (i.e. oligochaetes, polychaetes, ants) and the contribution for other soil and sediment functions.
 4.
The sensitivity and the ease of culture for some invertebrate species under laboratory conditions, as well as proliferation, producing a great number of individuals for testing in complex experimental designs and without tight regulatory requirements.


They have long served for providing insights into fundamental mechanisms of development, biomedical research (e.g., neurobiology, basic physiology, genetics, immunology, cancer biology), species diversification and genome evolution (e.g., Drosophila melanogaster, Caenorhabditis elegans; planarians and crustaceans) [92–95]; for studying the effects of ionizing radiation in neuronal function [96] and as model organisms in radiation hormesis studies [97].


Due to their important role in food webs, transferring carbon from producers to higher trophic levels (i.e., cladocerans, copepods), as detritivores contributing for degradation of organic matter through comminution (e.g., oligochaetes) and turnover of microbial communities (i.e., bacterivorous nematodes).


The role of some species as ecosystem engineers dynamically working the structure of soils and sediments (i.e. oligochaetes, polychaetes, ants) and the contribution for other soil and sediment functions.


The sensitivity and the ease of culture for some invertebrate species under laboratory conditions, as well as proliferation, producing a great number of individuals for testing in complex experimental designs and without tight regulatory requirements.


Aquatic invertebrates as benthic organisms and invertebrates living burrowed in soils or dwelling at the surface are among the group of organisms that may receive the highest radiation doses, since these environmental compartments are relevant environmental sinks of radionuclides. The mechanisms of action and the subsequent effects of ionizing radiation in invertebrates have been addressed mainly since the seventies, with a limited number of species, through laboratorial exposures to gamma radiation of single species, frequently at high-dose rates, with few environmental relevance for chronic exposure scenarios [98]. Real conditions include exposures to industrial radionuclides in areas affected by nuclear accidents, nuclear power plants, or in nuclear test sites, as well as through exposures to natural occurring radionuclides (NORs), as those found in uranium mining areas. In the later areas, the effects of radionuclides, mainly alpha-emitters, cannot be distinguished from that of metals, also present at high levels in the affected environmental matrices. The same difficulty exists in areas of nuclear accidents as the Chernobyl exclusion zone, where the release of different artificial radionuclides has occurred, although data available for activity concentrations in biota are almost limited to 90Sr, 137Cs, and some few other radionuclides [38].


Invertebrates are among the least sensitive organisms to ionizing radiation [62, 99]. Cassidy and co-authors [100] suggested that the reasons for these differences in sensitivity, between organisms of different taxonomic groups, may  include differences in DNA content, DNA repairing processes, and kinetics of cell cycle, within other aspects. The doses able to cause mortality or decrease life span are species dependent and frequently very high: as for example above 1000 Gy for Caenorhabditis elegans [101]. However, differences in sensitivity of different life stages were also reported (i.e., Johnson and Hartman [101]), with reproduction effects being seen at much lower doses (i.e., 4 mG/h for earthworm).


Ionizing radiation hormesis has been reported in a number of studies with invertebrates (dipterans, coleoptera), exposed to low doses from different sources (X-ray, gamma radiation, 137Cs) (see review by Vaiserman et al. [97]). Reduced mortality rates and long-life spans were highly dependent on the exposure conditions [102], for example, life-extended effects were only observed in house flies (Musca domestica) reared in groups, and thus under high locomotor activity and exposed to a 10 Gy dose. Several hypotheses were then postulated and tested to unveil the factors responsible for modulating radiation hormesis, using Drosophila melanogaster, as model species, as for example:  increased IR resistance, IR-induced sterility in females, apoptosis induction and changes in DNA repair genes and life-stage differential sensitivity were some of the proposals [97] and references quoted herein. X-ray irradiation of D. melanogaster eggs with 0.75 Gy, decreased the amount of DNA segments, by cleavage of S1 nuclease sensitive sites (<3 kb), resulting in a great DNA stability, changing the repair and/or transcription processes and thus affecting lifespan and the resistance of adults to IR [103]. Based on all the studies conducted, the radiation hormesis model proposes that the exposure to low doses of IR could induce several adaptive responses, which in turn will prevent environmental-induced health effects [97].


At the cellular level, oxidative stress and the activation of oxidative stress-response mechanisms have been reported as the major indirect consequences of exposures to IR of aquatic and terrestrial invertebrates. Won and Lee [104] observed a significant increase in the activation of several enzymes, as for example, superoxide dismutase (SOD), catalase (CAT), glutathione reductase (GR) and glutathione-S-transferase (GST) in the marine copepod Paracyclopina nana, exposed to gamma radiation doses equal or greater than 10 Gy (at a dose rate of 2 Gy/min). However, in this study, no data from additional molecular parameters, as those related with DNA damage or lipid peroxidation were provided, preventing us to infer if the activation of these enzymes was sufficient or not to prevent cellular damages. A dose-dependent increase in ROS was also recorded in another marine invertebrate species, as for example, the copepod Tigriopus japonicus and the rotifer Brachionus koreanus, for a range of concentrations from 50 to 200 Gy (irradiated at a dose rate of 2 Gy/min) [105, 106]. Concomitantly, the antioxidant response system was activated, and GST and GR activities were significantly increased for the copepod, while for the rotifer the same was recorded for the activity of GST. A cellular and lipid peroxidation (LPO)-related ROS was dose-dependent overproduction was also recorded in the freshwater cladoceran Daphnia magna after 8-day exposure to a dose rate of 100 mGy/h of gamma radiation. The overproduction of mitochondrial ROS was significantly enhanced at 40 and 100 mGy/h [1]. Dose rates of the same order of magnitude (10.7 and 42.9 mGy/h) were also able to cause lipid peroxidation in daphnids, after both 24 and 48 h of exposure. However, at the highest dose rate tested (106 mGy/h), the same effect was only registered after 48 h of exposure [55]. This observation, which was consistent with other studies (i.e., Fuller et al. [107]), gave rise to the hypothesis that ROS may also act as a signaling molecule, requiring a certain level within the cell to activate antioxidant defense mechanisms. Neutral lipid catabolism was also observed in the nematode C. elegans independently of the different doses and dose rates tested (7 and 52 mGy/h), and this effect was associated with a reduced longevity, as lipid homeostasis is responsible for endocrine signaling of longevity [108]. In fact, the up-regulation of different hormone receptors in daphnids was suggested as a signal of disruption of normal endocrine functions in response to IR exposure [1].


Regarding the interaction of ROS with proteins, Won and Lee [104] registered an upregulation of the hsp gene in the copepod P. nana, which was interpreted as being related with a possible response to protect key proteins (probably those involved in DNA repair signaling pathways) through the synthesis of chaperones.4 In the cascade of events promoted by ionizing radiation, Song and co-authors [1] recorded an enhanced expression of the Ube2 gene in D. magna, involved in the degradation of proteins, suggesting the activation of a mechanism responsible for the elimination of proteins damaged by ROS. A highly efficient antioxidant protection system may not be able to protect DNA from damage but it can delay protein carbonylation.5 Therefore, protecting cellular components involved in the repair of DNA double-strand breaks (DSB) was proposed as a main factor to explain the resistance of bdelloid rotifers to ionizing radiation [109].


DNA damage is a frequently reported effect in invertebrates exposed to IR from different sources. These damages can be either caused indirectly, mediated by ROS, or by direct deposition of radiation energy in DNA [1, 104]. In response to DNA damage, the expression of genes related with DNA repair systems (e.g., p53, RAD50, Mre11 coding for the DSB repair protein, Ku70, Ku80, and DNA-PK) was recorded in different invertebrate species, frequently with a non-monotonic response, but always with a significant and differential expression at low and higher dose rates (at 4, 100, and 200 mG/h) [1, 55, 104, 106]. In summary, genes involved in nucleotide excision, base excision, homologous recombinant, and non-homologous recombination repair pathways have been found to be all involved in the response of cells to IR. At high IR dose/dose rates, ROS may also induce DNA methylation,6 leading to the accumulation of damages, by silencing some genes. Song et al. [1] recorded an enhanced expression of DNA (cytosine-5)-methyltransferase 1 (Dnmt1), DNA cytosine-5 methyltransferase 3A2 (Dnmt3a2) genes, involved in maintenance of DNA methylation and in de novo DNA methylation, respectively, in D. magna.


The disruption of energy metabolism under the exposure to IR is another reported effect at the cellular level, once again in different invertebrate species [1, 55]. Direct interference with proteins of the electron transport chain, mitochondria ultrastructural changes caused by ROS and modulation of oxidative phosphorylation are within some of the mechanisms proposed, based on observations made in D. magna exposed to gamma radiation [55]. Genes encoding NADH dehydrogenase (Nd), succinate dehydrogenase subunit A (SdhA) of complex II, different cytochrome oxidase subunits (COX1, COX2, and COX3), cytochrome c oxidase copper chaperone (COX17) of complex IV and ATP synthase subunit mitochondrial (sun) of complex 4 were some of the genes involved in the electron transport chain found to be suppressed by gamma radiation [55]. At the end of the cascade of events triggered by gamma-radiation, the regulation of different apoptotic signaling pathways was observed in freshwater Cladocera, in parallel with DNA damage and regulation of repair mechanisms, cell cycle disruption and mitochondrial dysfunction [1, 55]. Although not significant, an increasing trend in apoptotic cell death with increasing dose rates of radiation was recorded in crustaceans, namely daphnids and in the Norway lobster (Nephrops norvegicus) cell cultures exposed to 60Co gamma-radiation [110]. Apoptosis is a downstream event, to oxidative stress and DNA damage occurrences, that is activated to eliminate damaged cells in an ultimate effort for protecting organisms.


The effects of ionizing radiation at the population level are poorly documented and it has been demonstrated that equal levels of effect at similar individual endpoints (e.g., growth or reproduction) may have different impacts on population dynamics [111]. Furthermore, it is still difficult to link the results of biomarkers of oxidative stress and genotoxic damage with phenotypic consequences (changes in morphology, growth, reproductive output, and viability of offspring) [112]. Data available allowed a tentative hierarchization of individual endpoints based on their radiosensitivity: mutation > reproduction > morbidity and mortality [113]. One step forward, modeling population responses it was shown that they differed depending on the affected individual reproduction endpoint (juvenile or adult survival, delay in maturity, or reduction in fecundity) [114]. Hatching was shown to be the most sensitive endpoint to chronic exposures to gamma radiation for aquatic invertebrates (EDR107 of 830 mGy/h for the polychaete worm Neanthes arenaceodentata) and fecundity for terrestrial invertebrates (EDR10 of 2600 mGy/h for Porcellio scaber). These species displayed similar EDR10 values for individual and population level endpoints (net reproduction rate). This was observed for the species that had a particularly sensitive individual endpoint.


The most concerning consequences of genotoxicity, that may support inferences about potential effects on natural populations, are those that affect the reproductive fitness of organisms. Reproduction has shown to be the most sensitive parameter in invertebrates (collembolans, worms, tardigrades, chironomids, and polychaetes) exposed to IR, when compared with survival or other endpoints at the individual level [109, 115–120]. It was suggested that the decrease in fecundity is not caused by the number of DNA DSB, but by the inactivation of the DNA repair systems [109]. In fact the incomparable ability of bdelloid rotifers to remain fertile, after extensive DNA damage, was attributed to high efficiency of repair systems and to mechanisms that protect proteins of these repair systems [121]. These authors also associated the resistance to ionizing radiation of these organisms with their resistance to desiccation resulting from their adaptation to ephemeral ponds. Desiccation, similarly to radiation, increases ROS production and DNA breakage.


Harrison et al. [112, 122], also working with the polychaeta N. arenaceodentata, hypothesized that chromosomal aberrations caused by gamma radiation doses of 2.0 and 4.0 Gy were responsible for gametal cell death and subsequent decreases in brood sizes of this species. In opposition, under laboratory conditions, significant effects were recorded in sperm quality, but not on sperm numbers, of males of the crustacean Echinogammarus marinus chronically exposed to doses rates of 1 and 10 mGy/day provided by the betta emitter 32P, for two weeks. Significant DNA damage was recorded in spermatozoa cells only at the highest dose rate. Furthermore, only a weak correlation was found between sperm quality parameters, fecundity, and embryo parameters analyzed [107]. Effects on ovary structure and oocyte development were also reported in the freshwater cladoceran D. magna in response to exposure to 1 and 100 mGy/h gamma radiation, dose rates.


Another possible cause of reproduction impairment in invertebrates, under exposure to ionizing radiation, may be related with the allocation of energy to molecular response mechanisms (e.g., activation of antioxidant defense system, DNA repair mechanisms) rather than to reproduction, with consequences on the fecundity of organisms [117].


An ED508 for reproduction of 21.9 Gy, one order of magnitude lower than that recorded for growth (144 Gy) was found for the collembolan Folsomia candida, under exposure to 137Cs gamma radiation at a constant dose rate of 8.3 Gy/min. Song et al. [1] also observed a non-monotonic reduction in the total number of offspring of the cladocera D. magna, concomitantly with no effects on survival, molting or ovulation frequency (at dose rates of 1 and 100 mGy/h). At the lowest dose, the effect on the cumulative reproduction output was mainly associated with an increase in the number of days needed to deliver four broods, while at the highest dose rate, the reproductive cycles were accelerated but the size of the broods was reduced. A similar observation was made by Parisot et al. [123] in the same organisms exposed to dose rates of 0.07–35.4 mGy/h of gamma-radiation, for 23 days. The same non-monotonic response was recorded D. magna representing 38 different genotypes collected in lakes located inside the Chernobyl exclusion zone with a range of dose rates between 0.1 and 181.2 mGy/h.


In a study conducted by Alonzo and collaborators [111], the freshwater species D. magna and the terrestrial earthworm Eisenia fetida, two species with different life history strategies (short lived/parthenogenic versus more long-term life/sexually reproducing hermaphrodite, respectively) were selected: (1) to model population growth in response to individual effects caused by the exposure to IR and (2) to investigate populations susceptibility using two different models to take into account single generation and multiple generation exposures. It was shown that in daphnids, the population growth was 1.5-fold more sensitive to changes in fecundity than in mortality. Daphnids population growth was also highly affected by delays in reproduction. Earthworms’ population growth was more sensitive to delays in reproduction, while effects in fecundity and mortality have a similar and lower impact on populations. Despite the different life strategies, the intrinsic rates of population increase were equivalent for both species, because the greater reproductive rate of daphnids is compensated by a shorter life span relative to earthworms.


After disturbances of great magnitude, the recovery of natural populations of cladocera may rely on the banks of resting eggs in the sediments of lentic systems. These resting eggs if irradiated may have its performance compromised, affecting the dynamic of the natural populations. Zadereev et al. [124] observed that although doses up to 100 Gy (variable dose rates) did not affect the survival and hatching of resting eggs of Moina macrocopa, the size and the structure of populations initiated from resting eggs exposed to this highest dose of gamma-radiation, were affected. Therefore, subsequent effects on the dynamic of the populations of this cladocera may be expected in lakes with highly contaminated sediments.


Under a real scenario of radionuclides contamination, no correlation was found between different reproduction endpoints (proportion of breeding females, fecundity, brood mass, maternal body mass) of the crustacean Asellus aquaticus, sampled at different lakes in the Chernobyl affected area, with the gradient of dose rates between 0.064 and 27.1 mGy/h registered at these ecosystems [107]. Also, upscaling to populations and communities, Murphy et al. [125] focused on the diversity of littoral macroinvertebrates communities at eight natural lakes in Belarus, with a range of external dose rates from 0.066 to 10.22 mGy/h once again did not find any correlation between population endpoints (abundance, taxon richness, Shannon-Wiener diversity index, and the Berger-Parker dominance index) and the range of external dose rates registered in the sampled lakes. This study suggested that the IR dose rates recorded had no detectable effects on the littoral macroinvertebrate communities of these lakes.