Genotoxicological studies in aquatic organisms: an overview
School of Biological Sciences, Plymouth Environmental Research Centre, University of Plymouth, Plymouth, PL48AA, UK
Abstract
Substantial progress has been made in the last two decades to evaluate the impact of physical and chemical genotoxins in aquatic organisms. This overview (a) summarises the major high lights in this stimulating area of research, (b) compares the developments in this field with the developments in mammalian genotoxicological studies, where appropriate, (c) introduces 18 different articles presented in this special issue of Mutation Research in the backdrop of main advances and, (d) hypothesises on future directions of research in this exciting field.
Keywords: Aquatic organisms; Eco-genotoxicology; Genetic ecotoxicology
1. Introduction
The quality of human life is directly or indirectly dependent on the health, vitality and vigour of the environment around us. Abiotic and biotic components, including air, water, soil, food and plants and animals constitute our environment, and through complex interactions, life is sustained. Deterioration in the health of any of these components can have serious knock-on effects on the continuity of life. The aquatic environment, which covers two-thirds of the planet, is inhabited by the majority of extant species in different ecological niches; moreover many of them are important sources of human food. It has been estimated that approximately 70% of the human population resides within 60 km of coastal regions and this percentage is increasing [1]. In addition, a significant proportion of the world’s largest cities are connected, either directly or indirectly, to the marine environment. The aquatic environment therefore plays vital roles for ecosystem functioning, human health and civilisation.
As a consequence of human population growth and industrial development, the production, consumption and disposal of anthropogenic chemicals and wastes continue to increase. The aquatic environment is often the ultimate recipient of this increasing range of anthropogenic contaminants, a large proportion of which are potentially genotoxic and carcinogenic substances. For example, during 1994 in the US alone, 22,744 different facilities released 2.26 billion pounds of toxic substances into the environment, of which approximately one-third were rodent carcinogens [2]. Most of these carcinogens were discharged as components of complex mixtures (e.g. liquid effluents, air-borne emissions, and solid wastes). It is therefore not surprising that in recent years there has been increasing concern regarding the genotoxic/carcinogenic hazards, and reports on the occurrence of malignancies in aquatic organisms [3], [4], [5], [6], [42], [45] and [96]. The occurrence of malignancies in aquatic organisms has also been correlated with a higher incidence of tumours in human populations and increased use of herbicides [7]. Therefore, natural biota might be used as sentinel or surrogate species for the evaluation of genotoxic chemicals in the environment and their risk to human health.
Although exposure of aquatic organisms to genotoxic contaminants could pose a risk to human health via the food chain, there is also an ecological risk that may lead to heritable mutations and loss in the total genetic diversity (either intra or inter species), with significant implications for long-term survival of natural populations [8], [9] and [10]. Mathematical models have suggested that fixation of mildly deleterious mutations could significantly contribute to loss of Darwinian fitness and eventual extinction of small populations [11] and [12]. However, although a correlation between air pollution, incidence of lung cancer and loss of biodiversity has been demonstrated [13], there have been no serious attempts to establish links between genotoxins-induced mutations and extinction of natural biota. A large number of mammalian germ cell mutagens have been identified and many of these agents have also been found to be somatic cell mutagens [14]. However, there is a fundamental gap in our knowledge on the long-term implications of these mutagens in the environment, including effects on population survival and recruitment.
When viewed against the backdrop of enormous reproductive surplus and natural wastage (due to predation, desiccation, changing salinity, temperature, acidity, etc.), it has been argued that any detrimental effects resulting from a localised pollution incident is likely to have a diminishingly small impact globally. This applies especially to invertebrates, which constitute more than 90% of the extant species and play important role in ecosystem functioning [8] and [15]. From a broader perspective, it is likely that combined effects of other stresses (e.g. changing temperature, acidity, salinity, oxygen concentration, etc.), loss of habitat and pollution (including genotoxin exposure) could produce detrimental conditions for population survival. Indeed, global climate change and other environmental stressors will have implications for mutagenesis studies [16]. However, in a very short geological time scale and in complex environmental conditions, where the contaminants occur in all probable combinations, it would be very difficult to pinpoint particular contaminants causing population extinction. Keeping aside the debate on the role of genotoxic exposure on population survival and scientific considerations in protecting the environment for sustainability, maintenance of diversity, environmental conservation and the fact that the health of human populations is dependent upon other species, it could be argued that the human population has both the moral obligation and the practical necessity to protect other species with which it shares its limited biosphere [17].
While substantial progress has been made in recent decades in defining the significance of exposure to genotoxins for human health risk assessment, there has been only very limited development with respect to understanding the impact of genotoxins on natural populations. This is despite the fact that Boveri proposed the relationship between chromosomal changes and the origin of tumours as early as 1914, using developing echinoderm embryos as a model [18]. In 1998, Mutation Research published two special issues dealing with the use of invertebrates and fish in genotoxicological studies [19] and [20]. In these special issues, attempts were made to record the advances in our capabilities to evaluate the interaction of genotoxins with the genetic material of aquatic biota and the development of realistic biological methods to identify expressed genotoxicity in these organisms. These issues covered the studies pertaining to biochemical and molecular end-points of genotoxicity, including alterations in carcinogen metabolising enzymes, DNA adducts, expression of oncogenes, use of fish in long-term carcinogenicity assays, including molecular dosimetry studies with a range of carcinogens and anti-carcinogens. The issues also reflected the ecological significance of genotoxic exposures in natural biota [19] and [20]. Since publication of these issues, considerable progress has been made in our capability to determine the impact of anthropogenic stress and in our understanding of the fundamental mechanisms involved in mutagenesis and carcinogenesis in aquatic organisms. In addition, there has been growing public demand for clean water, and concern over the presence of a variety of anthropogenic contaminants with diverse mode of actions in the aquatic environment. This has also led to significant regulatory developments across the globe. This special issue aims to update recent advancements in our capabilities to evaluate potential adverse impact of exposure of genotoxins in the aquatic organism. Authors (representing different laboratories) in this special issue have previously made significant contribution to this field. It is hoped that this special issue will provide a single source of current state-of-the-art information to the readers and workers in the field.
2. Genomics and mutagenesis
One recent advance in the field of genetics is the mapping of the human genome and other model species. The technological developments associated with this international project have further stimulated research in our field to elucidate the nature, cause and consequences of mutations. It is often not appreciated that the idea of mapping the human genome grew from the need to improve our ability to detect and study induced and inherited mutations [21] and [22]. The field of mutagenesis is indeed entering into an exciting phase and it is hoped that ‘toxicogenomics’, the integration of genomics, bioinformatics, and toxicology, will accelerate our understanding of the impact of anthropogenic agents at the molecular level and assist in hazard and risk assessment. With the aid of new technologies, understanding of changes in gene structure and expression would help us to evaluate the impact of anthropogenic stress and the potential of natural biota to adapt and survive in changing environments. Several critical genes are being identified, isolated, and qualitatively and quantitatively assayed to measure the impact of anthropogenic stress in the natural biota. It is envisaged that gene expression profiling may reveal molecular mechanisms of action of chemicals, despite the inherent limitation of genomic and proteomic experiments, which measure single end points (i.e. RNA or protein levels), albeit for thousands of genes at a time [23]. In this context, an important approach has been to use the gene expression profiling with DNA microarrays, to screen large numbers of genes in response to environmental changes. Although the time and high cost for analysis and interpretation of the data are one of the inherent drawbacks of this technology, the workers in the field are showing growing awareness of this type of tools as an alternative to classical toxicological tests for rapid screening. However, one can appreciate that currently, there is a significant knowledge gap in our understanding of the molecular events that govern toxicologically relevant outcomes. In ecotoxicological terms, one of the major difficulties will be to link these molecular effects with ecological consequences [24]. Two papers in this issue have used this technology to measure gene expression in two different fish species following exposure to different contaminants.
Genotoxic potential of pulp mill effluents has been evaluated in vitro and in vivo by several workers [25] and [26]. In this issue, Denslow et al. [27] exposed spawning largemouth bass (Micropterus salmoides) for short and long terms to various concentrations of paper and pulp mill effluents to evaluate gene expression and steroid levels. Exposure scenarios were phased such that exposure ended at the same time. Gene expression was mainly determined by differential display (DD) reverse-transcriptase (RT)-polymerase chain reaction (PCR). The data showed that a set of genes in female fish that were up-regulated during vitellogenesis, were turned off by exposure. A comparison was made with other fish exposed to estradiol to provide a global overview of the changes in gene expression associated with pulp mill effluent treatment and also allowed the isolation of differentially expressed transcripts. One of these transcripts, CYP1A was cloned, characterised by sequencing and used to develop conventional RT-PCR and real-time PCR assays, both of which confirmed its induction in treated males. In addition, levels of circulating hormones and vitellogenin were measured in exposed animals and a Northern blot was used to quantify vitellogenin mRNA expression.
In recent years, there has been growing concern over the presence of those contaminants which interrupt hormonal metabolism in human as well as in aquatic species [28]. Genotoxic and carcinogenic potential of some steroidal hormones in humans are well-documented [29]. It appears that in common with other toxicants, endocrine disrupting agents may exert their effects simultaneously via several mechanisms, either in parallel or in series. In this context, tri-n-butyltin (TBT), a known endocrine disrupting agent for some gastropods, has also been shown to be genotoxic, cytotoxic, immunotoxic and teratogenic in some marine invertebrates [30], [31] and [32]. In addition to the studies by Denslow et al. [27] on large mouth bass mentioned above, in this issue Brown et al. [33], describe macroarray experiments designed to test gene expression in liver of male plaice (Pleuronectes platessa) as a function of time following exposure to ethynyl oestradiol (EE2). The authors first obtained genes by suppression subtractive hybridization and then prepared macroarrays on nylon membranes. Male fish were exposed to a constant concentration (20 ng l?1) for 21 days and then were transferred to clean water for 10 days. Interesting observations suggested that there are two different temporal expression patterns for estrogen-controlled genes. One pattern evident for vitellogenin (VTG) 1 and 3, and zona radiate protein (ZRP) 1 and 3, peaks by day 16 but then declined even though exposure continued for another 5 days. A second pattern evident for Vtg2 and ZRP2 showed a maximum expression from days 16–22 and then declined only when EE2 was removed. It is not clear why the expression of genes initially up-regulated by oestrogenic exposure should then decline in the continuing presence of EE2.The authors speculate a feed-back mechanism which turns transcription off or destabilises the transcripts. Such a mechanism could operate at the molecular level by direct effects on transcription factors or RNAses or may influence hypothalamic pituitary signalling. Given that the molecular mechanisms of endocrine disruption are still being explored, there is certainly a need to evaluate the gene profiling and potential mutagenic effects of these agents in different aquatic organisms.
3. Oncogenes and tumour suppressor genes
Compared to mammals, there have been limited studies pertaining to identification and expression of oncogenes in fish and invertebrates [34], [35], [36] and [37]. Available information suggests that several of these genes have a high degree of nucleotide sequence and deduced amino acid similarity with the mammalian gene counterparts [38], [39] and [40]. While some information is available in the literature pertaining to structure and expression of oncogenes in aquatic organisms, there is a paucity of information on anti-oncogenes or tumour suppressor genes. The significance of tumour suppressor genes in human cancer is well documented. For example, mutations in the p53 gene are known to play an important role in over half of human cancers [41]. As mentioned above, although induction of tumours in fish and many invertebrates are well documented [3], [4], [5], [6], [42], [45] and [96], and different fish species have been used as models for carcinogenic studies, there are only a few studies on the structure and function of anti-oncogenes in aquatic organisms. The p53 gene has been sequenced from different fish species for potential use of mutations in the highly conserved domains to identify potential genotoxins in the aquatic environment. Comparisons of the deduced sequences have revealed a high homology within the conserved DNA binding domain among different fish species and the very high sequence conservancy in fish is similar to that in mammalian p53 [43]. However, it is important to study the functional relationship between tumours in fish and alterations in five conserved domains, four of which (II–V) are found in the central portion of the protein (specific DNA binding domain). Attempts to find mutations in these regions have not been very successful under in vivo conditions. Using a yeast in vitro exposure system, Cachot et al. [44], have identified and characterised both spontaneous and potentially carcinogen-induced mutations in the flounder (Platichthys flesus) p53 gene. Given that the detection of p53 mutations under in vivo conditions has been difficult, this study attempts to describe a necessary step to identifying potential ‘hotspots’ for benzo(a)pyrene diol epoxide (BPDE)-induced mutation. Although the situation in vivo may well differ from this in vitro study, the study would certainly contribute to our search for p53 mutations in non-human species.
Among invertebrates, the incidence of gonadal tumours in soft-shell clams (Mya arenaria) near contaminated sites is well documented [45]. However, attempts to induce the tumours by laboratory exposures to 2,3,7,8-tetrachlorodibenzo-p-dioxin (dioxin, TCDD) have been unsuccessful [45]. It was hypothesised that tumour promotion in clams occurs through the aryl hydrocarbon receptor (AHR) by mechanisms comparable to those described for vertebrates. Characterization of the clam AHR revealed that it does not bind the prototypical AHR ligand, TCDD [46]. Subsequently, it was shown that exposure of clams to dioxin under laboratory conditions leads to induction of differential gene expression [47]. In continuation of the earlier studies, Olberding et al. [48], in this issue, attempted to further elucidate the mechanisms of tumour formation involving a protein with significant sequence similarity to mammalian E6AP, a homologus to E6AP carboxy terminus (HECT) E3 ubiquitin-protein ligase. E6AP, in association with the high-risk human papillomavirus (HPV) E6 protein, is known to be involved in the abnormal degradation of the p53 tumour suppressor protein in human cervical cancer. Adopting different approaches, the authors have provided a detailed functional study of a protein that may prove crucial in understanding environmental non-AHR mediated tumourigenesis in clams and other aquatic organisms.
4. Adaptation, resistance and chemosensitisation
The term ‘stress’ is usually applied to situations where the ‘fitness’ of individuals (or populations) is reduced because of changed environmental conditions. In ecotoxicological terms, this fitness would include ‘Darwinian fitness’ (i.e. growth, fertility and fecundity) of populations. While stress is normally considered to be environmentally induced, either by physical (e.g. climatic factors and pollutants) or by biotic factors (e.g. parasites, competitors, predators), it could also be intrinsic in origin. Such intrinsic stress could originate from an increase in homozygosity as a result of genetic drift and/or inbreeding. These homozygous conditions, in conjunction with natural selection will cause genetic stress that could lead to a decrease in fitness [49]. Several mechanisms have been suggested, including an increase in homozygosity for recessive deleterious alleles and a decrease in the level of heterozygotes for overdominant loci [50]. Other mechanisms, such as a decrease in the level of genomic coadaptation or an increase in genomic instability have also been suggested to play important roles in decreased fitness [51] and [52]. From an environmental conservation point of view, many small populations are subject to genetic stress due to their small size resulting from drift and inbreeding. Given that many species of natural biota have become endangered, studies on the processes associated with genetic stress are becoming increasingly important.
In contrast to genetic adaptation or resistance, which is transmitted to future generations, organisms may acclimatise physiologically at the individual level and could become more resistant as a consequence of earlier exposure [53]. This phenotypic plasticity, which results due to epigenetic mechanisms, may modify the toxic effects of contaminant exposure in the natural biota. Distinguishing between these two phenomena (i.e. genetic and epigenetic) is important in predicting the potential short and long term costs to affected populations. Wirgin and Waldman [54], have attempted to summarise our current understanding of the mechanistic basis of resistance or tolerance in several North American fish species. The toxicity mediated by a range of aromatic hydrocarbons was also correlated to CYP1A1 and early life stage toxicity. This particular study has suggested that while some resistance could be genetic and hence transmitted to at least the F2 generation, others could be physiological due to some unidentified epigenetic mechanisms. In addition, the authors also observed differential resistance for different groups of contaminants (e.g. aromatic hydrocarbons and halogenated hydrocarbons). They have also attempted to discuss resistance in fish in the format of case studies providing new data on resistance in tomcod and pointing out some of the costs that have or have not been associated with tolerance.
The phenomenon of multixenobiotic resistance (MXR) in aquatic invertebrates is well-documented [55] and [56], and explains the apparent simultaneous resistance of many aquatic invertebrates to multiple xenobiotics in the environment. The molecular mechanism of MXR has been suggested to be similar to the well-known multidrug resistance (MDR) phenomenon in tumour cells resistant to chemotherapeutic drugs. As with MDR, the MXR-mechanism in aquatic organisms pumps out xenobiotics, including anthropogenic chemicals and hence prevents their toxic effects. However, it has been suggested that many chemicals, called chemosensitizers, may inhibit the function of this fragile mechanism. As a result, the organisms could accumulate xenobiotics, including genotoxins, and inflict damage in the DNA and other cellular systems [56]. Although these chemosenitizers could play an important role in determining the toxicity of anthropogenic chemicals in the natural environment, and given that there could be a large number of these chemicals in the environment (of both natural and anthropogenic origin), their ecotoxicological significance remains to be established. While the field of MDR is one of the intensively researched areas in cancer biology, only a limited amount of information is available in aquatic organisms and most of this information is restricted to invertebrate species. In this issue Smital et al. [57], have further extended their study to elucidate the ecological significance of MXR inhibition under in vivo conditions. They have demonstrated that such inhibition could lead to enhanced production of mutagenic metabolites in mussels and apoptotic cells in sea urchin embryos following exposure to model MXR inhibitors. They have also attempted to critically analyse the questions (a) whether the inhibition observed after exposure to environmental samples is due to saturation of MXR transport proteins or due to presence of potency of MXR inhibitors and, (b) whether potent environmental MXR inhibitors are natural or man-made? These novel studies will further contribute to our understanding of molecular mechanisms of MXR, chemosensitisation and their ecotoxicological significance.
5. Heritable mutations and transgenerational effects
The transmission of genetic damage to offspring is a primary concern for scientists and regulators in human health arena. For mammalian genetic toxicology, guidelines have been presented which involve a weight-of-evidence approach to determine the heritable mutations hence identification of potential human mutagens [58]. These guidelines broadly take into account the intrinsic mutagenic potential of chemicals, their ability to reach the differentiating or differentiated germ cells through the blood-gonad barrier and interact with germ cell DNA to induce heritable mutations in mammalian species. While several approaches have been adopted and models have been developed for germ cell hazard and risk assessment in rodents and in fruit flies (e.g. mouse specific locus and Drosophila sex-linked recessive lethal tests), not much progress has been made to evaluate the potential hazards and risks from germ cell mutagens in aquatic organisms. In addition, in recent years, it has also been realised that exposure of somatic cells to ionising radiation could give rise to genomic instability which could be expressed after several generations of the cell cycle [59]. However, little is know about transgenerational effects in germ cells which could be manifested in the offspring.
In this issue, Shimada and Shima [60], who have worked extensively on the Japanese medaka fish (Oryzias latipes) model, extended their work on the transgenerational effects of radiation on somatic mutation frequency in developing embryos. This novel, non-mammalian system facilitates the mutants to be analysed as mosaics during embryogenesis that have both orange and white leucophores (pigment cells) in +/wl heterozygous offspring. In other words, in this system for germ cell mutagenesis, embryos that have both wild-type orange leucophores and mutant white leucophores are scored as mosaic mutants. Using this system, the authors found a threefold increase in mutant frequency of (non-irradiated) maternal alleles following irradiation of sperm or late spermatid but not with irradiated spermatogonial stem cells. These somatic mosaics showed higher embryonic lethality, indicating that mutants result from gross chromosomal rearrangements that lead to lethality. Interestingly, the authors also found that this indirect genetic instability is not heritable in the F2 generation. This could be attributed to large chromosomal rearrangements that suppress recombination and subsequent mosaic mutants in the F2. The data suggested that irradiation of sperm and late spermatids can induce indirect mutations in F1 somatic cells, which led the authors to hypothesize that genomic instability arises during F1 embryonic development. This study in a model fish system supports similar findings pertaining to untargeted mutation in maternal alleles at tandem repeat DNA sequences arising in mice born from irradiated spermatozoa [61].
Polycyclic aromatic hydrocarbons (PAHs) are known genotoxins and carcinogens. Contamination of the marine and estuarine environments by PAHs is a global phenomenon. An estimated total input of 230,000 metric tonnes of PAHs is being released annually in the marine environment [62]. There is no indication that this input will decrease in the coming decades, bearing in mind accidental or deliberate oils leaks, spills, refinery operations and industrial discharges [63] and [64]. PAHs are generally known to accumulate to high tissue concentrations in invertebrates at the bottom of the food chains, where uptake rates greatly exceed rates of metabolism and elimination, compared with vertebrates, where metabolism and elimination can cope with uptake [65]. The presence of PAHs in the aquatic environment is of concern since they induce acute toxicity in organisms and the presence of PAHs in the sediments have been linked to liver neoplasms and other abnormalities in benthic fish species [66] and [96]. PAHs (e.g. benzo(a)pyrene and dimethylbenzanthracene) have also been shown to induce dominant lethal mutations in mice following intraperitoneal injection [67] and air pollution (which contains appreciable amount of PAHs and particulate matters) has been linked to higher rates of inheritable mutations in herring gulls (Larus argentatus) and in laboratory mice [68], [69] and [70]. However, there have been virtually no studies to evaluate inheritable mutations or transgenerational effects in aquatic organisms following exposure to PAHs, particularly among the invertebrates. This could be due to lack of characterisation of genomes and the reproductive strategies of the model invertebrate species. Using a fresh water, parthenogenetically reproducing clonal organism, Daphnia magna, in this issue, Atienzar and Jha [71] have evaluated the repair kinetics and transgenerational effects following exposure to B(a)P using randomly amplified polymorphic DNA (RAPD). The study suggested that while some of the changes in the DNA profiles were reversed in the offspring, indicating an efficient repair system, some of the changes were transmitted to the offspring. The exposure of B(a)P was also correlated with the reproductive success of the organisms. A clear link between induced genetic damage and potential population level effects has been demonstrated, for which there is lack of information in the literature. Despite the criticism of the assay for its lack of reproducibility [72], the authors claim that after rigorous optimisation of the PCR parameters, the assay performs well, qualitatively and quantitatively, and its reliability could be strengthen as suggested by other workers [73] and [74]. In fact, several studies have used PCR-based techniques to assess genetic variation and changes in aquatic biota exposed to different contaminants [75], [76], [77] and [78] and it has been suggested that the RAPD assay can detect mutations only if they occur in at least 2% of the DNA [79]. Given that most of the natural biota are poorly understood in terms of their genetic make-up (e.g. microsatellites, minisatellites, karyotypes, cell turn-over rates, etc.), it is likely that combined with robust experimental plans and analytical tools, such molecular approaches will play an increasingly important role in determining the impact of genotoxins in the aquatic environment [80].
6. In vitro studies in genetic ecotoxicology
The fundamental role of genotoxic agents as the primary basis for the development of malignancy in mammalian species is now well accepted. This knowledge has led to a tiered strategy for the detection of mammalian genotoxins in order to protect human health. It is therefore logical to utilise the insights of mammalian genotoxicology in genetic ecotoxicology. In this context, the three-tier approach of mammalian genotoxicology initially utilises a suite of in vitro tests (e.g. bacterial assays, mammalian cell cultures) for evaluating the inherent genotoxicity of a substance followed by expressed genotoxicity under in vivo conditions. Finally, if the substance is a somatic mutagen to mammals in vivo, tests are conducted to measure the interaction of the substance with germ cell DNA or other heritable mutations under in vivo conditions [58]. For eco-genotoxicology, or genetic ecotoxicology, a clear distinction is required between testing and environmental monitoring [15]. For testing the intrinsic genotoxic potential of environmental samples, several bacterial tests (e.g. Ames test, UmuC test, Mutatox? and SOS chromotest, etc.) are being employed for hazard assessment [2].
The use of in vitro systems in ecotoxicological studies provides the opportunity not only for extrapolation from in vitro to in vivo systems, but also information on biological responses at higher levels biological organisation [81]. Fish play a major role for the flow of energy in aquatic ecosystems, are exposed continuously to contaminants in the natural habitat and constitute an important part of human diet, especially in coastal regions. In this respect, cells derived from liver, gonads and skins (e.g. primary hepatocytes, rainbow trout gonads and skin) of fish species have been extensively used to investigate toxicity of several reference genotoxins and environmental contaminants, such as organic extracts of sediment samples [81], [82], [83], [84] and [85]. These cells retain some important traits of fish (e.g. poikilothermic behaviour, unique xenobiotics metabolism and low rate of repair mechanism) [84]. In addition, due to ethical reasons, there is a move to minimise the use of vertebrates, including fish for experimental purposes. Furthermore, due to the lack of established cell lines from aquatic invertebrates, there will be increasing use of fish cells in ecotoxicological studies. In this context, rainbow trout gonad cells (RTG-2), a permanent cell line capable of metabolising the xenobiotics without the need of an exogenous metabolic system has been shown to provide reproducible results in inter-laboratory validation tests for cytotoxicity [81]. Compared to mammalian cells they have been shown to be more sensitive for the induction of genetic damage [86]. In this issue, Castano and Becerril [87], using randomly amplified polymorphic DNA (RAPD), as used earlier by Atienzar and Jha, in water fleas, D. magna [71], have extended their study on rainbow trout gonad (RTG-2) cells to evaluate the genotoxic potential following exposure to a range of concentrations of B(a)P for different periods. Qualitative and quantitative analysis of RAPD bands suggested that this in vitro system is useful in evaluating genotoxic effects of direct as well as indirect acting mutagens following either acute or chronic exposures.
7. In vivo studies and biomonitoring
While bacterial or in vitro systems could be used as screening tools to define the intrinsic genotoxicity of a substance, definitive ecotoxicological risk assessments should consider the expressed genotoxic activity in ecologically relevant organisms. These will take into account environmentally realistic routes of exposure, the effects of metabolism and DNA repair efficiency [88]. However, despite growing concern over the presence of genotoxins in the aquatic environment, there is a lack of adequately validated test methods, which could be used effectively to evaluate genotoxicity, and associated toxicity, in aquatic organisms under environmentally relevant conditions. The development of such in vivo test systems is also essential in providing a scientific basis for comparing the relative risks of man-made versus natural genotoxins [88], given that large number of marine plants and animals produce very potent genotoxins [89].
Biomonitoring for genotoxic risk evaluation is defined as frequent or even continuous gathering of information in a given population that is relevant to that population’s health risk [90]. The information gathering ranges from information on exposure, carcinogen-induced phenotypic changes (e.g. DNA adducts or gene mutations), neoplastic lesions and cancer occurrence. The main objectives of population monitoring in cancer research is therefore to identify carcinogenic mechanisms by obtaining more precise estimates of the extent to which different factors (e.g. environmental, life style, genetic susceptibility) contribute to human cancer burden and potential germ cell mutations [90]. This information ultimately helps to prevent cancer and reducing the risk of inheritable mutations. In contrast to human health risk assessment, until recently, the integral effects of contaminants relied heavily on analytical techniques, in particular for environmental monitoring and effluent discharge compliance. However, simple detection of environmental contaminants by these techniques is not sufficient unless their biological effects are also properly evaluated. In addition, genotoxins in the environment can occur as complex mixture, and the risk associated with such mixtures cannot be adequately anticipated on the basis of the effect and behaviour of individual components. Since biological systems are the target of toxicant action, they could provide important information which is not readily available from chemical analyses of the environmental samples [88]. Thus, in view of the limitations of chemical monitoring, the importance of biological monitoring is being widely recognised.
Several manuscripts in this issue deal with in vivo laboratory and field studies in order to evaluate the potential impact of contaminants in different aquatic organisms. The fundamental difference between biomonitoring studies involving human and natural populations appears to be that while the former is question driven, the later is technique driven. In addition, most of the techniques used for aquatic species have been adopted from human studies. Bolognesi et al., in this issue [91], studied the effects of environmental contaminants in Mediterranean mussels, Mytilus galloprovincialis, at four different sites along the Ligurian coast, Italy. These mussels included native as well as transplanted (caged) specimens. Induction of micronuclei and DNA strand breaks (using alkaline elution technique) were used as biomarkers of genotoxicity. In addition, the levels of polycyclic aromatic hydrocarbons (PAHs) and heavy metals (e.g. mercury and cadmium) were also analysed. The wild mussels showed accumulation of contaminants as a function of pollution gradient and this was found to be correlated with the induction of micronuclei. The mussels showed seasonal variability for the induction of genetic damage, caged mussels showing higher damage than the wild samples.
Using single cell gel electrophoresis (Comet assay) and DNA adduct analysis (32P-post labelling), Winter et al. [92], evaluated genotoxic effects in feral and caged chub (Leuciscus cephalus) from three rivers with different water quality (due to presence of organics, metals and pesticides) around Birmingham, UK. In general, elevated levels of DNA damage was recorded with a decrease in chemical water quality, in both feral and caged animals indicating an impact of chronic pollution. Recorded seasonal DNA adduct data suggested a higher degree of damage in the feral compared with caged animals. Apart from chemical parameters, both the studies of Bolognesi et al. [91] and Winter et al. [92] suggest that both in fish and mussels, seasonal variation could play an important role in the observed genotoxic effects in freshwater as well as in the marine environment.
Lyons et al. [93], have carried out a field study around the coastal waters of the UK in which European flounder (P. flesus) were sampled from different estuaries. Hepatic DNA analysis of the samples suggested that fish populations in certain contaminated UK estuaries are being exposed to complex mixtures of genotoxic and potentially carcinogenic contaminants. This supports a previous study by the group which demonstrated that European flounder populations inhabiting industrialised UK estuaries are exposed to high levels of sediment-bound PAHs, and that a proportion of the bioavailable PAHs are metabolised to carcinogenic substances [94]. In the present study, these authors have also attempted to correlate the DNA adduct formation with histopathological biomarkers. Although it was difficult to associate the histological lesions with specific sites, the contaminated sites showed higher prevalence of these damages. Given that DNA adduct formation is a repairable damage, the study emphasises the need for complementing chemical data and effects at higher levels of biological organisation (e.g. histopathological biomarkers), when the adduct analysis is carried out. It has been shown that DNA adduct formation is associated with the incidence of hepatic lesions, including neoplasms, in fish at contaminated sites [95]. There is also strong evidence of a cause-and-effect relationship between exposure to genotoxins in sediment and water and neoplasm epizootics in wild fish populations [96].
Frenzilli et al. [97] have attempted to evaluate the levels of DNA strand breaks (using Comet assay) and apoptotic cells in Eelpout (Zoarces viviparous), a non-migratory fish species, sampled from Goteborg harbour area, Sweden, following a oil spill accident. The genetic damage and apoptotic cell analyses were carried out in nucleated erythrocytes. In addition, levels of PAHs exposure were also measured by analysing PAH metabolites in the fish bile. The study suggested a high level of damaged DNA, complemented by a peak in the bile PAH metabolites in samples from the most impacted site, 3 weeks after the oil spill but this damage was found to recover 5 months after the accident. However, the samples collected from the middle part of the harbour, which is chronically impacted from contamination, did not show any recovery. It is interesting to compare this study with a similar study following the Sea Empress oil spill accident in 1996 in Milford Haven, Wales, UK. The study by Harvey et al. [98] indicated that oil contamination did not induce any detectable elevations in DNA adduct levels in the invertebrates (i.e. Halichondria panacea and Mytilus edulis), but the contamination did appear to induce adducts in the fish species (i.e. Lipophrys pholis, P. platessa and Limanda limanda). Interestingly, in common with the study by Frenzilli et al. [97], data obtained 12–17 months after the spill suggested that the affected species recovered from the oil contamination [98]. Bearing in mind that both DNA strand break and adduct formation are repairable damages and, biomarkers of ‘exposure’ rather than ‘effects’, it remains possible that such exposure could lead to mutations in other somatic and germ cells. The selection of robust, multiple end points in phylogenetically different groups of organism should therefore be considered to get a holistic view of potential impact of such accidents when planning a long term studies.
While we are attracted to study contaminated sites to evaluate potential damage in the exposed biota, in order to elucidate and solve many of the challenges in eco-genotoxicology (e.g. identification of physical, chemical and biological factors influencing the induction of genetic damage under natural conditions; identification of sensitive species and life stages; individual variability and genetic susceptibility, etc.), there is also a need for the monitoring sites (or samples) which are considered to be pristine, undisturbed or have moderate levels of contaminations compared with heavily contaminated sites or samples [88]. This is considered to be necessary in order to generate background and historical control data while evaluating the impact of anthropogenic contaminants on natural biota [99]. Monitoring of such sites is also necessary for many estuaries in industrialised areas that require periodic maintenance dredging of navigable waterways to sustain recreational, commercial and military shipping [100] and [101]. Akcha et al. [102] in this issue, investigated the influence of both biotic (e.g. age, sex) and abiotic (e.g. sampling site) factors on the use of genotoxicity biomarkers (i.e. DNA strand breaks and adducts) in different cell types (i.e. erythrocytes and heapatocytes) for environmental monitoring. The authors have highlighted several factors, including seasonal variability, as carried out by Bolognesi et al. and Winter et al. [91] and [92], which require due consideration when conducting field sampling using flatfish species (L. limanda).
Interspecies variability for toxic and genotoxic response is well known and is attributed to the differences in uptake, accumulation, metabolism, excretion, sequestration and repair efficiency. Compared to laboratory studies, the situation in the field becomes more complex due to several confounding factors. It is well accepted that to protect human and ecosystem health, it is necessary to develop sensitive assays and to identify responsive cells and species, and their life stages [86]. Bihari and Fafandel [103], in this issue, have determined the spontaneous and B(a)P induced DNA single strand breaks in a range of aquatic species, and as mentioned earlier, supports the adoption of a multi-species approach in designing biomonitoring programmes. While evaluating the DNA strand breaks, the authors also obtained different elution rates for DNA from different species at different pH, suggesting that DNA denaturation under alkaline conditions are species-specific. This has important implications for standardising protocols for biomonitoring using different species.
For environmental monitoring purposes, one of the major biological challenges has been to bring potential test organisms into continuous routine culture. This is especially important when sound genetic analysis is to be carried out. In this context, the use of black slug (Arion ater), a terrestrial gastropod species, which has been used traditionally as an indicator for metal pollution, has been evaluated by Hamers et al. in this issue, for PAHs monitoring in the environment [104]. The organisms, collected from field were reared in the laboratory and orally exposed to B(a)P for different periods of time (i.e. 3 and 119 days). B(a)P hydroxylase (BPH) and bulky DNA adducts were analysed in the digestive glands and kidney cells respectively. After a short exposure to a relatively high B(a)P doses, a dose-dependent increase for BPH activity and bulky DNA adduct levels was found. The authors however could not find increased BPH activity and bulky adduct levels following long-term exposure of environmentally relevant doses of B(a)P. Based on this lack of response, the authors conclude that this species is not sensitive to B(a)P exposure and, therefore, not suitable for monitoring environmental exposure to PAHs. Given that there is a lack of information pertaining to the use of terrestrial gastropods for environmental monitoring studies, this is valuable information for the workers in the field.
The marine ecosystem is the largest system in terms of size and complexity. Compared to human populations, the aquatic organisms are distributed and restricted to well-defined and distinct occupancy, along the vertical and horizontal habitats of the ecosystem. In these ecological niches, it is hard to find a single species, representative of all. Although the deep sea environment is an important source for food and sink for anthropogenic wastes, it is most poorly studied ecosystem, primarily because of logistic problems to reach these places. In this context, the members of hydrothermal-vent communities in the deep-sea are well-adapted to survive in one of the most naturally contaminated (e.g. heavy metals and radionuclides, etc.) environment on this planet [105]. Given the extreme nature of the vent environment, it is of fundamental interest to study the spontaneous and induced genetic damage in these organisms. Dixon et al., in this issue [106], have reported interesting findings related to adaptation and the effects of hydrostatic pressure change on DNA integrity in blood and gill cells of a hydrothermal-vent mussel species (Bathymodiolus azoricus) collected from the Mid-Atlantic Ridge, an important study area for vent research in European waters. Based on their findings, the authors suggest that future studies should concentrate on populations living at greater depths but only after significant advances are made in logistics of collecting the samples which are subjected to decompression stress during retrieval and subsequent experimentation.
8. Risk assessment and environmental management
In this issue we have seen several examples of how the science is being used to assess the impact of anthropogenic stresses under laboratory and field situations in aquatic species. It is however a challenging task to translate these information to assess the health of the ecosystem as a whole. Determination of the ‘health’ of the ecosystem is therefore rapidly emerging as a new branch of science which focuses on the evaluation of health status, maintenance (i.e. sustainability) of a healthy one and treatment (i.e. restoration) of an unhealthy ecosystem [107]. One of the most popular tools in our efforts to assess the potential impact of anthropogenic stress is ecological risk assessment. Because of the knowledge gaps in our understanding of ecological significance of anthropogenic stressors, there is a lack of consensus for an acceptable, comprehensive decision-making framework which could establish the roles of science and policy in formulating environmental management principles. This has generated several myths that have numerous implications and important consequences for environmental decision making [108]. Some of these myths include selection of sentinel or sensitive species, utility of acute versus chronic data, extrapolations from laboratory to field situations, etc. While science is playing an important role in developing techniques to assess the impact on the environment, at the moment it cannot address all the issues surrounding environmental risk management [108]. Attempts are however being made to distinguish between these myths with the realty and different models are being proposed to evaluate ecotoxicological risk following exposure to genotoxic agents in different environment [109].
In this issue, Moore et al. [110] present their views on this very complex and articulated topic for integrated environmental management. The view, based on the core information from published work from laboratory and field studies, recognises the limitations and strengths of the literature and proposes the use of diagnostic clinical-type, laboratory-based ecotoxicological tests or biomarkers linking different levels of biological organisation, using sentinel animals as integrators of pollution. The information generated could be used by decision makers and environmental managers. The authors also hypothesize that the proposed approach would benefit with the development of computational simulation models of cells, organs and animals in tandem with available empirical data.
9. Future directions
Substantial progress has been made in the last two decades in the field of aquatic genotoxicology. Most of the early developments had involved analysis of chromosomal aberrations in selected organisms, which included fish, bivalve molluscs, polychaete worms and echinoderm larvae [111], [112], [113], [114], [115], [116], [117], [118], [119], [120], [121] and [122]. Most of these studies probably benefited from the techniques developed to study the chromosome number and morphology in 1960s and 1970s for taxonomic and evolutionary studies [15]. However, except for the actively dividing cells of embryo-larval stages of certain marine invertebrates [15], [30], [32], [88], [116] and [117], the successful use of chromosomal aberrations based on metaphase analysis in aquatic organisms has been very limited. This is apparently because of low mitotic rates in tissues (e.g. gill, spleen, etc.) of adult individuals, relatively small-sized and high chromosome numbers in most of the studied aquatic organisms and lack of well-established cell lines for invertebrates [15]. Given that only a few, selected groups of organisms have been explored so far, attempts to get complete metaphase spreads either from established cell lines or from whole organisms is likely to expand in future. This will also facilitate refinement of existing protocols, development of more in vivo test systems in phylogenetically different group of organisms and their successful use in laboratory and field studies. These systems will complement parallel developments in other areas of ecotoxicology and facilitate the adoption of an integrated, inter-disciplinary approach using multiple biomarkers to evaluate the impact of contaminants and other environmental stressors (e.g. UV radiation, hypoxia, nutritional deprivation, parasitic infection, etc.).
With limited success for the use of metaphase spreads for aberration analysis, the micronucleus (MN) assay in blood and gill cells of bivalve molluscs, fish and amphibians was developed and applied under both field and laboratory conditions [123], [124], [125], [126], [127], [128], [129], [130], [131] and [132]. The anaphase-telophase aberration assay was also developed in echinoderm and fish larvae [119], [120], [121] and [122]. The advantages and limitations of MN and anaphase-telophase aberration assays in aquatic organisms have been discussed in the literature [15]. While in mammalian systems, development and applications of molecular techniques (e.g. FISH, M FISH, SKY, COBRA, etc.) are proving very useful to illuminate the minute details of the complexity of the types of aberrations induced [133], the field of aquatic genotoxicology is still far behind in the application of these techniques to elucidate the qualitative and quantitative induction of chromosomal aberrations, one of the most important parameters for its known biological significance. Limited attempts to apply molecular cytogenetic techniques in aquatic organisms have not gone far enough to throw light on the nature and complexity of genetic damage induced by environmental agents [134] and [135]. In future, it is hoped that in parallel with mammalian studies, combination of whole chromosome painting with telomeric and centromeric probes, would provide better insights of induction of structural and numerical chromosomal aberrations in aquatic organisms. This will also provide a robust comparison with well-advanced studies in human and mammalian cells.
The application of single cell gel electrophoresis (i.e. Comet assay) has revolutionised the field of genetic ecotoxicology. This assay has provided the opportunity to study DNA damage, repair and cell death (apoptosis) in different cell types of aquatic organisms, without prior knowledge of karyotype and cell turn over rate. The advantages and disadvantages of this assay have been extensively described in the literature [15], [136], [137] and [138]. In human and mammalian studies, the combination of Comet assay with fluorescence in situ hybridisation (FISH) technique has allowed the quantification of single and double strand breaks, alkali labile sites, and has provided the insights of the role and fate of specific DNA sequences at whole genome or single cell level [139], [140] and [141]. Due to a lack of distinct banding patterns, the packaging of chromatin in lower organisms appears to be different from higher organisms [15]. In mammalian systems, intragenomic heterogeneity for the induction of genetic damage and repair exists [142], [143], [144], [145] and [146]. There is however not much information available pertaining to heterogeneity for DNA damage and repair in aquatic organisms. In fact, very little is known about DNA repair efficiency in different aquatic organisms. Future studies will therefore explore the potential differential sensitivity and elucidation of repair efficiency of apparently homogenous genomes from different cell types of aquatic organisms. This will include use of combination of Comet and FISH techniques [139], [140] and [141], identification of DNA repair deficient (mutant) individuals, construction of mutant cell lines and transgenic models [147]. In common with mammalian systems, these developments will adopt several molecular approaches to dissect the DNA repair mechanisms [141], [148] and [149].
Finally, with growing developments in ‘omics’ (i.e. genomics and proteomics), nanotechnology and biotechnology areas, we will be able to know more about how the organisms perceive changes at the molecular and genetic levels, and make functional responses within their environment. These functional responses will uncover pattern of gene expression in the organisms and their short-and long-term adaptations in response to environmental changes. The mechanistic understanding and adaptations at molecular, genetic and cellular levels could then fill the gap in knowledge to link the reproductive and survival strategies of the organisms under stressed conditions. The tools developed, and information generated, would also help environmental managers to develop and apply robust approach to protect human and ecosystem health. All these developments will make the field interesting, stimulating and challenging.
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