Mutagenesis

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Mutagenesis, Vol. 17, No. 6, 495-507, November 2002
© 2002 UK Environmental Mutagen Society/Oxford University Press

Marine invertebrate eco-genotoxicology: a methodological overview

David R. Dixon², Audrey M. Pruski, Linda R.J. Dixon and Awadhesh N. Jha¹

Southampton Oceanography Centre, Waterfront Campus, European Way, Southampton SO14 3ZH, UK and ¹ School of Biological Sciences, Plymouth Environmental Research Centre, University of Plymouth, Plymouth PL4 8AA, UK

	Abstract

Top Abstract Introduction Applications of genotoxicity... Microlesions Non-genotoxic causes of DNA... Conclusions References

 
The last 25 years have seen major advances in the field of mammalian genotoxicology, particularly with the advent of molecular methods, some of which have spilled over into the relatively new field of eco-genotoxicology, which aims to evaluate the impact of contaminants on the natural biota. Unlike mammalian genotoxicology, where the focus is centred on a limited number of model species, efforts in the marine field have generally lacked coordination and focus, with the result that progress has been somewhat slow and fragmented. However, it is recognized that at the DNA and chromosome levels, marine invertebrates express qualitatively similar types of induced damage to that found in higher organisms (e.g. point mutations, strand breaks and chromosomal aberrations). Given that many of these species (bivalve molluscs, crustaceans, polychaete worms, etc.) are linked directly or indirectly to the human food chain, this is an important reason why one should be concerned about their exposure to environmental mutagens and carcinogens, particularly as many of these organisms have the capacity to (i) transform these agents to biologically active metabolites and (ii) accumulate toxicants in their cells and tissues at concentrations several orders of magnitude above that found in the environment. This review covers the advantages and limitations of those cytogenetic and molecular assays that have been used to address the question of genotoxicity in the cells and early life stages of selected marine invertebrate species. It concludes with the recommendation for the adoption of standardized test procedures, leading to a tiered approach in future eco-genotoxicity testing.

	Introduction

Top Abstract Introduction Applications of genotoxicity... Microlesions Non-genotoxic causes of DNA... Conclusions References

 
Whereas human mutation studies have been supported by a rigorous conceptual framework, stemming from our understanding of the part played by mutagens and carcinogens in genetic disease, birth defects and cancer, we are largely ignorant of the part played by environmental mutagens in affecting the health and long-term survival potential of marine organisms. Hence our concerns in this area are largely based upon an extrapolation from the mammalian model. With the possible exception of those few species already threatened with extinction (great whales and some seals), any additional mortality or abnormality that might result from mutagen exposure in the marine environment is unlikely to have a serious impact at the population/species level apart from on a local scale. It can safely be said that, when viewed against the enormous wastage that takes place naturally amongst the early reproductive stages of many coastal and offshore invertebrates, caused by predation, desiccation, reduced salinity or simply being swept out to sea, any effect resulting from a localized pollution incident is likely to be vanishingly small (Wurgler and Kramers, 1992). However, it has been argued that contaminant exposure could lead to detrimental effects at the population level, such as the loss of genetic diversity, which could have serious implications for species survival and ecosystem functioning (Bickham et al., 2000).

The marine environment provides a sink for many natural and anthropogenically derived chemicals. For example, in 1994 the OECD estimated that ~1500 new chemicals are being added annually to the 100 000 already present in the natural environment (Steinberg et al., 1994). The agricultural, pharmaceutical, oil and gas industries are coming under increasing pressure to implement sustainable environmental management practices and, in particular, to monitor the effects that waste products/discharges may have on the natural environment and habitat destruction (e.g. Natura 2000 sites).

Invertebrates constitute >90% of extant species and play a major role in ecosystem function. In the marine environment, sessile (non-mobile) species are particularly suited to monitoring studies, hence the emphasis on marine mussels (Mytilus spp.) (Bayne, 1976) in national and international programmes (e.g. Mussel Watch) (Goldberg et al., 1978). Bivalve molluscs, including the common blue mussel, Mytilus edulis/Mytilus galloprovincialis, have been widely used for the detection and monitoring of chemical pollution in coastal waters (see for example O'Connor, 1996). One interesting aspect of the use of bivalves in aquatic toxicology is that, compared with other invertebrates, the induction of cancers (both solid tumours and cancers of haematopoietic origin) has been relatively well characterized under both laboratory and field conditions (Gardner et al., 1991; Peters et al., 1994). In one area of the USA where herbicide use had increased, elevated incidences of gonadal tumours were reported in both bivalves and humans (Van Beneden, 1994). These findings suggest that bivalve molluscs represent suitable sentinel or surrogate species for neoplastic effects in man.

	Applications of genotoxicity assays in aquatic invertebrates

Top Abstract Introduction Applications of genotoxicity... Microlesions Non-genotoxic causes of DNA... Conclusions References

 
This section covers those test/screening methods that have been used over the past two to three decades to address the question of genotoxicity in the cells and early life stages of marine invertebrates. For convenience, these approaches can be divided into two categories based on the target size: (i) effects visible at the chromosome level (macrolesions), and (ii) those operating at the biochemical or molecular levels (microlesions) (Brusick, 1978). The first part deals with methods for the detection of macrolesions that require cells to be in a dividing state: structural and numerical chromosomal aberrations, sister chromatid exchanges and micronuclei. The second part deals with methods for the detection of microlesions, which could be applied to cells in interphase. These include the single cell gel electrophoresis (Comet) assay, the gel electrophoresis assay, the alkaline elution and unwinding assays and the analysis of DNA adducts.

Macrolesions
Structural and numerical chromosomal aberrations (CAbs). Of all environmentally induced genotoxic effects, CAbs are the best understood in terms of their consequences for the affected cell or individual. Stemming from the widespread use of chromosome numbers and morphology in taxonomic and evolutionary studies in the late 1960s and 1970s (see for example White, 1977), CAbs have featured prominently in marine invertebrate studies up to the present day (see for example Kadhim, 1990; Jha et al., 1996, 2000).

Metaphase aberrations. While metaphase analysis provides the most comprehensive approach to aberration detection, major limitations in the marine invertebrate field stem from (i) a lack of in vitro cell lines, (ii) inadequate information relating to the karyotype and (iii) a lack by marine invertebrate chromosomes in general to band successfully (Table I). Furthermore, a lack of well-defined meristematic regions and a low cell turnover rate, especially under field conditions (Dixon, 1983; Dixon et al., 2001), limits the use of this approach. However, based on embryo/larval studies, our understanding of the chromosomal characteristics of representatives of two ecologically important invertebrate taxa, the bivalve molluscs and polychaetes, has now reached a stage where it is possible to conduct detailed karyotype analysis with a high degree of accuracy and confidence [e.g. Mytilus spp. (Dixon and Flavell, 1986; Jha et al., 1995a, 2000a,b), Neanthes arenaceodentata (Pesch et al., 1981) and Platynereis dumerilii (Jha et al., 1996, 1997, 2000c; Hagger et al., 2002)].

View this table: [in this window] [in a new window]   Table I. . Differences between mammalian and marine invertebrate cells

 
Two other fundamental differences exist between the cells of marine invertebrates and those of mammals which have seriously restricted the use and development of CAb methods. One is a lack of in vitro cultures and cell lines, which, in part, reflects the failure of mammalian mitogens and growth factors to induce cell division in marine invertebrates, and the other relates to differences in the way that chromatin is packaged: homogeneously dispersed in marine invertebrates and heterogeneously clustered in higher groups. Nevertheless, some workers have had success in maintaining cells in the form of primary cell cultures, e.g. gill cells of Mytilus (Cornet, 1993), but this approach has not developed beyond a preliminary stage. For a review of some problems associated with the culture of marine invertebrate cells see Rinkevich (1999). Similarly, there is little literature relating to chromosome banding (G, C and R banding), but in those species examined to date, differential staining has been restricted to specific chromosomal regions, leaving the bulk of the karyotype undifferentiated (see for example Dixon and MacFadzen, 1987; Martinez-Lage et al., 1994). This lack of banding success has restricted our knowledge of chromosome morphology, which would have been helpful in analysing CAbs more precisely.

Anaphase analysis (AA). The anaphase aberration test was first recommended for use in mammalian genotoxicology three decades ago by the Ad Hoc Committee of the Environmental Mutagen Society and the Institute of Medical Research (1972), but it was another decade before it was first used in an environmental context (see for example Hose et al., 1983; Nichols et al., 1984). With very few exceptions, AA has been used in laboratory studies, which is undoubtedly a reflection of the difficulty in accessing large numbers of dividing cells from field collected material. The only exception is early life stages, but again there are possible pitfalls, particularly in relation to the use of preserved samples.

The first marine use of the AA test was the field study by Longwell and co-workers (see for example Longwell and Hughes, 1980). While not strictly within the purview of this review, this large-scale fish-based study was hampered by the use of formalin, a chemical that is now known to cause preservation artefacts, especially in cells that are undergoing division. It was almost 20 years before this approach was tried again. This time, Hose and Brown (1998) applied AA to the developing eggs of the Pacific herring after the Exxon Valdez oil spill disaster. However, due to the problems inherent with working with uncontrolled field samples, these authors concluded that the method was best suited to laboratory or mesocosm (contained environment) studies using culturable species.

Anaphase aberrations, for the detection of lagging chromosomes (a questionable abnormality), multipolar anaphases, bridges and fragments, has been used over the past two decades in a number of marine invertebrate studies, focusing particularly on sea urchin embryos and larvae (Hose and Puffer, 1983; Hose et al., 1983; Anderson and Wild, 1994), who showed it to be a more sensitive indicator of mutagen exposure (e.g. benzo[a]pyrene) than fertilization or larval development effects, including the highly sensitive 48 h larval bioassay. Recently, we successfully applied AA to the developing embryos of Pomatoceros lamarckii, a tube-dwelling marine worm, whose unspawned gametes were shown to have the potential to serve as a monitoring system/integrator of genotoxin exposure which is suited to both laboratory and field studies (see for example Dixon et al., 1999). In keeping with the observations of Hose and Brown (1998), this work yielded evidence of a link between an increased incidence of AA in two-cell stage embryos and an increased level of abnormality in 48 h old larvae, coupled with decreased survival (Dixon et al., 1999; J.T.Wilson, unpublished PhD thesis).

Aneuploidy and polyploidy. Meiotic and mitotic non-disjunction has been known for a long time to play an important role in the initiation of malignancies and other diseases in human neoplasia (Hook, 1985; Hassold, 1986; Parry, 1998). A significant number of the pharmaceuticals, plasticizers and agrochemicals that regularly find their way into the marine environment are known to be aneugenic, based on the results of standard mammalian assays (reviewed by Parry, 1998). In fact, many agricultural and other pesticides (aimed at insect, fungal and plant pests), many of which end up in the marine environment, were designed primarily to inhibit the process of cell division through effects on the mitotic spindle apparatus. Given this knowledge, it is perhaps surprising that more effort has not been expended in developing test systems for studying the effects of aneugens in marine species. However, regulatory bodies have not pursued this and there is a lack of a coordinated approach by those workers in the field.

Where aneuploidy has been investigated, the approach has been limited to simple chromosome counts [M.edulis (Dixon, 1982), Mercenaria mercenaria (Stiles et al., 1991), various gastropod species (Barsienne and Lovejoy, 2000) and Macoma baltica (Thiriot-Quiévreux and Wolowicz, 2001)]. All these studies reported finding an increased incidence of numerical aberrations linked with contaminated conditions. Furthermore, Leitao et al.(2001) have recently shown that chromosome loss is associated with reduced growth rate in juvenile Pacific oysters. Clearly, these findings point to aneuploidy being an important pollution end-point, which now deserves using high throughput, flow cytometric techniques. However, in keeping with the micronucleus test (see later), such studies must be supported by information relating to cell turnover rate, since a failure to detect any effect may simply be a reflection of a failure of the cells to divide. Cell division rate is known to be a sensitive indicator of environmental stress (see for example Stebbing, 1981).

In the mammalian literature there has been some debate about the importance of polyploidy as a genotoxic effect (Mitchell et al., 1995). The only marine invertebrate work is that of Dixon (1982) and Wilson et al. (in press). In the first case it was shown that mussel embryos (M.edulis) originating from a polluted dock exhibited a higher frequency of aneuploidy but not of polyploidy compared with animals collected from a pristine environment. In the case of a laboratory exposure study using the embryos of the polychaete P.lamarckii, polyploidy was only observed (at 0.5% frequency) after acute exposure to di-butylphthalate at a concentration of 1x10^-5 M, but not at lower doses or in the controls (Wilson et al., in press). Clearly, based on this limited amount of work, polyploidy does not appear to be a sensitive genotoxicity end-point in marine invertebrates.

Application of the FISH technique. Given the ability to detect a wide range of structural and numerical abnormalities in both metaphase and interphase preparations, it is somewhat surprising that more work has not been done using fluorescence in situ hybridization (FISH), especially given the large number of invertebrate sequences that are available in the sequence databases (GenBank, etc.). A lack of probes for individual chromosomes for marine invertebrates has limited the application of this technique for chromosomal aberration analysis. With the exception of one study using a vertebrate telomere probe (Jha et al., 1995b), the remaining marine invertebrate studies have concentrated on the use of a ribosomal DNA probe [dogwhelk (Pascoe et al., 1995), M.galloprovincialis (Martinez-Exposito et al., 1997), amphipod (Libertini et al., 2000) and mollusc and polychaete (Viturri et al., 2000a,b)], which was intended to overcome inconsistencies associated with the use of silver staining and other chemical methods when visualizing NORs. Recently, we used the FISH method to study rates of aneuploidy in interphase nuclei derived from the embryos of the polychaete P.lamarckii (Wilson et al., in press). Two problems that became apparent during this study were: (i) the difficulty of distinguishing between true aneuploidy and replicating chromosomes (a problem when dealing with asynchronous cell populations); (ii) the tendency for some FISH signals to be obscured when overlain by embryo cytoplasm, something that has not been reported in mammalian studies. Regardless of these shortcomings, an increase in aneuploidy frequency was observed following exposure to a range of recognized and suspected aneugens (e.g. colchicine and TBZ), which demonstrated the utility of the FISH method in aneuploidy scoring. This and other new developments originating from the field of molecular cytogenetics (e.g. chromosome painting) are now available in an eco-genotoxicological context.

Micronucleus (Mn) assay. The Mn assay has featured significantly in marine invertebrate studies and has been applied mainly to different species of bivalve exposed under both field and laboratory conditions (Majone et al., 1987; Brunetti et al., 1988, 1992a,b; Scarpato et al., 1990; Wrisberg and Rhemrev, 1992; Wrisberg and van der Gaag, 1992; Burgeot et al., 1995, 1996; Mersch et al., 1996; Venier et al., 1997; Bolognesi et al., 1999; Hagger et al., 2000). A limited number of studies have also been performed using echinoderm larvae (Vacquier and Brachet, 1969; Hose et al., 1983; Pagano et al., 1983; Saotome et al., 1999).

Gill and blood cells (haemocytes) of bivalve molluscs have been used widely for Mn analysis (Figure 1). The use of haemocytes has several advantages: (i) these cells form part of the open vascular system and are therefore likely to be homogeneously exposed to toxicants; (ii) laboratory studies have suggested that these cells are sensitive to reference genotoxicants; (iii) the cells can be obtained irrespective of the reproductive phase of the animals (e.g. in contrast to embryo/larval stages); (iv) the assay could be complemented by other cytotoxicity (e.g. neutral red retention assay) (Bolognesi et al., 1999; Lowe and Fossato, 2000) and genotoxicity assays (e.g. DNA strand break measurement) (Steinert et al., 1998; Bolognesi et al., 1999); (v) the assay is applicable to both laboratory and field conditions, including cage transplant studies.

View larger version (43K): [in this window] [in a new window]   Fig. 1. . Micronucleus assay. A micronucleated agranular haemocyte from M.edulis.

 
Despite the potential advantages of haemocytes from bivalve mussels for the Mn assay, it appears that the system has not been properly validated under laboratory conditions prior to its use in field studies. In particular, little attention has been paid to the influence of cell turnover kinetics on Mn expression, given that the induction of cytogenetic damage is a cell cycle-dependent phenomenon. Furthermore, studies enumerating Mn formation have adopted arbitrary sampling times (see for example Wrisberg and Rhemrev, 1992; Mersch et al., 1996), while there is also evidence for high levels of inter-individual variability, as reported, for example, in gill cells (Venier et al., 1997). In field studies too, mussels have tended to be exposed for variable time periods before sampling, making comparison and interpretation difficult. Moreover, most investigators have pooled the results from several individuals following exposure to the chemicals or (uncharacterized) environmental samples, without taking into account the potential for inter-individual variability.

One of the limitations of studies carried out using haemocytes of mussels is that most authors have not focused on a specific cell type. This is mainly because of a lack of detailed knowledge of haemocyte sub-populations and their origin in invertebrates. In bivalve molluscs, haemocytes can be classified morphologically into agranular and granular cells (Cheng, 1981). The haemocytes of M.edulis, for example, can also be divided into basophilic and eosinophilic groups, the former including hyaline (agranular) and granular cells, the latter only granular cells. The granular haemocytes of M.edulis contain two distinct types of granules, which can be distinguished by size, lectin staining characteristics and enzyme content (Dyrynda et al., 1997). Despite these haematological complexities, some studies have attempted to restrict counts to the non-granular cell population while quantifying the induction of Mn (Venier et al., 1997). These authors, while evaluating the effects of benzo[a]pyrene, found granular haemocytes less sensitive compared with agranular haemocytes. In this study, binucleated cells were detected along with other nuclear abnormalities, the levels being significantly higher in exposed individuals compared with controls. In addition, it has also been suggested that the inducibility of Mn in bivalve haemocytes appears to be limited in scale, since Mn levels have not been observed to be more than three times greater than controls, suggesting a `plateau' or `saturation' effect (Wrisberg and Rhemrev, 1992).

The use of bivalve gill cells for Mn analysis also has several limitations. Firstly, compared with haemocytes, gill cells can only be obtained once from an individual since their removal usually involves a degree of damage. Second, solid tissue preparations require enzymatic treatments (e.g. protease or dispase) plus washes and incubation in different buffers to obtain single cell suspensions (Venier et al., 1997; Bolognesi et al., 1999), which makes the method more difficult to perform. Furthermore, the cell suspensions obtained from gill filaments are heterogeneous in composition (like haemocytes), which may influence their sensitivity and introduce bias into the scoring procedure. At least two cell types have been identified in mussel gills. A predominant, larger cell type, with large, round and well-spread nuclear chromatin, which is distinct from a smaller cell type which is characterized by having compact nuclei and a higher nucleus:cytoplasm ratio (Scarpato et al., 1990; Venier et al., 1997). In addition, Venier et al.(1997) reported the presence of some large cells with abundant cytoplasmic granules and with a peripheral nucleus similar to that of the granular haemocytes, which occurred in different proportions in different individuals. It was suggested that these were haemocytes, since mussels have an open circulatory system and haemocytes are able to move into the connective tissues (see for example Dopp et al., 1996). For Mn scoring it is important, therefore, that only large, agranular spherical cells with prominent nuclei are analysed from gill cell preparations. In view of these cellular characteristics, it is recommended that given a shorter time involved in slide preparation, ease of cell identification, lower baseline frequency and higher induction factor (Wrisberg and Rhemrev, 1992; Mersch et al., 1996; Bolognesi et al., 1999), haemocytes are the most suitable cell type for Mn studies.

Studies have also compared the Mn induction in gill and blood cells of mussels and it appears that gill cells tend to have higher baseline and induced frequencies (Mersch et al., 1996; Bolognesi et al., 1999). This can be explained by the fact that the gill epithelium represents the primary target for contaminants. When comparing the induction of Mn in different cell types of the same individual, one must take into account the exposure period and cell turnover time. Lack of information on cell cycle kinetics limits the value of a number of past studies and limits our understanding of the induction of cytogenetic damage in bivalve gill cells and haemocytes. Clearly, there is a need for basic research to elucidate these cytological aspects in marine invertebrates.

Background or spontaneous levels of Mn can vary depending upon species, season and staining procedures (see also SCGE discussion). Brunetti et al. (1988, 1992a,b) have suggested a significant influence of physiological (age, growth and seasonal changes) and physical factors (temperature and oxygen factors) on Mn frequency in mussel gill cells. When comparing results from different studies, it is important that all these factors be taken into account. It has been suggested that the variation in response with season is probably due to temporal variations affecting the mixed function oxygenase (MFO) levels in the tissues (Wrisberg and Rhemrev, 1992). This is something that should be kept in mind when using this and other assays under field conditions at different times of the year. Given that physiological condition varies with reproductive status and reproductive timing varies across the country, due to water temperature differences (Seed, 1976), it is likely that even when samples are collected in the same season from different areas of the country, both the baseline and induced levels of genetic damage may differ quantitatively.

In view of the limitations of the earlier reports using haemocytes from mussels, it is essential that: (i) the system is properly validated against a range of concentrations of reference genotoxic agents relevant to the marine environment (to ensure sensitivity, reproducibility and reliability); (ii) the kinetics of induction of micronuclei over a period of time with different exposure scenarios are analysed; (iii) the inter-individual variability, if any, is evaluated by systematic and repeated extraction of haemolymph from the same individuals and then the system is adapted to the field situation (contaminated and uncontaminated sites); (iv) the relative sensitivity of the biological assay is compared with chemical analysis for the detection of genotoxicity. We have adopted such an approach when using the blood cells of mussels (Fig. 1; A.Jha, A.Bowen, A.Bloomfield and J.Hagger, in preparation). While several technological advances have been made to clearly and unambiguously identify the induction of Mn in mammalian cells under in vitro conditions (well preserved cytoplasm, Mn not touching the main nucleus, similar or weaker staining and size of Mn 1/3 of the main nucleus), limitations of this technique under in vivo conditions due to lack of basic biological information in aquatic organisms has first to be overcome while gradually improving the technique to improve its sensitivity and therefore usefulness in aquatic systems.

As mentioned earlier, a few laboratories have attempted to develop the Mn assay using embryo/larval stages from echinoderms (mainly sea urchins). Sea urchin bioassays have been widely used as a sensitive indicator of toxicity in marine studies (Kobayashi, 1971). Sea urchins, like bivalve molluscs, are cosmopolitan in their distribution and by selecting a range of species a regular supply of gametes can be obtained for laboratory testing purposes. It is no surprise, therefore, that this source of dividing cells has been exploited for Mn studies (Pagano et al., 1982; Hose, 1985; Anderson and Wild, 1994; Saotome et al., 1999). In common with other embryo/larval assays, the advantage of this system is that it offers the opportunity to study other key events in embryogenesis, such as fertilization, embryogenesis and larval development. These reproductive end-points can be studied in parallel to determine contaminant effects on the reproductive success of the organism, one of the key objectives of ecotoxicology. Consequently, such short-term bioassays offer the opportunity to screen genotoxicity, teratogenicity and embryotoxicity together, in addition to any induction of Mn. This type of assay has also been used to study anaphase aberrations (Hose, 1985; Anderson and Wild, 1994). Problems encountered with the use of echinoderm assays are: (i) preparation of monolayer cells on slides from 48 h old embryos/larvae for the evaluation of the cytogenetic damage (i.e. Mn or anaphase aberrations); (ii) exposure period (bearing in mind that cytogenetic damage is a cell cycle-dependent phenomenon); (iii) lack of information pertaining to metabolizing capacity for xenobiotic substances at these early stages of development. The method used to make slides from embryonic cells has been criticized on the grounds that it may introduce artefacts (Kligerman, 1980). In addition, the sensitivity of the assay may be compromised when embryos are exposed for long periods to toxicants which may (i) result in the dilution of the induced damage and (ii) increase the risk of pathological effects.

Sister chromatid exchanges (SCE). Unlike chromosomal aberrations and micronuclei, which are indicators of effect, SCE, with their uncertain aetiology and significance, are considered to be a biomarker of exposure. The SCE technique has been demonstrated to be a sensitive exposure end-point when applied to a range of marine invertebrates and fish species (see for example Zakour et al., 1984; Dixon and Prosser, 1986; Jones and Harrison, 1987; Jha et al., 1996). However, SCE remain largely a sensitive adjunct to other cytogenetic methods. At a practical level, exposing cells in vivo to bromodeoxyuridine for two cell cycles presents a serious problem, particularly under field conditions. However, Dixon and Pascoe (1994) showed that it was possible to measure SCE in early stage mussel (M.edulis) embryos originating from adults collected from a range of contaminated field sites. However, constraints imposed by seasonal reproduction and the failure to obtain intact metaphases from these cells severely limited the utility of the method.

While the range of species employed in SCE studies has been limited, both life stage and cell type differences in response have been reported. Early life history stages have featured prominently in marine invertebrate studies, but where the adult stages have been investigated these were found to have different sensitivities and SCE response characteristics to those of larvae (Dixon and Clarke, 1982; Jones and Harrison, 1987). For example, in 24 h old Mytilus larvae, rather than eliciting a generalized increase in response, mutagen exposure prompted an increase in SCE frequency in a subset of cells only, so-called high frequency cells (Harrison and Jones, 1982). Furthermore, in some species/life stages the SCE response is limited to a simple doubling of the baseline rate (see for example Pesch et al., 1981; Harrison and Jones, 1982). Table I summarizes the differences between marine invertebrate and mammalian cells which have limited their application in cytogenetic studies in general.

	Microlesions

Top Abstract Introduction Applications of genotoxicity... Microlesions Non-genotoxic causes of DNA... Conclusions References

 
Single cell gel electrophoresis (SCGE) or Comet assay
Laboratory and field studies. SCGE is currently the most widely employed method to detect microlesions in marine invertebrate genotoxicology. Most of this work has focused on the marine mussel M.edulis (see for example Mitchelmore et al., 1998a,b; Steinert et al., 1998a,b; Wilson et al., 1998; Pruski and Dixon, 2002), apart from a few studies that have featured other bivalve species, Crassostrea virginica (Nacci et al., 1996), Patunopecten yessoensis and Tapes japonica (Sasaki et al., 1997) and the shrimp Palaemonetes pugio (Lee et al., 2000). Compared with other methods, SCGE has been shown to be highly sensitive, with a sensitivity similar to that observed in mammals (Table II). Despite the small number of field studies that have been carried out to date, it has been clearly demonstrated that SCGE can be used to detect a reduction in water quality caused by chemical pollution (Steinert et al., 1998a; Frenzilli et al., 2001), as well as synergistic interactions involving sunlight and pollutants (Steinert et al., 1998b). However, significant levels of inter-individual variability, including seasonally related variations in response, have been reported from some studies, both laboratory and field (Nacci et al., 1996; Wilson et al., 1998; Shaw et al., 2000; Frenzilli et al., 2001). This underlines the need for care when designing experiments and selecting controls (for this and other methods) and emphasizes a requirement for more baseline monitoring to be carried out over long time intervals. In a recent study, mussels from some polluted sites were found not to show significantly greater DNA strand breakage levels than individuals from reference sites (Shaw et al., 2000; Large et al., 2002), which was interpreted (based on transplant studies) as an adaptive response to prevent the detrimental effects (including energetic costs) of ongoing DNA damage in populations that were being chronically exposed to pollution (Shaw et al., 2000; Large et al., 2002). Clearly, if this behaviour were confirmed and shown to be a widespread phenomenon, this would pose a major problem when applying SCGE to field populations.

View this table: [in this window] [in a new window]   Table II. . Summary of SCGE results obtained using cells from a marine bivalve and two mammalian species

 
SCGE has also been applied in vitro. Dose–response increases in DNA strand breakage were recorded in cells of M.edulis exposed to both direct-acting (hydrogen peroxide and 3-chloro-4-(dichloromethyl)-5-hydroxy-2[5H]-furanone) and in-direct acting (benzo[a]pyrene, 1-nitropyrene, nitrofurantoin and N-nitrosodimethylamine) genotoxins (Mitchelmore et al., 1998b; Wilson et al., 1998). Recently, work in our laboratory has shown that the Comet assay can be used to visualize DNA damage in spermatozoa of a tube-dwelling polychaete P.lamarckii that had been exposed to a range of different contaminants (J.T.Wilson, unpublished PhD thesis). As spermatozoa lack the ability to repair DNA damage, this application of the Comet assay has considerable potential as a sensitive indicator for use in future laboratory and field studies.

SCGE in studies of DNA repair. In mammals, the SCGE assay has been used extensively in DNA repair studies (Speit and Hartmann, 1999; Collins et al., 2001). In a recent study of the effects of cadmium on the DNA integrity and repair efficiency of gill cells of M.edulis, we showed that the method provides a useful means of following the kinetics of DNA repair (Pruski and Dixon, 2002). While cadmium was not shown in itself to be directly genotoxic, a marked decrease in DNA repair capability was found in mussel cells that had been pre-exposed to cadmium. As a number of environmental contaminants (arsenic, cadmium and mercury) have now been shown to have the ability to interfere with DNA repair (Hartwig, 1998; Kitchin, 2001), the comet-based approach has potential when investigating indirect effects, such as occur when dealing with complex mixtures. Furthermore, since DNA strand breakage can have a diverse and complex aetiology, changing the focus away from mutagenicity per se towards the kinetics of DNA repair offers an attractive alternative, which represents both a useful marker of exposure and a less ambiguous end-point than strand breakage. However, variations in DNA repair efficiency do occur naturally, as a result of seasonal variations in the levels of detoxification enzymes, reproductive status and as part of the overall ageing process (see below), which emphasizes the need to exercise great care when interpreting results. To illustrate this point, Figure 2 shows SCGE results from a study in which we investigated the efficiency of DNA repair in large and small M.edulis following exposure to hydrogen peroxide (10 µM). By comparing strand breakage levels at time intervals after exposure with and without treatment with the DNA repair inhibitors aphidicoline and cytosine-ß-D-arabinofuranoside (ARA-C), it became apparent that while young mussels showed an ability to repair at least some of the damage they sustained from the pro-oxidant exposure, no similar evidence of DNA repair was recorded in the case of the older size class. Clearly, age-related differences in response and repair potential need to be taken into account when designing experiments, particularly in relation to field studies.

View larger version (13K): [in this window] [in a new window]   Fig. 2. . The SCGE assay. Effect of size/age on cell recovery following oxidative stress in M.edulis. Suspensions of gill cells were prepared from (A) small (2.7 ± 0.2 cm length shell) and (B) large (4.5 ± 0.3 cm length shell) animals collected from Hill Head (Hampshire, UK), a nominally clean site, and treated with hydrogen peroxide (10 µM) for 1 h with or without DNA repair inhibitors (10 µM aphidicoline and 10 µM ARA-C, in combination). Cells were then embedded in low melting point agarose and the slides were either subjected immediately to the SCGE assay or left to recover in CMFS for 6 h in the presence or absence of the inhibitors prior to electrophoresis. Experiments were conduced at 4°C. Results are expressed as mean + SEM, n = 5. White bars, without inhibitors; black bars, with inhibitors. **, Statistically different from negative control (i.e. without inhibitors), Student's t-test, P < 0.001.

 
Critical evaluation of the Comet assay. Mitchelmore and Chipman (1998) recently reviewed the value of SCGE for laboratory studies and environmental monitoring. The principal advantages and disadvantages associated with the method are shown in Table III. Compared with those other methods referred to above, SCGE provides a well-researched and widely utilized tool for addressing various aspects relating to genotoxicity in the marine environment. Its main advantages are that it is sensitive, rapid and therefore, a relatively inexpensive method, which can be applied both in vivo and in vitro to a wide range of cell types (namely gills, digestive gland, haemocytes, germinal tissues, embryos and larvae) and, importantly, without any requirement for cell division (see also a review by Fairbairn et al., 1995).

View this table: [in this window] [in a new window]   Table III. . Advantages and disadvantages connected with the Comet assay

 
While guidelines relating to the use of the Comet assay have been published for mammalian genotoxicology (Tice et al., 2000), no standardized protocol currently exists for marine invertebrate studies. As pointed out above, variations in protocol can lead to major differences in results. The composition of the lysis buffer and length of unwinding and electrophoresis, as well as the running conditions, are all important variables that can influence the final outcome. Another critical element is the choice of dissociation procedure. A loss of DNA integrity has been reported for cells that have been treated with trypsin (Birmelin et al., 1998), which is commonly used in the enzymatic dissociation protocols when dealing with solid tissues. Likewise, when using a non-enzymatic mechanical method to isolate mussel gill cells, we observed increased amounts of DNA damage after longer dissociation times (Pruski and Dixon, submitted for publication).

In addition to those DNA lesions inflicted by mutagens, strand breaks also arise as a consequence of DNA repair (transient strand breaks) and apoptosis (see for example Speit and Hartmann, 1995; Steinert, 1996; Choucroun et al., 2001). Apoptosis has also been identified as a complicating factor in relation to the Mn test (Meintières et al., 2001). DNA repair can be prevented by the addition of inhibitors (Mitchelmore and Chipman, 1998). We found using the annexin V affinity assay that >20% of mussel gill cells underwent apoptosis when exposed to 100 µM H₂O₂ (Pruski and Dixon, 2002). This effect needs to be taken into account when interpreting the mutagen responses of marine invertebrates, particularly since apoptotic cells are indistinguishable from normal cells when subjected to the usual dye exclusion tests which are used to determine cell viability. However, apoptotic cells do give rise to comets that have a distinctive rounded `hedgehog' shape (Hartmann, 2000), thus providing a useful indicator of any heterogeneity in cellular response (Figure 3). An ability to discriminate between apoptotic and non-apoptotic cells is a distinct advantage when working with the cells of marine invertebrates, because, in contrast to what has been observed for mammalian cell cultures, there is high inter-cell variability, which may be explained by the presence of several cell types (a particular problem when dealing with differentiated tissues) or other properties of the cell, including the cell cycle status (e.g. cells that are in S phase contain replication forks that behave as single-strand breaks and thus increase the background level of DNA damage) (Olive, 1999).

View larger version (24K): [in this window] [in a new window]   Fig. 3. . The SCGE assay. Appearance of gill cells from M.edulis processed in the SCGE assay. (A) Undamaged cell (most of the DNA is located in the head of the comet); (B) damaged cell (the DNA present in the tail reflects the DNA damage as single- and double-strand breaks and alkali-labile sites); (C) apoptotic cell (the nucleoid core is not distinguishable).

 
While cryopreserved cells have been used successfully in mammalian studies (see for example Visvardis et al., 1997), the high salt content of marine invertebrate cells (equivalent to 3.5% salt) leads to extensive cellular damage during freezing (D.R.Dixon, unpublished results). Similarly, transportation of live animals can itself lead to increased levels of DNA damage, resulting from stress effects brought about by, for example, lysosomal destabilization. Given these practical problems, ideally both the agarose gel embedding and electrophoresis steps need to be done at the time of animal collection, which introduces complexity into the field application of the method. However, once made, slides can be stored in lysis buffer for several months without any apparent increase in DNA damage (Steinert et al., 1998a). After electrophoresis, slides can also be dehydrated in absolute alcohol before air drying and long-term storage (Pruski and Dixon, submitted for publication).

By altering the buffer pH it is possible to distinguish double-strand breaks (using the so-called neutral Comet assay at pH 8.3) from single-strand breaks (alkaline version at pH > 13). As alkali-labile sites (such as apurinic/apyrimidinic sites) lead to DNA single-strand breaks at pH 12.5 (Fortini et al., 1996), the standard alkaline version of the assay does not allow discrimination between naturally occurring alkali-labile sites and environmentally induced strand breaks. However, using a further methodological refinement, Horváthová et al.(1999) showed that it is possible to discriminate between natural alkali-labile sites (pH 13) and single-strand breaks (pH 12.1). The ability to discriminate between these three different types of strand break enormously extends both the sensitivity and utility of the method.

Other methods for the detection of DNA strand breakage. The alkaline elution technique has been applied in a number of genotoxicity studies based on the Mediterranean mussel M.galloprovincialis using both haemocytes (Bihari et al., 1992; Vukmirovic et al., 1994) and gill cells (Bolognesi et al., 1992). These studies demonstrated the applicability of the alkaline elution technique for detecting DNA damage caused by in vivo exposure to different test compounds. Similarly, the alkaline unwinding assay has been reported as a rapid and sensitive method for measuring the induction of DNA strand breakage in the cells of marine invertebrates, under both laboratory and field conditions. For example, Nacci and Nelson (1992) used this assay to detect DNA damage in the gill tissues of M.edulis and Crassostrea virginica that had been transplanted into an urban estuary setting contaminated with polychlorinated biphenols, polyaromatic hydrocarbons and heavy metals. Significantly, their findings were supported by both in vivo and in vitro experiments using reference mutagens. Other uses of the alkaline unwinding assay include investigations on M.galloprovincialis (Accomando et al., 1991; Viarengo et al., 1991) and the starfish Asterias rubens (Everaarts, 1995).

The gel electrophoresis assay. Up to now we have dealt with established assays based on the mammalian toxicology model, which have been shown to have their limitations, particularly when used under field conditions. A lesser known assay, based on agarose gel electrophoresis, the GE assay, which has only been used to a limited extent in the mammalian field, has been found by us to have considerable potential when screening for DNA damage in the tissues of marine invertebrates, particularly under field conditions (Figure 4). Unlike SCGE, the GE assay has the advantage of being easy to use since there is no requirement for extended processing prior to the preservation of the samples. However, the GE assay shares much in common with the Comet assay in terms of principle and quantifying the effect, namely the measurement of the distance migrated and the dispersion of broken DNA strands in an electric field, but in the case of GE this applies to a pooled sample of cells and not a single cell isolate. The best tissue preservative we have found is BLB buffer (5% SDS, 250 mM EDTA, pH 8, and 50 mM Tris, pH 8), which, unlike some other preservatives in common use (e.g. alcohol), does not cause in itself a high level of strand breakage.

View larger version (146K): [in this window] [in a new window]   Fig. 4. . The GE assay. Genomic DNA samples from Atlantic vent mussels B.azoricus run on a 1% neutral gel to reveal differing levels of DNA damage after different treatments and recovery types. Lanes A and B, molecular weight marker; lane C, untreated Mytilus reference; lane D, Lucky Strike vent field [remote operated vehicle (ROV) collected, 1750 m water depth]; lane E, Lucky Strike, ROV collected and exposed to MMC; lane F, Menez Gwen vent field, 820 m, ROV collected; G and H, Menez Gwen, acoustically retrievable cage collected after 12 h and 1 week in the laboratory at 1 bar pressure. It is evident that the best quality DNA, equivalent to untreated Mytilus came from the cage collections and showed evidence of a marked improvement with time spent in the laboratory.

 
In principle, the GE assay is extremely simple and consists of five steps: collection/exposure, preservation, DNA extraction, electrophoresis and quantification. The GE assay has the important advantage that it can be applied to virtually any cell type or tissue. Agarose gels between 0.1 and 3% have the ability to separate DNA fragments over four orders of magnitude on the molecular size scale (Sambrook et al., 1989), but in our studies a 1% gel was found to provide a satisfactory compromise, having the power to resolve DNA fragments in the size range 0.5–23.1 kb; DNA >23.1 kb is effectively intact genomic DNA. The GE assay has featured several times in the mammalian literature (Martin and Black, 1998; Frenzilli et al., 2001) with various methodological refinements (see for example Drouin et al., 1996). In our experience, using hydrothermal vent organisms, this method gave useful results for moderate and high doses of oxidizing agents, but its sensitivity was less than that found for SCGE at low doses (i.e. in the 10 µM range).

In a toxicological context, the vast majority of DNA lesions, and single-strand breaks in particular, are neutral in effect, as most are quickly repaired by direct rejoining or base excision repair. In contrast, double-strand breaks are much more serious in effect and form the basis for many structural chromosomal aberrations and molecular (i.e. genic regulation) uncoupling events that can have serious consequences for the cell and, potentially, the carrier organism as a result of cell death, damage or even neoplastic transformation. These potentially more important double-strand breaks can be visualized on gels of neutral pH. In the case of marine bivalves and polychaetes, where there is a naturally high incidence of ALS (alkali-labile sites), this type of gel has proved more informative than the more widely used alkaline version (Dixon et al., in press).

DNA adducts. The DNA adduct approach to environmental monitoring was first validated using fish species (see for example Stein et al., 1990). Levels of adducts in English sole from Puget Sound, WA, were compared with levels of high molecular weight polyaromatic hydrocarbons in sediments, which showed a positive correlation. There have been a number of studies using the ³²P-post-labelling technique for the detection of DNA adducts in M.edulis and M.galloprovincialis, a close relative (see for example Sole et al., 1996; Venier and Canova, 1996; Harvey et al., 1997; Canova et al., 1998; Harvey and Parry, 1998). These studies demonstrated both the sensitivity of the assay and considered the possible biological implications of the effect. However, a causal link between the presence of the DNA adducts and the initiation of tumours in these organisms remains to be demonstrated, although DNA adduct formation has been found to be associated with an increased incidence of hepatic lesions, including neoplasms, in fish at contaminated sites (Reichert et al., 1990). Interestingly, while DNA adducts were recorded from intertidal fish, no similar response was observed in invertebrates following the Sea Empress oil spill in February 1996 in south Wales, UK (Lyons et al., 1997; Harvey et al., 1999). As with SCE, the presence of DNA adducts in these organisms must therefore be considered at the present time to be strictly a biomarker of exposure. As with other methods, it is important to validate the methodology in the laboratory for any particular species, taking into account age, life stage, sex and cell type differences prior to any application in the field.

	Non-genotoxic causes of DNA damage

Top Abstract Introduction Applications of genotoxicity... Microlesions Non-genotoxic causes of DNA... Conclusions References

 
As with other organisms, DNA damage may arise as a result of processes unrelated to mutagen exposure and clastogenicity. For example, both physical and chemical stressors (e.g. high temperature and increased salinity) are known to cause DNA damage through processes linked to lysosomal enzyme destabilization as part of a general stress response. In such cases, evidence of damage is usually reversible and is lost once an organism is returned to its normal conditions. In Figure 4 such an effect can be seen in DNA derived from the deep sea vent mussel Bathymodiolus azoricus following two different recovery methods. However, one method, namely that using a remote operated vehicle (ROV), was found to be more severe. It was only after animals had been given time (>24 h) to recover from this collection stress that differences relating to their specific environmental conditions could be detected (Pruski and Dixon, submitted for publication).

	Conclusions

Top Abstract Introduction Applications of genotoxicity... Microlesions Non-genotoxic causes of DNA... Conclusions References

 
The basic underlying principle of these tests is whether a chemical (or its metabolites) induces damage to DNA. Driven by the aim of protecting human health, in conjunction with regulatory developments, a spectrum of tests has been developed (>200 to date) (Houk, 1992), of which a selection are being rationalized with attempts to harmonize them at the international level. While some attempts are being made to protect the environment from increased radiation exposure (Pentreath, 1998), due to the diversity of the species involved little attempt has been made to make quantitative estimates of the possible consequences of exposure to anthropogenically derived mutagens to the natural biota, either on land or in the aquatic environment. An attitude that prevails is that if human beings are being protected, the environment will be also. In addition, policy makers require a proven cause and effect before taking any action, rather than adopting precautionary measures. In recent years, however, there have been some regulatory developments which have recommended the genotoxicity testing of chemicals and effluents prior to their discharge into the marine environment. For example, a requirement for evaluating the mutagenic, carcinogenic and teratogenic properties of substances entering the aquatic environment was recommended by the 1995 North Sea Conference (OSPAR Commission, www.ospar.org). Similarly, the US Environmental Protection Agency has recommended adding mutagenicity testing protocols to the current suite of chemical methods used to detect the presence of specific chemicals or effluents (Houk, 1992). The International Standards Organisation (ISO) is now recommending a standard bacterial method (i.e. the UmuC test) for testing the genotoxicity of wastewater. In this context, it is worth mentioning that while in vitro or bacterial test systems could be used as screening tools to define the `intrinsic' genotoxicity of a substance, definitive ecotoxicological risk assessments should also consider the `expressed' genotoxic activity in ecologically relevant organisms. This approach will thus take into account environmentally realistic routes of exposure, the effects of metabolism and DNA repair efficiency (Jha et al., 2000a).

In mammalian genotoxicology a tiered screening approach has been employed and universally accepted by the regulatory bodies. The core strategy involves using in vitro assays which are designed to be sensitive and to detect intrinsic genotoxicity. If clear evidence of genotoxicity is seen in one or more of these assays an assessment is made in vivo in order to determine whether this intrinsic genotoxicity is expressed at the level of the whole animal (Elliott, 1994). In relation to eco-genotoxicology, a clear distinction is needed regarding testing, in which case many of the standard tests already employed for human hazard assessment can be used, and biomonitoring, in which case some of the methods described here have an important role. Certain standard bacterial tests (Ames test, UmuC test, Mutatox^®, SOS chromotest, etc.) and in vitro assays (based on fish cell lines) are already available to evaluate intrinsic genotoxic potential, while the SCGE, CAbs and Mn assays, as applied to marine invertebrates, have now reached a stage where they could be used to detect expressed genotoxicity and for environmental monitoring purposes. The next 25 years promises to be an exciting and challenging period for eco-genotoxicology, which will also see developments in the areas of environmental genomics and proteomics.

	Acknowledgments

 
We are grateful to Jim T.Wilson (Southampton Oceanography Centre) for his help in various ways during the assembly of this review. This review contains results obtained during the EU-VENTOX project (EVK3CT1999-00003).

	Notes

 
² To whom correspondence should be addressed. Email: d.dixon{at}soc.soton.ac.uk

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