7 |
Methods to Assess Effects on Brackish, Estuarine, and Near-Coastal Water Organisms |
| M. H. Depledge | |
| Odense University, Denmark | |
| S. P. Hopkin | |
| University of Reading, United Kingdom |
In 1956, 2000 cases of alkyl-mercury poisoning resulting from ingestion of contaminated fish and shellfish were reported among fishermen and their families at Minimata Bay, Japan (Clark, 1986). This and other similar poisoning events stimulated marine pollution studies throughout the 1960s and 1970s. Research focused primarily on detecting the dangers posed to human health from contaminants in marine food products.
During the past decade, at least two major changes in emphasis have occurred. First, marine ecosystems are worth preserving in their own right and not simply to limit indirect threats to human health. Second, the science of ecotoxicology has revealed that the effects of pollutants in natural ecosystems are diverse, complex, and often unpredictable. Dissatisfaction is increasing with the lack of ecological relevance of many standard ecotoxicological tests and a realisation that extrapolation from laboratory findings to real-world situations is often impractical (Giddings, 1986). This situation has spurred efforts to find more ecologically relevant methods to assess the effects of pollutants in marine, estuarine and brackish waters (Hopkin. 1993). An objective of this chapter is to review these methods.
A key problem in ecotoxicology is that many of the terms used are ambiguous. Therefore, the terminology used here will be defined.
The Group of Experts on the Scientific Aspects of Marine Pollution (GESAMP) has defined marine pollution as "the introduction by man, directly or indirectly, of substances or energy into the marine environment (including estuaries) that result in such deleterious effects as harm to living resources, hindrance to marine activities including fishing, impairment of quality for the use of seawater and reduction of amenities" (GESAMP, 1990). In this chapter, brackish water bodies, ranging in size from coastal lagoons to the Baltic Sea, are also considered subject to marine pollution.
A "contaminant" is a substance that can be detected in an ecosystem above its background concentration, but which has not been demonstrated to give rise to one or more of the adverse effects mentioned above.
When discussing adverse effects of pollutants, some consider the accumulation of pollutant residues in the tissues of organisms to be adverse. Others consider an effect injurious only if changes occur in physiological processes in organisms, such as alterations in cellular morphology, metabolic activity, or physiological rates. Ecologists might restrict this definition still further to only those pollutant-induced effects that give rise to ecologically significant changes, i.e., those at the population level. These latter effects are of prime interest, and they will receive the most attention in this chapter. The other effects mentioned are also important, since they represent stages in the progression of ecological adaptation.
Effects of pollutants can be detected at several different levels of biological organization, ranging from the level of the whole ecosystem to that of the subcellular and molecular. Before selecting methods to detect specific changes, the level of organization (individuals, a population, a community, or an ecosystem) for study must be clearly defined. For instance, detecting changes at the molecular level in one tissue may have little significance for the health and survival of the entire individual. Likewise, alterations at the level of individuals may not be evident at population or community levels (Moriarty, 1983). Assessments may be conducted at any location and at any moment in time, and may be repeated periodically to detect insidious changes apparent only as trends over long time periods. Such an approach may help to estimate the future changes from proposed introductions of new chemicals into the environment. While a desirable goal may be to visit a site only once, make non-destructive measurements, and determine the consequences of specific concentrations of one or more substances on the biota at that site, attainment of this goal is not possible because most methods are empirical, and require temporal or spacial comparisons between control and test sites.
A most difficult issue to address is that when an effect can be characterised as truly adverse. This matter will be addressed only superficially herein; rather the focus is on the detection and quantification of biological changes induced by pollutant exposure. The reader is referred to Forbes and Forbes (1993) for a fuller discussion of the scientific, managerial, and political issues involved in deciding when effects caused by pollutants are of sufficient impact to warrant remedial action.
The most ecologically relevant measurements to assess ecotoxicity are those that describe changes in ecosystem structure and function (Kelly and Harwell, 1989). However, such measurements are often difficult and time consuming to make, and are seldom predictive. Thus, by the time that a significant change can be measured, for example, in nutrient cycling, the ecosystem may have been severely damaged already. Another limitation of this type of measurement is that relating the degree of ecosystem change to a particular level of environmental contamination (even major pollutants) may often be impossible.
Descending through the biological hierarchy, a range of measurements can be made at the level of communities, such as those of algae or encrusting species such as bryozoans, hydroids, or barnacle assemblages (Blanck et al., 1988; Bayne et al., 1985). However, similar measurements are much less practical when dealing with larger, free-moving organisms, such as macroinvertebrates, fish, marine mammals, and birds that have longer generation times.
Holwerda and Opperhuizen (1991) presented a contemporary overview of physiological and biochemical approaches to the toxicological assessment of environmental pollution. However, much of the work presented does not define the relevance of biochemical and physiological effects to consequences in populations or communities. Blanck et al. (1988) postulated that pollutants that do not exert a selection pressure can cause no significant biological effects to an ecosystem, since these substances are unable either to restructure communities or to change the genotypic distribution in the populations. Thus, inter-individual variability in responses to pollutants becomes of major importance, since it is the key to understanding the mechanisms of selection underlying pollution-induced (as well as naturally occurring) ecological change (May, 1986; Depledge, 1990a).
At the cellular and molecular levels, numerous pathological changes and biochemical markers have been identified that signal exposure to pollutants (McCarthy and Shugart, 1990; Huggert et al., 1992). Relationships between particular pollutants at known concentrations and pathological or biochemical responses to which they give rise have been established readily; however, as with physiological effects, relating effects in individuals to alterations in populations and communities has proved to be difficult at best.
When comparing data obtained from monitoring of brackish and fully marine habitats, certain factors should be kept in mind. Species diversity is often greater in fully saline coastal waters than in estuaries and brackish waters (Remane and Schlieper, 1971); thus, pollution may result in the early loss of sensitive species. Many estuarine and brackish water species can tolerate the fluctuating conditions of such environments because they are able to maintain life in such environments (Depledge, 1990b). Hence, these tolerant species may be preadapted to tolerate greater stress such as that arising from pollution (Gray, 1974; Howell, 1984; Depledge, 1990b). Alternatively, the biota of estuarine and brackish waters may already be living near their tolerance limits due to continuous exposure to natural stressors. Some researchers, therefore, have concluded that estuarine and brackish water species are more vulnerable than their marine counterparts to additional pollutant-induced stress. This matter has yet to be resolved, and is discussed more extensively by McClusky et al., 1986.
Salinity differences may also strongly alter the bioavailability of pollutants to organisms, and hence may impact toxicity. With cadmium in water, for instance, reducing salinity also reduces chloride complexation, thereby increasing the bioavailability of cadmium to organisms (Mantoura et al., 1978). However, this situation may not pertain if interactions occur with calcium in the water (Bjerregaard and Depledge, 1992; Depledge, 1990b). Differences in pollutant bioavailability in waters of different salinity emphasize the utility of bioindicator studies in which only the pollutant load accumulating in the organism is assessed.
In the early 1970s, measurements of pollutant residues in aquatic organisms were recognized to be a valuable addition to analyses of water and sediments (Phillips, 1980). The idea behind such biomonitoring approaches is to use an organism's pollutant load as an index of exposure. According to Phillips (1980), the advantages of measuring pollutant residues in aquatic organisms are:
An important reason to measure the concentrations of contaminants in marine organisms is that the analysis of seawater and sediments is difficult and expensive. Usually, contaminant concentrations are much lower in water and sediment samples than in biota (often below analytical detection limits). Furthermore, water current may disperse pollutants, even though organisms remaining at a locality have been contaminated. This situation makes difficult the prediction of the extent of accumulation in marine organisms from the results of analysis of abiotic samples.
Ratios of organism to water concentration usually span several orders of magnitude, even at the planktonic level (Skwarzec and Bojanowski, 1988). For example, tissues of the sea skater, Halobates micans, contain remarkably high concentrations of cadmium, despite the presence of much lower concentrations of cadmium in the water that the organism inhabits (Schulz-Baldes, 1989). The metal appears to be concentrated in the surface microlayer of the seas, and is accumulated by Halobates. The insect clearly provides a far more significant route for cadmium transfer to its predators than would have been predicted from analysis of seawater alone.
Likewise, some pyrethroid insecticides cause substantial acute mortality in estuarine populations at concentrations well below the detection limits in seawater (Schimmel et al., 1983; Clark et al., 1989). Monitoring of water, sediment, and biota from estuaries in South Carolina (US) indicated little potential impact on the fauna resulting from insecticide exposure (Trim and Marcus, 1990). Nonetheless, 30 insecticide-related fish kills representing 11.5 percent of all estuarine kills occurred during the monitoring period. Thus, monitoring data allowed no early detection of either pesticide impact or identification of pollutant sources (Trim and Marcus, 1990).
Quantifying the relative proportions of pollutants taken up from food, seawater, or sediments is important to assess the ecotoxicological significance of particular contaminant loads (Depledge and Rainbow, 1990). The influence of temperature and reduced salinity on bioavailability should also be considered (Newman and McIntosh, 1989), as in the case of hydrocarbon residues from the Amoco Cadiz oil spill which are locked up in anaerobic sediments from which they are being released slowly to contaminate oyster tissues (Berthou et al., 1987). Because this process is so complex, accurate estimates of the rates at which the residues will accumulate in oysters can be made only with much uncertainty.
Both biotic and abiotic variables are known to influence pollutant kinetics, and must be taken into account when interpreting biomonitoring data (Phillips, 1980). Although some strong correlations between environmental pollution and pollutant content in tissues of some species have been found (Phillips, 1980), the interpretation of pollutant body burdens may not be as simple as originally anticipated. Depledge and Rainbow (1990) emphasized that the systemic mechanisms of handling metals and an organism's physiological condition can determine the significance of the body burden of a specific metal. Thus, differences in the partitioning of metals among different tissues may largely influence toxicity, but may be masked when measurements are restricted to those of whole body concentrations (Depledge and Rainbow, 1990). Until a greater understanding of trace metal handling by organisms emerges, monitoring metal loads in biota may not always be a suitable approach to mapping the influence of metal pollution in the environment.
Very few species meet all, or even most, of the criteria for an ideal bioindicator species. Bivalve molluscs appear to be one of the most suitable groups; by far the greatest research effort has been directed towards them (Phillips, 1980). Bivalves are widespread and commonplace, are easily collected in large numbers, and are sedentary. Because bivalves are filter feeders, they pass large volumes of water over their body surfaces, and accumulate pollutants almost continuously. Thus, they act as integrators of exposure over long time periods.
Mytilus edulis has been analyzed most frequently, since it is common, and has an almost global distribution. However, caution should be exercised in using Mytilus edulis as the identifier, because it is now known to be a complex of three species (Lobel et al., 1990). Mytilus has become a part of global Mussel Watch programmes (Bayne, 1989; Cossa, 1988; Cossa, 1989), and has demonstrated interesting trends in pollutant concentrations. For example, Fischer (1989) demonstrated that concentrations of cadmium in Mytilus edulis in Kieler Bight in the western Baltic in 1984 had declined to about 30 percent of the levels in 1975. Reductions in cadmium and lead concentrations in mussels and oysters in the US over a similar time period were found by Lauenstein et al. (1990). The long durations of many of the programmes will ensure that long-term trends in bioavailability of inorganic and organic pollutants to bivalves and their predators can be separated reliably from natural fluctuations (Sericano et al., 1990).
Crustacea have also been used extensively in biomonitoring. Rainbow et al. (1989) identified sites in Scottish coastal waters where concentrations of copper in amphipods were higher than findings in abiotic samples would have predicted, suggesting that copper may have been discharged into the sea from whisky distilleries. This work has continued with studies on copper and zinc in Orchestia gammarellus on North Sea coasts, where several pollution hotspots have been identified (Moore et al., 1991).
Concentrations of copper and zinc in barnacles from contaminated sites are among the highest of any marine organism (Chan et al., 1986; Powell and White, 1990). Transplant experiments have shown that the barnacles reflect environmental levels by rapidly accumulating zinc and copper present at contaminated sites (Al-Thaqafi and White, 1991). They exhibit an uptake-storage mechanism of metal detoxification unlike most other marine Crustacea that exhibit some degree of metal regulation resulting in maintenance of a more or less constant body concentration. Sedentary, adult barnacles are potentially ideal long-term in situ monitors of pollution, particularly of copper and zinc.
Depledge (1990a) emphasized the importance of considering inter-individual differences in pollutant load among the representatives of a population of organisms. Thus, knowing the mean value of a pollutant concentration for a population sample is often insufficient. When sampling organisms or tissues to assess the bioavailability of pollutants at a particular locality, the distribution of pollutant loads in the population must be ascertained for predictions of ecological significance. Depledge and Bjerregaard (1989) and Lobel et al. (1989) pointed out that the frequency distributions of particular trace metal concentrations may not be normal, a fact almost invariably ignored in biomonitoring surveys. The problems that arise by assuming a normal distribution when actually the data are skewed has been clearly demonstrated by Lobel et al. (1982), and Blackwood (1992) has provided a useful discussion of the legitimacy of using the log-normal distribution for describing environmental data.
For zinc in Mytilus edulis collected from the Tyne estuary (UK), the mean zinc concentration was only 75 percent of the mid-range value. In a comparison of three sites contaminated with zinc to varying extents, the lowest tissue concentrations recorded at each site were similar (0.83, 1.5, and 1.11 µmol Zn/gm), whereas the highest concentrations were markedly different (3.32, 10.0, and 20.5 µmol Zn/gm). The distributions of tissue concentrations from animals at each site were skewed positively. Statistical techniques are already available to quantify and compare the residual variability of trace metal concentrations in biological tissues (Lobel et al., 1989).
When measuring effects of chemicals on ecosystems, the uniformity, intensity, frequency and duration of exposure, and uniqueness of the chemical (i.e., whether found normally in ecosystems but in lower concentrations or whether it is man-made) are all important to consider. These and other factors have been discussed in detail by Kelly and Harwell (1989). An extensive debate has been ongoing as to whether changes in ecosystem structure precede changes in function or vice versa; however, this issue remains unresolved (Ford, 1989). The effects of chemical stressors are manifest initially as a loss of sensitive species and changes in the relative abundance of rapidly reproducing taxa. However, changes in ecosystem processes do not always occur as pollution increases, nor is the rate of ecosystem change constant when it does occur (Levine, 1989). Since some ecosystem processes (e.g., nitrogen and phosphorus cycling) can be taken over by other species as more vulnerable species are eliminated, structural changes still offer probably the best prospect of signalling pollutant-induced change in most aquatic ecosystems.
Numerous authors have attempted to find indices of overall ecosystem health, such as the rate of photosynthesis, the photosynthesis-respiration ratio, and the activity of electron transport systems (Woodwell, 1962; Ivanovici and Weibe, 1981; Levine, 1989). Critical reviews of the concept of ecosystem health have recently been issued by Calow (1992) and Rapport (1992). A key problem with this approach is that little attention has been paid to relating changes in ecosystem health to particular degrees or types of pollution. Babich et al. (1983) attempted to overcome this limitation by suggesting establishment of dose-response relationships for ecosystems in which a pollutant load which causes a 50 percent or a 10 percent reduction in an ecosystem process is determined (EcD50 or EcD10, respectively). So far, these indices have proved to be of no practical utility.
Critical concentrations for pollutants in soils and sediments have been proposed to be set al levels that protect 95 percent of species from poisoning; however, this concept has limited acceptance for the marine environment (Hopkin, 1993).
For marine benthic communities, several studies have revealed important changes in species composition following exposure to organic pollution from pulp mills and oil spills (Pearson, 1970, 1971, 1975; Rosenberg 1972; Rosenberg, 1973; Sanders et al., 1991). In these and other instances, the affected communities are dominated by a very few species, usually of annelid worms. Such pollution indicator organisms are potentially useful to assess the extent to which benthic communities are affected. Ecological succession occurring in response to a pollution gradient is complex, and may be influenced by a great many factors other than pollutant exposure (Pearson and Rosenberg, 1978), so the approach should be used cautiously. Nevertheless, changes in community composition appear to be reasonably consistent when addressed using the above approach.
Reductions in species diversity, retrogression to opportunistic species, and shifts to smaller sized species are all well-documented responses of communities to stress (Forbes and Forbes, 1993). Diversity indices usually integrate the relative abundances and numbers of species in a community, and provide an estimate of community complexity. A fall in diversity usually indicates significant pollutant-induced change (Gray, 1980), but predation, competition, spatial heterogeneity, and successional change may also influence diversity indices, and have led to valid criticisms by Gray (1980). Thus, Ford (1989) concluded that general agreement exists that diversity indices do not reliably reflect pollutant disturbances over time or space.
Practical difficulties associated with the use of diversity indices include problems with the accurate identification of organisms to the species level. While this characterization may be accomplished in well-studied, temperate, brackish water, and estuarine ecosystems, greater problems may be encountered in coastal regions. Furthermore, practising this approach in the subtropics and tropics (where species diversity increases enormously and where many species have yet to be classified) may be impossible at present.
One of the simplest methods to detect pollution-induced changes in communities of marine benthic organisms is to analyse the log-normal distribution of individuals per species in sediment samples. This approach assumes that a community equilibrium exists (Connell and Sousa, 1983; Williamson, 1987). For example, temporary departures from a log-normal distribution may occur during seasonal recruitment of juveniles. Gray (1980), therefore, recommended that monitoring year-to-year variations in species abundance could be conducted most effectively during the winter in temperate habitats or at those times of the year when populations are not actively recruiting. A detailed study has recently been conducted by Ferraro and Cole (1992) on the taxonomic level that needs to be adopted to assess the impact of pollutants on macrobenthic communities.
In many samples of benthic communities, the most abundant class is not that represented by one individual per species, but often lies between classes with either three or six individuals per species. Thus, the curve relating the number of individuals per species (x axis) to the number of species (y axis) is often strongly skewed. This curve can be converted to a normal one by plotting the number of individuals per species on a geometric scale (usually 2x). Plotting the geometric classes on the x axis (Class I = 1, Class II = 2 to 3, Class III = 4 to 7, Class IV = 8 to 15, and so on) against cumulative percent of species on the y axis invariably gives a straight line. At polluted sites, a break in the line often exists, indicating departures from an equilibrium community. A persistence of this break over several sampling occasions is indicative of pollution-induced disturbance. The log-normal distribution procedure has been described by Gray (1981).
Pollution-induced community tolerance (PICT) can be used in the assessment of pollution effects. Blanck and Wangberg (1988) investigated PICT in periphyton communities established on 1.5 cm2 glass discs. When exposed to severe arsenate stress, the communities developed 17000-fold tolerance to the pollutant, as measured by photosynthetic activity. Arsenate exerted a selection pressure leading to replacement of sensitive species by tolerant ones. This effect caused overall arsenate tolerance of the community to increase. The stress also affected the rate of increase of biomass. An important consideration when using this approach is that exposure to one pollutant may confer tolerance to another (i.e., co-tolerance). Although this effect was reported to be insignificant for arsenic (Blanck and Wangberg, 1988), situations can be envisaged in which tolerance to a particular trace metal or organic pollutant may be due to exposure to a quite different trace metal or organic compound having the same or similar mechanism of action, leading to the incorrect conclusion that a community has been exposed to a chemical which actually it has never encountered.
Warwick et al. (1990) studied the structure of benthic communities in relation to pollution in Hamilton Harbour, Bermuda. Multivariate analysis of the fauna detected differences in community composition that could be related to the pollution gradient in the Harbour. For the statistical analysis of the macrobenthos, aggregation of species data to family level was acceptable. However, for nematodes, aggregation from genus to family level resulted in a significant loss of information.
One of the most detailed studies on the responses of marine organisms to a putative pollution gradient was conducted on the sublittoral fauna of the Frierfjord-Langesundfjord (Norway) by Gray et al. (1988). This valuable research includes multivariate statistical analyses, which discriminate among sites on their faunistic attributes, and univariate measures of community stress, and provides a very comprehensive, comparative account of methods for assessing community responses. The effects observed were actually attributed to seasonal anoxia rather than pollution.
Gray (1979) commented on the surprising paucity of information on the effects of pollutants on populations. Since then, some notable studies have been added to the scientific literature. Nonetheless, population-level effects of pollutants are perhaps the most neglected area of ecotoxicological research.
An excellent analysis of stress (including pollutant-induced stress) on natural populations was carried out by Underwood (1989). He noted that experimental studies on populations are difficult to conduct because of inertia (a lack of response in a population following exposure to pollution), resilience (the capability of the perturbation from which a population can recover), and stability (the rate at which a population recovers from a stress). The extent to which a population has the above attributes determines its range of responsiveness to intermittent, temporary , and acute and chronic pollutant exposure.
The timing of exposure to stress relative to the life cycle of individuals within the population is also critical. For instance, experimental deletion of patches of seaweed from a sublittoral kelp bed at certain times of the year leaves the kelp population depleted until after the following reproductive season. However, if kelp patches are removed when the plants are reproducing, little effect can be observed (Kennelly, 1987).
Effects of pollutants on populations may involve loss of individuals due to smothering with oil, poisoning, or incapacitation, which separately render some organisms more susceptible to predation or to mortality caused by other environmental stressors. Alternatively, population-level effects may emerge due to alterations in individual growth rates, reduced fecundity and longevity, and disturbance of endogenous biological rhythms (Gray, 1979; Moriarty, 1983; Depledge, 1984).
A most effective means of detecting changes in populations over time is in the use of life tables (Morris, 1959; Varley and Gradwell, 1960; Varley et al., 1975). Life tables permit analysis of the role and importance of different mortality factors in relation to the overall mortality rate of individual populations (Begon and Mortimer, 1986). However, distinguishing between pollution-induced effects and those related to fluctuations in natural environmental factors may be difficult.
Luoma (1977) commented that the existence of one population of a species that is more resistant to a toxicant than another is direct evidence that the concentration of the toxicant in the environment of the resistant population is sufficient to elicit adverse biological effects. The presence of a toxicant-resistant population of one species in an ecosystem further suggests that other species may have been affected by the resistance-eliciting substance. Luoma (1977) stated that populations resistant to a particular pollutant are found only in areas of known contamination. An illustration is populations of the polychaete, Nereis diversicolor, found in English estuaries contaminated with copper, cadmium, lead, and zinc that are more tolerant to high concentrations of copper in seawater than are N. diversicolor found in less contaminated areas (Bryan and Hummerstone, 1971, 1973; Bryan, 1974). Thus, the implied selection of resistant genotypes within a population exposed to a contaminant constitutes the basis of the PICT assessment method. However, for N. diversicolor, enhanced genetic resistance to cadmium or lead has not been detected.
Futyuma (1986) reviewed data from several fields of research showing that genetic mutations conferring resistance occur independently of exposure, and, therefore, are not induced by pollutants. In support of this conclusion, Weeks and Depledge (1992) recently reported marked differences in mercury tolerance among three populations of the amphipod, Platorchestia platensis, living in similar habitats within 20 km of one another. None of the populations has been exposed to significant trace metal contamination in situ.
Gray (1979) stated that, in habitats under severe pollution stress, the species that dominate are those that have flexible life histories. Species having less flexible life histories increase in abundance under conditions of slight pollution. Thus, Gray (1979) concluded that the presence of a particular species in a polluted area may be more closely related to its life history strategy than to tolerance to adverse conditions. Nonetheless, when an intense selection of genotypes takes place, the population that survives at a polluted locality would be expected to have a different genetic composition when compared with populations of the same species in clean conditions. Modern genetic techniques have the potential to detect such phenomena (Bickham et al., 1986).
The expression of particular genes confers the ability to survive in polluted conditions. Thus, phenotypic differences between populations of a species occupying clean and polluted conditions might be expect to occur. Such differences might be manifested as alterations in morphology, behaviour, and biochemical-physiological characteristics. Depledge (1992) has argued that of these factors, particular biochemical-physiological traits are most likely to confer resistance to chemical toxicity .He proposed that, by looking for changes in the relative proportions of resistant versus sensitive phenotypes (or "physiotypes" as they were called to emphasize the importance of biochemical-physiological characteristics), detection of changes in the tolerance distribution of individuals within populations exposed to marine pollutants may be possible. Similar studies have been conducted for many years to assess the development of resistance to pesticides among insects (Wood and Bishop, 1981) and by evolutionary biologists (Via and Lande, 1985; Schluter, 1988; Anholdt, 1991; Forbes and Forbes, 1993).
Invertebrates that have been shown by breeding experiments to be genetically resistant to high concentrations of metals include marine polychaetes (Grant et al., 1989). These authors have suggested that genetically based tolerance to metals in polychaetes could be mapped, and used as an in situ monitor of the ecological impact of pollutants.
Several biochemical techniques are now available to detect genetic changes in organisms exposed to pollutants (Shugart et al., 1992). For example, the DNA alkaline unwinding assay can be used to detect DNA-damaging substances in marine animals exposed to environmental pollutants. In gills of blue mussels caged at the New Bedford Harbour Superfund Site (Massachusetts, US) which is highly contaminated with many organic and inorganic substances, a significant increase in DNA strand breaks was detected after three days of exposure (Nacci and Jackim, 1989). In fish, DNA-xenobiotic adducts persisted in the liver for much longer than aromatic hydrocarbon (AH) metabolites that induced them (Varanasi et al., in press). Thus, these effects on DNA may have a practical use in integrating cumulative exposure of fish to AHs.
Phillips (1980) described the use of organisms transplanted from clean sites to polluted sites or vice versa to examine the accumulation or depuration of pollutants. The value of such an approach was demonstrated in studies in Hong Kong on the uptake and release of PCB isomers in transplanted mussels (Perna viridis) (Tanabe et al., 1987). Similar experiments have been attempted in which parameters of the well-being of organisms have been measured after transfer from one site to another (e.g., scope for growth in transplanted mussels) (Tedengren et al.,1990).
The responses (in terms of changes in species composition) of fouling communities to pollution stress can be monitored in situ by reciprocal transplants. Climax communities are allowed to develop on submerged surfaces in a clean and a polluted site, and are then moved between sites. An experiment in Australia at Woolongong Harbour (uncontaminated) and Port Kemblar Harbour (polluted by discharges from heavy industry ) demonstrated rapid changes in community structure in response to pollution (Moran and Grant, 1991). Within two months after transfer, communities moved from Woolongong to Port Kemblar were similar in structure to those that had developed wholly at Port Kemblar. Most changes occurred in the short term when sensitive species were killed by periodic discharges-an effect difficult to predict by measuring levels of pollutant in water. Space previously occupied by these species was quickly colonized by opportunists tolerant of pollutants, giving rise to changes in community structure.
Changes at the ultrastructural level can also be examined in organisms transplanted between clean and contaminated sites. Thomas and Ritz (1986) showed that much of the zinc accumulated by the barnacle Elminius modestus transplanted to a zinc-contaminated site was stored with phosphate in intracellular granules. These granules increased in size in proportion to the length of exposure. Few granules were lost from contaminated organisms when they were transplanted to a clean site. The full potential of transplant experiments in ecotoxicology has yet to be explored.
Caging organisms in mesocosms provides a means of bridging the gap between the laboratory and the field (Kuiper and Gamble, 1988). Mytilus edulis caged in the effluent from a factory producing titanium dioxide, accumulated titanium in excretory granules (Ballan-Dufrancais et al., 1990). Mitochondria in the bivalves contained lesions and reduced cristae, indicating a serious effect of the effluent on the respiratory metabolism of the organisms (Coulon et al., 1987).
The most powerful tools for the investigation of pollutants in situ are biomarkers. Depledge et al. (1992) and Depledge (1993) have discussed the rational basis of the biomarker approach. In the past, a variety of cellular biochemicals have been measured in tissue samples from aquatic animals (less so plants), and their concentrations have been related to exposures to specific pollutants or to various classes of pollutants. This approach has achieved very little, because biomarker responses are of unknown ecological significance. For example, a rise in mixed function oxidase (MFO) in the livers of fish taken from a polluted area may signify pollutant exposure; yet, the fish may continue to grow and reproduce normally, and the MFO response may be viewed as part of an acclimatization process to altered environmental conditions rather than a manifestation of an injury.
Much can be gained by relating biomarker responses to changes in Darwinian fitness parameters (Depledge, 1993). As an illustration, Sanders et al. (1991) measured stress protein responses and scope for growth (SFG) in Mytilus edulis exposed to various concentrations of copper in the laboratory .By relating the two, the investigators showed that stress protein responses occurred prior to reductions in SFG. As copper exposure increased further, SFG decreased and stress protein responses were even more marked. Thus, the stress protein biomarkers could be used to signify a change in Darwinian fitness that may have consequences for the entire population.
Recently the biomarker concept has been extended from purely biochemical measurements to include those of cellular pathology, physiological processes, and even behaviour of organisms exposed to varying concentrations of pollutants (Depledge et al., 1992; Sanders et al., 1991). This enhancement creates the possibility of using a hierarchy of biomarker measurements (Depledge, 1993). Initially, effects of pollutants might be detected by relatively non-specific biomarkers, usually high in the hierarchy (e.g., behavioural and physiological biomarkers). Detection of abnormalities with these non-specific biomarkers at a site at risk from pollution might then justify the measurements of more costly, lower hierarchy, specific biochemical and cellular biomarkers (e.g., MFO activity, metallothioneins, intracellular granules, and tissue lesions) to seek to identify the class of pollutant responsible for the exposure. The magnitude of biomarker responses together with determination of tissue residue concentrations of pollutants would contribute to the overall assessment of pollutant impact.
Biomarkers may also have an important role in unravelling the interactions between natural environmental stressors (e.g., hypoxia, thermal and salinity stress) and pollutants, and the effects of mixtures of contaminants in areas receiving several pollutants (Livingstone et al., 1988). Where more than one stressor (natural factors plus one or more pollutants, or complex mixtures of pollutants alone) is present, the substances may act synergistically so that the combined effect is greater than the sum of the effects of the individual chemicals (Depledge, 1987; Walker and Johnston, 1989). However, in some circumstances, stressors (including pollutants) can act antagonistically. TBT and hydrocarbons, when present together, caused a much lower reduction in clearance rate in Mytilus edulis than had been predicted on the basis of proportional additivity (Widdows and Donkin, 1991). In their excellent summary of this problem, Widdows and Donkin (1991) described how reductions in SFG in Mytilus edulis in contaminated sites can be apportioned between specific pollutants. In a study conducted in Bermuda, Widdows et al. (1990) showed that the overall reduction in SFG of Mytilus edulis could be proportional such that, at the most contaminated sites, TBT accounted for 21 percent and hydrocarbons 74 percent of the observed effects.
When measuring biomarker responses in organisms sampled from the field (or when measuring pollutant residue levels in biota), only those organisms that have survived at the locality are considered. In moderately or severely polluted environments, situations can be envisaged in which the mean biomarker response is high initially and then declines, a change of response that might be due either to the return of clean conditions or to the dying out of the most severely affected organisms, so that only resistant or less severely stressed organisms remain. This complex matter deserves further study.
Biochemical markers (also called "biomarkers") of pollutant exposure and effects have been reviewed by McCarthy and Shugart (1990), Stegeman et al. (1992), and Fossi and Leonzio (1993). These works focus on the use of biomarkers to signal pollutant exposures rather than to detect injury to ecological systems. This limitation may reflect the complexities of validating the latter approach. Table 7.1 gives examples of biomarkers that have been used most often and pollutants known to initiate responses (Stegeman et al., 1992).
As an example, microsomal cytochrome P450 enzymes normally function at low rates in the liver of fish, but are induced by chemicals with relatively flat molecular structures such as aromatic hydrocarbons (AH), dioxins, and co-planar PCBs. Thus, P450 induction has been advocated as a reliable indicator of organic chemical contamination in marine systems (Stegeman et al., 1990). In the liver of dab and other fish, the induction of cytochrome P450 (IA family = P450IA) is highly correlated with activity of the enzyme 7-ethoxyresorufin O-deethylase (EROD) (Goksyor et al., 1989). Although the EROD assay is easier to perform than Western blotting or ELISA (enzyme linked immunosorbent assay) for P450IA, circumstances exist whereby pollutants (e.g., PCBs) can inhibit EROD measurement (Goksyor et al., 1991).
Table 7.1. Examples of biochemical markers of pollutant exposure
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| Biomarker | Pollutants initiating response |
| P450 | Aromatic hydrocarbons, dioxins |
| Metallothioneins | Cd, Cu, Zn, Hg, Co, Ni, Bi, Ag |
| Stress proteins | Thermal pollution, TBT, Cu and other |
| metals, PAH, UV radiation | |
| Glutathione transferases | PAH, PCB, BNF |
| Lipid peroxidation | Cd, PCB |
| Heme and porphyrins | Pb, As, Hg, PCB, dioxin, HCB |
|
|
|
A very useful approach to assess dose-dependent pollutant toxicity in the marine environment is to observe and measure chemical-specific (e.g., tributyl tin, TBT) morphological changes in a widely distributed organism (molluscs, such as, the dogwhelk, Nucella lapillus). The extent of the changes reflects the degree of environmental exposure.
Organotin exposure in the marine environment occurs primarily as a result of TBT leaching from antifouling paints applied to the hulls of boats to inhibit settlement of marine invertebrates, especially barnacle and mollusc larvae. Other sources of organotin include PVC manufacture, wood preservatives, and general biocides (Fent et al., 1988).
A combination of traditional laboratory toxicity studies, transplant experiments, and field observations have shown that TBT is probably the most toxic man-made substance ever to have been deliberately introduced into the marine environment (Bryan et al., 1986). TBT is toxic to many marine organisms, and has had a severe economic impact on oyster fisheries and farms (Waldock et al., 1983). In Australia, shell curling in oysters was induced in a hitherto pristine lake by mooring just two TBT-coated boats for only one month (Scammell et al., 1991). In some European estuaries and coastal waters, TBT has had a dramatic effect on populations of deposit-feeding bivalves and the dogwhelk, Nucella lapillus, that are extremely sensitive to TBT (Langston et al., 1990). Dogwhelks are now absent from many seashores in the UK, where they were common before the introduction of TBT paints (Bryan et al., 1986).
TBT causes female dogwhelks to grow a vas deferens and a penis. These appendages block the opening of the female genital duct so that egg capsules cannot be released (Gibbs and Bryan, 1986). TBT appears to affect the hormone system that determines the sex of prosobranch molluscs (Gibbs et al., 1988). Similar findings have been obtained with other Nucella species and related genera elsewhere in the World (Bright and Ellis, 1990; Stickle et al., 1990; Smith, 1980).
The phenomenon in which male sex characters are imposed on the female snail is called "imposex," and is widespread in stenoglossan gastropods (Smith, 1980). The mean size of the female penis relative to males in a population of dogwhelks provides a sensitive indicator of the degree of imposex and hence of exposure of dogwhelks to TBT. Thus, determining the level of imposex provides a much easier and cheaper method to assess the bioavailability of TBT than direct measurement of the chemical in seawater (Langston et al., 1990; Spence et al., 1990).
A description of the major physiological responses of marine organisms to chemicals is presented by Vernberg and Vernberg (1974); Vernberg et al. (1977, 1982), Dorigan and Harrison (1987), and Holwerda and Opperhuizen (1991). The rationale underlying extensive physiological monitoring, revealing disturbances of almost all physiological systems, has been fully discussed by Depledge (1989), providing encouragement that a strategy for changes in physiology to be related to alterations in the Darwinian fitness of individuals.
Acquisition of physiological data, whether in the laboratory or in situ, has been severely hampered by the lack of suitable transducers capable of recording data over long periods, from several test organisms, simultaneously, without imposing undue stress. However, the introduction of computer-aided monitoring systems and non-invasive transducer techniques is beginning to alleviate this situation (Depledge and Andersen, 1990; Aagaard et al., 1990).
Mayer et al. (1992) recently reviewed methods to assess chemical-induced changes in whole animal physiology. Scope-for-growth has proved to be the most useful in situ assay procedure. However, much work remains before physiological monitoring can be performed routinely. As with population- and community-level approaches, responses detected at the physiological level are often difficult to relate to specific substances at known concentrations. Nonetheless, an integrated approach involving biochemical, physiological, and behavioural biomarkers may eventually prove valuable.
Detailed accounts concerning the relationships between tissue pathology and diseases in marine organisms associated with pollution have been provided by Bucke and Watermann (1988), Hinton et al. (1992), Weeks et al. (1992), Hinton and Lauren (1990), Cormier and Racine (1990), Yamashita et al. (1990), and McMahon et al. (1990). A specific example of this approach was the study by Carr et al. (1991), which showed that the number and extent of hepatic lesions in winter flounder from the polluted waters of Boston Harbor were correlated with low tissue concentrations of ascorbic acid and hepatic glycogen (effects not observed in reference populations from clean sites). Grotesque deformities in smelt (Osmerus eperlanus) from the Elbe Estuary were observed by Pohl (1990).
Deformed fish had significantly higher lead and cadmium concentrations in their livers than did normal specimens, but the levels were still quite low in comparison to those at other polluted sites. Concentrations of each of the many pollutants in the Estuary were below critical levels, but when present together, the mixture of contaminants may have been related to the deformities.
The preceding account highlights the wide range of approaches available to assess the effects of pollutants in marine, estuarine, and brackish water environments. Identification of the type of injuries to be detected is necessary before selecting a measurement technique. No single approach is satisfactory .This review emphasizes the value of an integrated assessment approach involving several methods, each focusing at a different level of biological organisation. For an assessment useful in ecotoxicological decision-making, the fate of pollutants must be related to effects to which they give rise. Transplantation experiments and the use of biomarkers of toxicity that link responses to pollutants with changes in Darwinian fitness of organisms offer the greatest potential for future development as assessment tools.
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