Ecosystems provide ecologic goods and servicesincluding clean water, air. and soil-for wildlife and human activities such as agriculture, hunting, fishing, recreation, and aesthetic values (SCOPE 1996). In addition to their own intrinsic value, wildlife species are critical parts of the structure and functioning of ecosystems that support human activities. Evidence has been accumulating that environmental contaminants are affecting their populations (Colborn and Clement 1992). Ecotoxicologic studies have been the primary field providing data on ecologic effects of hormonally active agents (HAAs), particularly from laboratory-based studies examining effects of HAAs. Other chapters in this volume focus on the effects of HAAs on individual animals and humans; this chapter examines their effects on populations, communities, and ecosystems.
In this ecologic perspective, the committee has focused more on effects than on mechanisms. This is because ecologic effects are usually the aggregate of effects on individuals and, therefore, they are even harder to pin down than are effects on individuals. For individuals, controlled laboratory experiments with many replicates are often possible, whereas for wild populations, and especially for biologic communities and ecosystems, multiple replication is usually very difficult. Thus, ecologic studies can elucidate effects of suspected HAAs on populations, communities, and ecosystems, but cannot usually determine whether a chemical is an HAA.
A useful way to understand and model ecologic effects is through life-history dynamics. Although ecologic mechanisms might be difficult to pin down, a known effect on individuals can be extrapolated to populations through life-history models. The result of such modeling can then suggest what population (and even community-level) effects might be expected and, thus, could be valuable guides to research.break
In the first section, the theoretical and science-based conceptual framework for studying the effects of contaminants is examined, and criteria for determining whether those agents do affect the structure and functioning of communities and ecosystems are discussed. In the second section, case studies illustrate what is known about the ecologic effects of contaminants on fish, birds, alligators, and mammals. The criteria established in the first section are used to evaluate evidence that ecologic effects have occurred as a result of environmental contaminants, especially HAAs. Although the mechanism of action has not been established for some of the adverse effects, and it is uncertain if or how they involve the endocrine system, it is clear that some synthetic hydrocarbons have had adverse effects on some populations of some species. The population effects often have been due to the effects of small concentrations of persistent and bioaccumulative synthetic halogenated hydrocarbons on reproduction and development, such as embryo lethality and deformities (Giesy et al. 1994a). Finally, conclusions and recommendations are presented.
Nature of Ecologic Effects
Ecologic Effects in Humans and Wildlife
Preceding chapters have devoted much attention to humans, but ecologically, humans are only one of many species. With respect to human health, the unit of concern is usually the health and well-being of individuals, including individuals who are past their reproductive years. Occurrence of cancer and other diseases in individuals is an end point commonly used, and the human-health risks of exposure to toxic agents are often assessed in terms of the increase in probability that an individual will contract a disease after exposure to a given agent. Epidemiologists do study populations, but the motivating concern is the health of individuals. Wildlife are more than sentinels because their physiology is similar to that of humans. Their well-being is an important end in itself.
Ecologic effects are those that manifest themselves at the population level or higher. Effects on individuals of wildlife species are important, but they fall under the heading of wildlife toxicology or physiology and are considered elsewhere in this report (Chapters 5-9). Indeed, the potential for HAAs and other agents to produce ecologic effects is mainly through the aggregation of their effects on the physiology of individuals. Although effects on individuals are measured, and the experimental unit is the individual organism, when statistical measures are applied and rates or probabilities of responses are reported, effects are manifested at the population level of organization.
For wildlife species, the most common unit of concern and study is the population. The end points of concern are changes in population size or reproductive capacity. Although it is often necessary to examine adverse effects in individuals to understand population changes, the population is the most impor-soft
tant level of organization (NRC 1993; Burger and Gochfeld 1996). Ecologic effects can also manifest themselves at higher levels of organization: biologic community, ecosystem, and landscape. A biologic community is an assemblage of species that interact or at least co-occur. An ecosystem consists of one or more biologic communities and the physical and biologic environment. Ecosystems can be as small as a pond or as large as one of the Great Lakesindeed the entire planet could be considered a single ecosystem (Keeton 1972). A landscape is an array of biologic communities and ecosystems (Forman and Godron 1986). Each of these levels of organization embodies information that is not included in lower levels; thus, they have emergent properties that cannot necessarily be predicted from information on the lower levels. Because one ecosystem can include hundreds or thousands of species and habitats, evaluating the effects of a single disturbance is complex and more difficult than determining the effects of a chemical or chemicals on an individual organism (NRC 1981 ).
Population effects are usually more difficult to detect than are effects on individuals, and it is usually difficult to identify their causes. Because there is much background variability in population numbers and structure and because it is usually more difficult to obtain accurate population data for most species than it is to get information about the physiology of individuals, detection of effects at higher levels of organization is more difficult still, although changes in populations of key predator or prey species, for example, can have effects on some other species in a community. For example, contaminants in fish have affected fisheating birds in the North American Great Lakes, as described below.
Differences between how humans and wildlife species are viewed can influence the kinds of studies and analyses needed to evaluate evidence that HAAs in the environment cause harm. To be sure, the potential for population-level effects in humans would cause great concern, but the potential for individual abnormalities and increased incidence of disease is a much greater cause for concern with respect to humans than it is for wildlife. Because of these differences, this chapter focuses on population- and higher-level effects on nonhuman species. But the methods and analyses described or evaluating effects of HAAs on individual humans described in other chapters of the report can be applied to individuals of wildlife species as appropriate.
Effects and Measurement End Points
Ecologists have used many measures to examine the structure and functioning of populations, communities, and ecosystems (Table 10-1 ). Population-level measures are perhaps the most useful for determining adverse effects because they are the most direct; the population is the level of organization at which effects are likely to occur first. The measures include population size, age structure, sex ratios, recruitment, and biomass (NRC 1981). In most cases, knowing population characteristics before and after a disturbance will suggest hypothesescontinue
regarding the ecologic significance of the event. Understanding changes in population metrics as a function of contaminants might provide warning of potential population changes (NRC 1986a, 1995).
Several measurement end points can be used to determine the effect of any perturbation or chemical exposure on the structure and functioning of communities and ecosystems (NRC 1981. 1986a, 1993). Unlike measures of individual and population abnormalities, most of these require greater interpretation before cause-and-effect relationships can be demonstrated and effects of contaminants on wildlife understood. For example, species diversity is generally a useful measure of the structure of a community or ecosystem. In its simplest form, species diversity refers to the number of species in a given community or ecosystem. However, simply knowing of the species before and after a disturbance is not enough information by itself to decide whether an adverse effect has occurred (NRC 1986a).
Ideally, there should be some easily measurable end points to determine whether specific chemicals, such as HAAs, adversely affect the structure and functioning of communities and ecosystems. However, major problems are involved in identifying ecologic hazards, such as determining geographic bound-soft
aries, identifying target populations, choosing end points at the community and ecosystem levels, designating indicators or indices, recognizing temporal and spatial scales, identifying the effects of successional stages, and identifying appropriate reference points or systems for comparison (NRC 1986a: Norton et al. 1992; Burger and Gochfeld 1996).
Pyramid of Effects
Because communities and ecosystems are composed of many species and habitats, pyramid effects can occur, including bioaccumulation, biomagnification. cascading, keystone effects, and matrix effects.
Bioaccumulation is the storing of chemicals over time, leading to increasingly greater concentrations in the tissues of an organism. Given the same relative exposure, organisms that live longer usually accumulate more (have a greater body burden) than will organisms that have a shorter life span (Phillips 1993).
Biomagnification is the increase in exposure and accumulation that is observed as one advances up the food chain (Maedgen et al. 1982: Calabrese and Baldwin 1993; Genoni and Montague 1995). Chemical accumulations in organisms at high trophic levels can be far greater than they are in low trophic-level organisms (e.g., Buckley 1986).
Cascading effects are effects that follow from one effect on components of an ecosystem (Carpenter et al. 1985: Lipton et al. 1993). HAAs that selectively harm top carnivores can affect predator-prey relationships, and thus can indirectly change a species' reproductive performance by changing the effectiveness of its predators. Related is the idea of the keystone species (Paine 1966), which posits that removal of one predator can change the population sizes of several other species in a community. Other predator populations might increase because of lack of competition; prey populations might increase (at least initially) because of a lower predation rate and prey composition could shift because of increases in competition among prey organisms.
Matrix effects are those that occur in adjacent communities or ecosystems by virtue of their proximity.
Although the statements above are general, the processes they describe have been abundantly demonstrated. The committee has not provided a catalog of specific examples of bioaccumulation and biomagnification because many factors affect those processes. The factors include abiotic ones, such as seasonal climatic variations, and biotic ones, such as the animal's lipid content. It has thus been difficult to quantify bioaccumulation and biomagnification in most cases.
Susceptibility, Variability, and Evolution
Nearly all traits (anatomic, physiologic, or behavioral) that have been examined in humans and in wildlife show intraspecific variations (Darwin 1859: Fal-soft
coner 1960). These variations can lead to differences in susceptibility to chemicals, including HAAs. In terms of human-health risks, susceptibility refers to differences in genetically and nongenetically mediated susceptibility, and it can be affected by lifestyle and the environment (Omenn 1982; Woodhead et al. 1988). Ecologists have lagged behind in examining genetic and epigenetic (for example, hormonal) susceptibility of wildlife populations. Susceptibility differences could be important because they could be another form of selection within populations; that is, individuals that are less susceptible to toxicants would be more likely to contribute to the next generation (Fox 1995). Such an effect, although possibly important, would be virtually impossible to attribute to any particular cause, and so would not be useful in evaluating evidence of ecologic effects of HAAs. However, various techniques of genetic analysis, including those of molecular biology, could be useful for identifying potentially significant changes in population genetics and gene function (such as differential methylation and thus, gene expression) before their effects become apparent in such measures as population size or reproductive capacity. Moreover, because of differences in structure, some ecosystems are more susceptible than others (Burger 1997).
The generation time of organisms affects their response to selective pressures (evolution) (Lewontin 1965). For example, organisms with short generation times-insects and bacteria, for example-can rapidly evolve resistance to pesticides and antibiotics (NRC 1986b). They also should have the potential to adapt to HAAs. On the other hand, long-lived organisms-including turtles, elephants, and some bird and fish species-would be much slower to evolve adaptations to toxic pollutants; and they also are more susceptible because of bioaccumulation and biomagnification. Thus, in the absence of specific information, more attention should be paid to long-lived species in looking for ecologic effects of HAAs than to very short-lived species at low trophic levels.
Temporal and Spatial Scales of Effects
Temporal and spatial scales of exposure and effects are critical in evaluating evidence on the effects of environmental contaminants. Species sometimes are distributed evenly over the available and suitable habitat or they can occur in patches or discrete populations. The distribution pattern influences both their exposure and the possible effects of any toxicant. Some species experience continuous exposure; others have periods of low exposure, either because of physiologic mechanisms (hibernation) or behavior (migration). One aspect bears special comment with respect to HAAs in the environment. Many purported HAAs are distributed unevenly regionally and globally (see Chapter 3). Uneven distribution makes evaluation of the evidence more difficult, and it provides the potential for widespread and variable effects.break
Risk Assessment and Ecologic Effects
Risk assessment is one method of evaluating the effects of HAAs on ecologic systems, from the individual to the entire ecosystem. Although in the study of ecology the method borrows the four-part paradigm used for human-health risk assessment (NRC 1993), it differs significantly in that the problem-definition stage is far more complicated because of the inherent complexity of ecologic systems (Bartell et al. 1992; Norton et al. 1992: Burger and Gochfeld 1996) and because of the large variety of life-history patterns (Burger 1994). The choice of end points to evaluate is difficult because ecologic-risk end points include those for individuals, populations, communities, ecosystems. and even the biosphere (Sheehan 1984a; Suter 1990: Peakall 1992).
Use of Indicators
Because of the inherent complexity of ecologic systems, it is sometimes useful to identify biologic indicators of exposure and ecologic effects (Hunsaker et al. 1990a,b; Suter 1990; Kendall and Akerman 1992). Several investigators have addressed this topic for communities and ecosystems, including Sheehan (1984a,b). Cairns et al. (1992), Heimbach et al. (1992). Kendall and Akerman (1992), Peakall (1992). Weber et al. (1992), and Linthurst et al. (1995). Biologic indicators are often used to see whether toxicologic effects that occur in the laboratory mirror what happens in nature.
Framework for Evaluating Ecologic Effects
The preceding discussion leads to the framework that the committee used to evaluate information on ecologic effects of HAAs. Ecologic effects can be manifested at different population levels and can be preceded by changes in the population and genetic structure of various species. It is important (but not required) first to establish whether an agent is present in the environment, at what concentration, and how it is distributed. If the agent is not present or the ecologic systems of interest are not exposed to the agent. then one can conclude that agent is not a factor in that system. However, care must be taken when analytic techniques for detecting and measuring chemicals are less sensitive to disruption by the chemicals than are organisms.
It is important to establish plausible modes of action for a suspected agent. This, usually done through laboratory studies, is especially important if the presence or concentration of the suspected agent in the environment or organisms' exposure to it cannot be established easily, as might be the case if a suspected chemical can have an effect at extremely low concentrations.
If changes in population size, structure, growth rate, or other aspects of population dynamics are established, then it is necessary to link the suspectedcontinue
agent to those changes. It is also useful to rule out other possible causes of the changes. Next, changes in higher-level ecologic factors, such as community composition or primary productivity, must be plausibly related to established population-level effects or directly to the suspected agent. Whether the mode of action of the suspected agent is through a hormonal mechanism must be established. It is rarely, if ever, possible to establish all of the above relationships. Thus, one must use judgment, based on a weight-of-evidence approach, as described in the case studies below.
Effects on Populations and Communities
Studies of selected communities and individual populations can be used to illustrate what is known about the ecologic effects of environmental contaminants on fish, birds, alligators, and mammals. The studies discussed here were selected because they are the ones for which the committee found the strongest evidence for the potential of HAAs to disrupt community and ecosystem functioning.
Fish are good indicators of contamination for several reasons. Runoff and erosion carry chemicals from land to water, so aquatic organisms are exposed to all the contaminants in their drainage basin. Because fish are food for wildlife and humans, they are monitored for concentrations of toxic chemicals. Fish thus serve as good indicators of toxic substances in the aquatic environment. Because it is a long-lived top predator, the lake trout (Salvelinus namayculsh) has been suggested as an indicator of ecosystem functioning (Edwards et al. 1990).
Great Lakes Fish
Observations of fish from the Great Lakes have been instrumental in shaping much of what is understood of the potential for persistent, organic pollutants to accumulate in and have adverse health effects on wildlife. The lessons learned have been of much broader use than only for the lakes themselves (Gilbertson 1992). The Great Lakes are typical of many ecologic systems in that they have been subjected to a great deal of human incursion, including commercial and sport fishing (often overfishing); the introduction of native and alien species, such as forage fishes, Pacific salmon (Oncorhvnchlus spp.), and sea lampreys (Petromyzon marinus). and various genetic strains of those species (Strittholt et al. 1988): habitat loss; and chemical pollution.
For nearly 2 centuries the Great Lakes have received industrial and municipal wastewater. The effects of that contamination were not suspected until population effects were observed in the top predators of the ecosystem-birds (Giesy et al. 1994a). It was not until the late 1960s that analytic methods (gas-liquidcontinue
chromatography) were available to measure the presence of synthetic halogenated hydrocarbons (Reinert 1970). Even so, it was not yet understood that small concentrations (parts per trillion or parts per quadrillion) of the compounds could harm the Great Lakes biota (Thomann and Connolly 1984) or that the chemicals could persist, bioaccumulate, and biomagnify. Since the 1960s, many sources of major contaminants, such as polychlorinated biphenyls (PCBs) and 2,3,7,8-tetrachlorodibenzo-p-dioxin-equivalent (TCDD-EQ) have been eliminated or greatly curtailed. and much of the Great Lakes ecosystem is recovering (Makarewicz and Bertram 1991). However, contaminant residues, especially from petroleum hydrocarbons and metals, remain in sediment and could continue to harm fish in some locations (Fitchko 1986). For example, the chinook salmon (Oncorhylchus tshawytschal) population in Lake Michigan crashed in 1989 as the result of an outbreak of bacterial kidney disease (BKD), and populations have not yet recovered to their former size-despite an increased rate of stocking (Garling et al. 1995). The bacteria that cause this disease are normally present in the kidneys of the salmon, but they generally do not harm the fish. It is only when the fish are subjected to additional stress that BKD is expressed to the point of affecting survival and growth. PCB concentrations in chinook salmon (Williams and Giesy 1992: Williams et al. 1992) are lower than those found to cause immune suppression in rainbow trout (Oncorhynchus mykiss) (Arkoosh et al. 1994).
Lake trout (Salvelinus namaycush) also have suffered a population decline because of poor reproductive success (FWS 1981; Mac and Seelye 1981 b). Annual rearing mortalities in lake trout fry of as much as 97% were described for hatchery-reared fish between 1978 and 1981 (Mac et al. 1985). Several studies have attributed this to toxic organic residues in eggs (Mac and Seelye 1981a; Seelye and Mac 1981; Mac and Schwartz 1992; Mac et al. 1993). In addition, some of the fish produced fry that developed ''blue sac" disease, an edematous condition that results in fluid filling the yolk sac, causing a bluish color, and leading to death of the egg.
In the early 1980s, PCBs and dichlorodiphenyltrichloroethane (DDT) were believed to be responsible for the effects. However, efforts to correlate the degree of fry mortality in lake trout with relatively high concentrations of PCBs (11 mg/kg bw) and DDT (7 mg/kg bw) were unsuccessful (FWS 1981; Seelye and Mac 1981). Subsequent studies identified the total concentrations of TCDDEQ to be a more plausible causative agent (Symula et al. 1990; Walker and Peterson 1994).
The failure of recruitment of the lake trout populations could be the result of several factors-behavior, predation, genetics. changes in diet caused by changes in prey species composition, or inappropriate stocking practicesbut the evidence supports the hypothesis that toxic chemicals in eggs were, at least historically, responsible for some of the population decline (Walker and Peterson 1991; Walker et al. 1991). Current concentrations of TCDD-EQ in lake trout fry are near the threshold for mortality (Walker and Peterson 1994). Thus, it is likelycontinue
that the concentrations of TCDD-EQ in lake trout eggs were well above the threshold in the recent past, and that the current concentrations might continue to exert adverse effects on survival.
The effects of contaminants on fish populations of the Great Lakes are difficult to demonstrate because of the complex interactions among species and their environment. For example, during the 1970s and 1980s, populations of alewife (Alosa pseudoharengus) also were declining because of predation (Dunstall 1984) and die-off (Brown 1984). Salmon switched to other prey, such as bloaters (Coregonus hoyi) and smelt (Osmerus mordox) (Elrod 1983: Swedberg and Peck 1984; Muth and Busch 1989; Miller and Holey 1992). Thus, changes in reproductive performance and population size in salmonids might result from complex changes in predator-prey relationships instead of or in addition to the effects of exposure to halogenated hydrocarbons. The fish have been exposed to several compounds at once, so it has been difficult to correlate exposure to any one compound with the observed effects. Many contaminants have been identified as hormonally active, but it is not known whether hormonal or other endocrine disrupting actions promote all of the observed health and population-level effects. However, the data indicate that environmental contaminants can affect fish, especially the reproductive capacity (see Chapter 5). of fish species in the Great Lakes and probably have affected their populations.
M-74 Syndrome in the Baltic Sea
A syndrome analogous to the reproductive impairment observed in Great Lakes salmon has been observed in Baltic Sea fish. The number and fertility of eggs are normal, but the fry become listless and settle to the bottom and die. This phenomenon has been seen in Atlantic salmon (Salmo salar) and brown trout (Salmo trutta) in the Baltic Sea, and the condition has been called M-74 syndrome because it was first reported in these two species in the Baltic Sea in 1974. The M is an abbreviation of the Swedish word for environment, miljö. The effect is observed in wild and feral fish captured from the Baltic Sea for spawning in hatcheries. The intensity of the phenomenon varies among individuals and from one year to another (Norrgren et al. 1993). Several causes of the syndrome have been suggested, including genetics and nutrition, but most researchers believe the phenomenon was attributable to exposure to toxicants, particularly the dioxinlike polychlorinated diaromatic hydrocarbon (PCDH) (Norrgren et al. 1993). The syndrome has not been observed every year or to the same extent from one occurrence to the next, and tagged individuals have produced fry that exhibit the syndrome one year but not in others. Some scientists postulate that M-74 syndrome was not caused by exposure to contaminants: others believe it might be a combination of effects, due in part to contaminants, but that some accessory factor might be necessary for the phenomenon to be observed.break
Several researchers have investigated the concentrations of contaminants in herring oil (Falandysz 1986; Cooper et al. 1991) and the effects on herring (Clupea harengus harengus) (Hansen et al. 1985) and on salmonids fed oil extracted from the herring (Andersson et al. 1993). They demonstrated that there were contaminants in the herring oil, but they could not cause M-74 syndrome in the fish fed the herring oil. Several researchers surmised that a nutrient deficiency caused by the change in diet from a mixture that included shrimp to an almost exclusive herring diet was responsible for the reproductive effects (Norrgren et al. 1993).
Adult fish exhibiting M-74 syndrome are characterized by flesh that is white or grey rather than the normal pink or orange. The eggs produced from those adults are less colored. Some of the researchers suggested that a carotenoid deficiency was responsible (Andersson et al. 1993: Norrgren et al. 1993). This was supported by the color of the adult flesh and eggs, which, when analyzed for carotenoid pigments, were found to be deficient in astazanthine. That discovery led to several studies (Andersson et al. 1993; Norrgren et al. 1993) in which the diet of adult salmon was augmented with astazanthine or eggs were dipped in a mixture containing astazanthine. The treatments increased the carotenoid content of the eggs. but they did not prevent M-74 syndrome.
A great deal of research into the diet of the salmon as well as general chemical and physical conditions in the Baltic Sea has been conducted (Holm et al. 1993: Henriksson et al. 1996). Those studies show that, depending on the amount of fresh water entering the Baltic Sea, the degree of oxygen saturation in the benthic water varied from year to year. In fact, in some years, large areas of the benthic Baltic Sea were anaerobic. When this occurred, the shrimp (Crangon crangon) population was much reduced. It was thought that shrimp could be an important prey item for Atlantic salmon and brown trout. It was postulated that, in the years when shrimp populations were depressed, the two fish that exhibited M-74 switched their diet to one of mostly herring without consuming any shrimp. However, although shrimp make up a significant proportion of the diet of Atlantic salmon in the Atlantic Ocean, they are not a significant proportion of the diet of salmonids in the Baltic Sea (P. Vuorinen, Finnish Game and Fisheries Research Institute, Helsinki, Finland, personal communication, 1997). Thus, it does not seem likely that this is an alternative to the hypothesis that M-74 is caused by contaminants.
Working independently on salmon populations in the Finger Lakes of New York, Fisher et al. (1994) observed a syndrome in salmon (Salmo salar) that was similar to M- 74. They named the syndrome "Cayuga Syndrome" for the lake in which it was observed. A number of instrumental analyses were conducted to determine the concentrations of chlorinated insecticides, PCBs. and dioxin equivalents in the lake, but none of the compounds seemed to be occurring at sufficiently high concentrations to cause the syndrome. However, the investigators were ablecontinue
to demonstrate that the syndrome was caused by the presence of thiaminase (an enzyme that degrades thiamine) in the herring of the salmon's diet. When the fry from adults exhibiting the syndrome were fed diets supplemented with thiamine, the syndrome was eliminated. Thus, the investigators concluded that xenobiotics were not involved in the syndrome, but rather it was a thiamine deficiency caused by the consumption of herring in the diet. However, it should be noted that interactions between PCBs, TCDD, and thiamine metabolism have been reported in laboratory studies with rats (Yagi 1979; Yagi et al. 1979: Pelissier et al. 1992).
Birds are perhaps the most studied group of organisms with respect to environmental contamination because they are conspicuous and relatively easy to observe and collect; they were among the first groups in which the effects of environmental pollutants were observed. Most of the effects observed have been on piscivorus (fish-eating) birds, which are exposed to pollutants in contaminated fish. Two significant population effects have been changes in sex ratios in gull populations and population declines among gulls, cormorants, and terns.
Skewed Sex Ratios
The effects of exposure to environmental pollutants have been demonstrated in studies of skewed sex ratios in gulls exposed to organochlorines. such as DDT. PCBs. and TCDD. Skewed sex ratios in favor of females were documented by the increased number of nests containing an abnormally large number of eggs. or supernormal clutches. Nests containing five or more eggs are considered to be the result of female-female pairing, because a single female gull normally lays one to three eggs. It is believed that females associate with one another when males are unavailable because reproductive success relies on the presence of a pair of nest attendants. One adult must remain at the nest to guard chicks from predation while the other forages for food. It is also possible that females associate with one another to avoid aggression directed toward unmated birds (Shugart et al. 1988). Some studies have found that the reproductive success of supernormal clutches was less than normal (Schreiber 1970; Hunt and Hunt 1973). but other researchers have documented successfully fledged young from supernormal clutches (Kovacs and Ryder 1983).
The most dramatic and best-documented example of a skewed sex ratio occurred between 1968 and 1978 in the western gull (Larus occidentalis) population on Santa Barbara Island in California (Hunt et al. 1980). The adult sex ratio in that population was measured by laparotomy of 856 captured birds to be 0.26 males to females. The investigators also calculated the male-to-female ratio by estimating the number of nests on the island (896), the number of nonbreeding birds (200), and, based upon the number of nests with more than three eggs, thecontinue
percentage of female-female pairs (15%). Using those estimates, the male to female ratio was 0.67. Because many birds laid fewer than normal numbers of eggs in 1978, the investigators believe that the estimate for female-female pairs might have been too low. Therefore, the real ratio of males to females was probably between 0.26 and 0.67.
A supernormal clutch incidence of 0.6% to 1% also was documented in northeastern Lake Michigan herring gulls (L. argentatus) between 1978 and 1981 (Shugart 1980; Fitch and Shugart 1983). Both the California population and the Great Lakes populations of gulls were exposed to large concentrations of organochlorineparticularly DDTcontamination from the 1950s to the 1970s (Fry and Toone 1981). Historical studies of supernormal clutches in gulls to determine whether incidences have actually changed between the pre- and post-DDT eras indicate that the number of supernormal clutches has decreased significantly for many species of tern throughout the United States (Conover 1984b). However, three gull species have experienced a significant increase in supernormal clutches since the 1950s: western gulls, herring gulls nesting in the Great Lakes, and Caspian terns (Sterna caspia) breeding in the United States.
The decrease in the number of males in these populations could be attributable to several factors, some related to persistent organochlorine contaminants. For example, differential mortality between male and female gulls could occur because of differences in body burden of the toxicants. Male western gulls weigh about 25% more than females and they feed higher on the food chain (Pierotti 1981), which makes them more likely to accumulate pollutants throughout their lifetimes. It also has been suggested that estrogenic contaminants could be causing feminization of male embryos with resultant chemical sterilization and failed recruitment into the breeding population (Fry et al. 1987). However, behavioral studies of western gulls (Hunt et al. 1984) and attempts to correlate gonadal feminization with organochlorine contamination in a glaucous-winged gull (L. glaucescens) population in Puget Sound, Washington (Fry et al. 1987), have been inconclusive.
Thus, although environmental contamination has been correlated with the skewed sex ratio observed in several North American gull populations, it is still not known how pollutants cause the differential effects. The explanations presented above are just a few ways in which pollutants could affect populations. and there is no indication that effects are caused by a hormonal mechanism. There might be alternative causes that do not involve environmental contaminants.
Several species of fish-eating birds of the Great Lakes region have experienced population declines, which have been attributed to exposure to environmental contaminants (Keith 1966; Anderson and Hickey 1976; Fox et al.continue
1991a,b). Keith (1966) found that reproductive success in herring gulls in the early 1960s around Lake Michigan was approximately one-third of normal: Gilbertson (1974) found that it was about 10% of normal for birds nesting near Lake Ontario. Nesting colonies of Forster's terns (Sterna forsteri) at Lake Michigan experienced hatching success as low as 26%, and no fledgling success for several years (Hoffman et al. 1987). Although the effects observed with the gulls were partly attributable to eggshell thinning (see Chapter 5). 30% of eggs that were incubated to term failed to hatch. Egg-transfer experiments with herring gulls (Peakall and Fox 1987) and Forster's terns (Kubiak et al. 1989). in which eggs were transferred between nests in "clean" and "contaminated" sites, indicated that the declines in hatching success were the result of toxicants in the eggs and abnormal parental behavior. Environmental contaminants have been shown to affect egg and embryo survival (see Chapter 5), but it is not known whether pollutants are the cause of the behavioral changes observed in adults. This is important to consider because changes in parental behavior could have consequences for other aspects of avian survival outside of the breeding season.
Eggshell thinning has contributed to the decline of several bird species in the Great Lakes region and elsewhere. For example, Great Lakes colonies of double-crested cormorants (Phalacrocorax auritus) suffered widespread reproductive losses due to accidental egg breakage caused by eggshell thinning associated with 1,1 -dichloro-2,2-bis(p-chlorophenyl)ethylene (DDE) (Anderson and Hickey 1972), and the populations declined by 80% (Postupalsky 1978). Although there are no studies on community-level effects that result from population effects, it is likely that they were present because cormorants are strictly piscivorus, and lowering their populations by 80% must have had an effect on the fish populations they prey upon (and potentially on fish size and species composition). However, the effects of these declines on the populations of fish that are the prey of cormorants is almost impossible to discern, because so many other factors have changed as well (e.g., introduction of exotic species, stocking of predatory fish, sport and commercial fishing, other effects of pollutants, and lamprey predation).
In Georgian Bay, 95% of cormorant eggs broke or disappeared before incubation was complete (Weseloh et al. 1983). Dissection of nearly full-term eggs indicated that half of the eggs contained chicks with deformities (Ludwig et al. 1996), and the deformities were typical of those caused by exposure to organochlorine compounds (Fox et al. 1991a,b; Gilbertson et al. 1991). Chicks with deformities often die soon after hatching, resulting in reduced fledgling success in the colony, and ultimately, in changes in age ratios and overall population size.
Another notable incident occurred in 1986, when there was a severe flood in the Saginaw River ecosystem that released PCBs from the sediments in that area (Ludwig et al. 1993). This exposure led to the occurrence of deformities in Caspian tern chicks at a frequency 168 times greater in 1987-1989 than was seen in 1962-1987. Increases in deformities and lessened hatching success for a 2 yr period are likely to have effects on local tern population demography for years,continue
but it might not be detectable by simply enumerating the total population because immigration occurs from other colonies. The evidence that there continues to be toxicant-related embryonic death, deformities, and immune dysfunction in chicks strongly suggests that there are still population-level effects in the Caspian terns (Mora et al. 1993), even though there are no changes in the number of individuals. Populations of cormorants also declined substantially (see Table 10-2 for references).break
Cormorant embryos exposed naturally to environmental concentrations of PCBs. polychlorinated dibenzodioxins (PCDDs), and polychlorinated dibenzofurans (PCDFs) have a statistically significant increased incidence of brain (intercerebral) asymmetries correlated with the exposure (Henshel et al. 1997).
Bald eagle (Haliaeetus leucocephalus) populations also declined in the 1950s and early 1960s, primarily because of exposure to DDT and its metabolites (Wiemeyer et al. 1984). Although many populations have increased in much of the Great Lakes region since then, some populations have not recovered. Others have increased because of migration from other areas and not from reproduction (Best et al. 1994: Bowerman et al. 1994). Craniofacial malformation in eagle embryos and chicks is attributable to contaminated eggs derived from the parents' consumption of contaminated fish and closely resembles abnormalities observed in cormorants and terns. However, the effects of exposure to other pollutants. such as mercury (Giesy et al. 1995) cannot be ruled out. Although there is an inverse correlation between contamination with DDE and PCBs in eggs and plasma of nesting eagles and reproductive success (Bowerman et al. 1994, 1995), it is impossible to decide the degree to which each contaminant affects the birds or the degree to which effects are additive.
Populations of some bird species have been monitored in the Great Lakes region for the past half century, and there is clear evidence for population decline. reproductive impairment, or both, in several fish-eating species. There is also convincing evidence for community-level effects: Many of the declines in bird populations seem to be caused by eating of contaminated fish. Although the declines initially were believed to be caused by DDE-induced eggshell thinning and pollutant-induced adult toxicity, the populations still exhibit subtle effects. such as deformities probably caused by dioxin-like PCBs (Giesy et al. 1994a.b). As populations of some species have recovered, a significant incidence of teratogenicity is still observed. In several species, the effects have been associated with putative HAAs, such as coplanar PCBs, DDT and its metabolites, and other organochlorine compounds. Laboratory experiments with birds indicate that several agents can alter reproduction and development (see Chapter 5) and that the toxic effects of exposure to multiple compounds can be additive. However, a range of other factors can cause declines in population numbers, such as human disturbance and exploitation, loss of habitat, inclement weather, predation, disease. and competition for space or food (Burger and Gochfeld 1996). Any or all of those factors could contribute to lowered reproductive success. Nonetheless, it is remarkable that so many species of fish-eating birds in the Great Lakes have experienced the same individual and population problems. Such changes have not been observed so consistently in such a wide geographic area, over such a span of years. in colonial fish-eating birds of other regions.break
The effects of exposure to pollutants on the health and reproductive biology of the American alligator (Alligator mississippiensis) of Lake Apopka, Florida, have been extensively studied (see Chapter 5), and anomalies in genital and gonad morphology and alterations in plasma sex-steroid concentrations have been the reason for declines in its population size in this area over the years. The most significant decline occurred after the Tower Chemical spill into the lake in 1980 (Woodward and Moore 1990), when the pesticide dicofol (contaminated with DDT and its metabolites 1,1-dichloro-2,2-bis(p-chlorophenyl)ethane (DDD), DDE, and chloro-DDT) was accidentally released. There was a rapid, 90% decline in the population of juvenile alligators between 1980 and 1984 (Figure 10-1), and, beginning in 1984, the viability of alligator eggs from this lake decreased significantly (Figure 10-2). Recent data suggest that the juvenile population on Lake Apopka is increasing slowly, and egg viability has continued to vary, with most years exhibiting an average below that observed on other lakes (Rice et al. 1996). It has been suggested that the initial decline in the alligator population was a directcontinue
consequence of the lethal effects of high concentrations of dicofol-DDT (or its metabolites) on hatchlings and juveniles living in the lake. whereas the current reduced population density of those age groups is a result of poor clutch viability (Guillette and Crain 1996).
An analysis of embryonic mortality in alligator eggs suggests that more than 80% of all embryonic mortality takes place during the first month after fertilization of the egg (Masson 1995). There have been reports of greater concentrations of p,p'-DDE in alligator eggs collected between 1984 and 1985 from Lake Apopka compared with eggs from two other lakes (Heinz et al. 1991), but greater concentrations of organochlorine compounds could not be correlated directly with poor egg viability. The mean concentrations of p,p'-DDE observed5.8 ppm weight-to-weight (w/w) (1984: n = 3 eggs; range, 3.4-7.6 ppm) and 3.5 ppm w/w (1985: n = 23 eggs; range, 0.89-29 ppm)were above the concentrations known to harm avian eggs and embryos (Cooper 1991). In addition to p,p'-DDE, alligator eggs (n = 23) collected in 1985 from Lake Apopka had detectable concentrations of p,p'-DDD (<1.8 ppm), dieldrin (0.02-1.0 ppm), and cis-chlordane (<0.25 ppm) (Heinz et al. 1991).break
As described in Chapter 5, contaminants of Lake Apopka are believed to be the cause of abnormal gonadal morphology and altered sex-steroid concentrations in juvenile alligators. It has been suggested that the gonads of male and female alligators were permanently modified in ovo. and that normal sexual maturation was unlikely (Guillette et al. 1994). Studies by the U.S. Environmental Protection Agency (EPA) (1994b) have shown that juvenile and hatchling alligators from Lake Apopka have high concentrations of p,p'-DDE, primarily in fat (1.6-8.5 ppm) and in liver tissues (0.013-0.17 ppm). That also has been observed in hatchling alligators exhibiting developmental abnormalities.
Because alligators take 12-15 yr to reach sexual maturity (Ferguson 1985: Joanen and McNease 1989), the community and long-term population effects of changes in the current population size and age ratios in Lake Apopka will not be known for many years. It is likely that a reduced population size, coupled with the gonadal and sex-steroid abnormalities observed in juveniles, will affect the reproductive success of these alligators in the future and could pose significant long-term population and community threats.
Data on the effects of mammalian exposure to environmental contaminants are limited and come mainly from work with ranch mink (Mustela vison). river otter (Lutra lutra), and harbor seal (Phoca vitulina). Cause-and-effect relationships are difficult to establish for any toxic chemical in wildlife because of the many difficulties inherent to field studies. Therefore, subtle effects often must be inferred using epidemiologic criteria in concert with laboratory studies.
In the 1960s and 1970s, mink fur production decreased from ranches that used Great Lakes fish in the diet. Studies of ranch mink fed a diet of carp (Cyprinus carpio) from Saginaw Bay, Michigan, have suggested that population declines in the ranches were caused by impaired reproduction and reduced kit survival caused by PCB and TCDD in the diet (Heaton et al. 1995a), because fur production decreased at the time when pollutant concentrations were highest in the region and subsequently increased slightly in the 1980s, when concentrations of PCBs and other pollutants were lower. It is believed that pollutants could also be the cause of observed population declines in the wild (Allan et al. 1991). For example, there are fewer mink in areas where Great Lakes fishes can be consumed than in areas where they cannot, such as above dams that limit the accessibility to lake-run salmon. Fish in Great Lakes-influenced sections of three Michigan rivers were shown to contain greater concentrations of PCBs and TCDD-EQ than did fish that did not have access to the Great Lakes (Giesy et al. 1994b).break
Population declines in mink also have been reported in Georgia, North Carolina. and South Carolina (Osowski et al. 1995). A range of contaminants from the tissues of mink were analyzed, and concentrations of PCBs (0.216 µg/g) and mercury (3.5 µg/g) in liver tissue were in the range known to cause reproductive impairment, growth deficit, and behavioral impairment. Similarly, Foley et al. (1988) reported that PCB concentrations in wild mink from New York were in the range known from laboratory studies to cause reproductive damage (Platonow and Karstad 1973). A significant correlation between body burden in mink with body burden in fish collected from their home ranges (Foley et al. 1988) indicates that the food chain is being affected. These data, coupled with the evidence that exposure to PCBs impairs mink reproduction, reduces kit survival, and lowers body weights of kits (Platanow and Karstad 1973; Heaton et al. 1995 a,b), suggest that environmental pollutants have affected mink populations, which could pose community threats.
The river otter populations in Wales, England, and other regions in Europe declined in the 1950s, and home ranges were constricted (Mason and Macdonald 1986). The changes have been attributed to contaminants in the food chain, because pollution was great during that time, and as the contaminant concentrations in these areas have declined, the otter populations have begun to increase (Mason and Macdonald 1993a). However, recolonization (movement back into the areas) has been slower than expected, probably because of continued exposure to large concentrations of dieldrin, DDE, and PCBs (Mason and Macdonald 1993a, 1994). Using PCB concentrations in feces as an indication of contaminant concentrations in tissues, Mason and Macdonald (1993b) related dietary intake of PCBs to tissue concentrations, and related tissue concentrations to those that attend reproductive problems in otter. They concluded that exposure to PCBs was responsible for the slow recolonization rates in England and Wales (Mason and Macdonald 1993 a,b).
In the Netherlands, the otter populations have become nearly extinct. In 1989, the Dutch government adopted a policy to restore their populations, and a project entitled ''Development of Otter-based Quality Objectives for PCBs" was commissioned in 1993 to support this effort. The two-part project involved a critical review the field data from European otter-habitats (Smit et al. 1994) and an assessment of biomagnification of sediment-bound PCBs in otter habitats, biologic factors determining the bioaccumulation of PCBs in otters, induction of dose-dependent physiologic effects in material from captive and feral otters from different European countries, the health status of otters in Denmark in relation to PCB exposure, and the feasibility of monitoring of PCB exposure in feral otters (Smit et al. 1996). The project reported the following observations: significant biomagnification in the food chain of the otter in the Limfjord area in Denmark;continue
increased incidence of disease (viral infections, bacterial disease, pathologic deviations, and endoparasites) in otter populations, which was correlated with PCB-induced vitamin A deficiency: and current environmental PCB concentrations that are great enough to cause adverse effects (Smit et al. 1996).
Many marine mammals have undergone population declines in the past 40 yr. In some cases, the causes are well known (e.g., hunting of whales and some pinniped and collisions with boats and loss of habitat for manatees), but in others, such as the Steller sea lion (Eumetopias jubatas) in the Bering Sea, serious population declines are not fully understood (NRC 1996b). Another example is the decline observed in the beluga whale (Delphinapterus leucas) population of the St. Lawrence, which was estimated in the 1980s to be only 10% of the 1885 level (Reeves and Mitchell 1984) and to be lower than reference populations in the less contaminated Arctic (Sergeant 1986). Despite protection since 1980, the beluga whale populations have not recovered, due mainly to low calf production and low survival of young (Colborn and Smolen 1996). Low calf production and survival have been hypothesized to be due to exposure to HAAs (De Guise et al. 1995). Elevated concentrations of organochlorines, such as PCBs and DDT- and toxaphene-related compounds (Martineau et al. 1987; Muir et al. 1990 a,b), have been found in the blubber of beluga whales since the early 1980s. Moreover. such organochlorines have been associated with a variety dysfunctions, which might provide an explanation for population declines.
Chapter 7 of this volume describes a correlation between feeding on PCB-contaminated fish and impaired immune response in seals and whales. It is believed that this could account for the drastic declines in the common harbor seal in the Wadden Sea off the north coast of the Netherlands and elsewhere in Europe (Reijnders 1981: Osterhaus and Vedder 1988: de Swart et al. 1996). These seals had high concentrations of PCBs in their tissues and experienced immuno-suppression and subsequent viral infections. Similarly, a disease complex involving a primary lesion of the adrenal glands and secondary reactions in other organs has been observed in the Baltic grey seal (Halichoerus grypus) and the ringed seal (P. hispida) in the Baltic Sea and along Swedish west coast (Olsson et al. 1994). This complex is believed to be the cause of the dramatic decrease in these species in the 1960s and 1970s. The harbor seal population in these areas also suffered about a 60% decrease in 1988 because of epizootic disease caused by phocine distemper virus (Olsson et al. 1994). Historical skull bone material from these seals indicates the presence of unnatural stress factors associated with epizootic diseases. The tissue from the seals was contaminated with a variety of metal and nonmetal elements, and it has been suggested that these contaminants, particularly DDE and PCB methyl sulfones, might be the cause of the population declines (Olsson et al. 1994).break
In a review of marine mammals, Colborn and Smolen (1996) reported that 13 species of cetacean and pinniped experienced recent population declines or die-off. Reproductive and endocrine impairments were reported for eight of the species, and immune problems were seen in four of them. However, such impairments are as likely to be the result of the conditions that caused the deaths as to be the cause of them, and, in general, we do not have good information on the causes of the population declines and die-offs.
Summary and Conclusions
Analysis of available data shows that environmental contamination with known HAAs has affected wildlife populations and, in some cases, communities. There is evidence that certain synthetic, persistent, bioaccumulative hydrocarbons have caused effects on wildlife reproduction, but the mechanism of action of the HAAs is generally not understood well enough to determine whether they act through hormone receptors or through other pathways. Ecologic problems are inherently complex. However, the potential damage to ecosystems and their components from HAAs is too severe to ignore. What is known is that certain chemicals that are released into the environment can cause population-level effects on wildlife. It is at least possible that changes in bird populations in the Great Lakes region (and perhaps elsewhere) and the declines of some seal populations in the North Sea are due in part to the hormonal activity of pollutants. although the evidence is not conclusive. The evidence that implicates HAAs in population declines of lake trout in the Great Lakes is less strong, but it still seems quite likely that HAAs have been among the several factors in those declines.
Long-term studies of populations subjected to HAA exposures are needed to assess the effects of these chemicals in altering population size, age structure, and dynamics. Observational and experimental studies of the linkages between chemical exposures and alterations of key aspects of life histories (e.g., fecundity. survivorship, longevity, and age of first reproduction) should be undertaken to understand how chemical exposures affect long-term ecologic attributes of natural systems. Ultimately, the physiologic and biochemical bases of these linkages. once established, should be determined.break