Endocrine Disruption in Wildlife

The various EDCs differ greatly in their potencies relative to natural hormones and in their affinity for target receptors. Classification of EDCs has been performed according to their known or suspected activity in relation to hormone receptors and pathways. While most attention has focused on EDCs that are mediated through the ERs (estrogen agonists) and affect development and reproductive functions in wildlife and humans, numerous laboratory and field studies show that many EDCs can also target the androgen receptors (ARs), the thyroid hormone receptors (THRs), glucocorticoid receptors (GCRs), progesterone receptors (PRs), aryl hydrocarbon receptor (AhR) and retinoid X receptor (RXR) and other signaling pathways [1].

In addition to nonylphenol (NP), a wide variety of chemical compounds are known to act as ER agonists, including:

The pesticides methoxychlor, aldrin, dieldrin, certain polychlorinated biphenyls (PCBs), bisphenol A (BPA; a high-production-volume chemical used to make polycarbonate plastic)
Pharmaceutical estrogens, such as diethylstilbestrol (DES) and ethinyl estradiol (EE2; a major active component in birth control pill)
Natural steroid hormones excreted by humans and livestock (estradiol [E2], estrone [E], etc.)
Phyto-estrogens (which occur naturally in many plants, most notably in soybeans in the form of genistein and related substances)
A number of chemical mixtures (reviewed by [51])

There are a few known ER antagonists, or antiestrogens, including certain OH-PCBs [52]. AR antagonists comprise chemical compounds such as vinclozolin, procymidone, linuron, fenitrothion and chlorinated pesticides such as p,p′-DDE and lindane as well as some of the phthalate plasticizers (a group of chemicals used to soften polyvinyl chloride plastics), phytosterols (present in pulp mill effluents), and certain PCBs (reviewed by [53,54]). Polycyclic aromatic hydrocarbons (PAHs) are suspected of having a range of weak ED effects (depending on structure) via mediation through ER, AR, and Ah receptors (reviewed by [18]). Chemicals such as PCBs, perchlorates, and BFRs are AhR agonists and characteristic disruptors of thyroid hormone homeostasis [55]. The BFRs polybrominated diphenyl ethers (PBDEs) are known also to disrupt thyroid hormone transport and metabolism [56].

Thus, in addition to the reproductive system, many other different receptors and tissues can be affected by EDCs, including endocrine glands such as pituitary, thyroid, thymus, and adrenal and a number of other endocrine-mediated physiological systems, including the immune and neurological systems, although underlying mechanisms are poorly understood. Increasing evidence from laboratory and field studies demonstrates that the neuroendocrine stress response is a sensitive target for disruption by a range of environmental contaminants, at a number of discrete loci. For example, it has been established that interrenal dysfunction, involving an impairment of the secretion of corticosteroid hormones such as cortisol, can be caused in wild fish and other vertebrates by chronic exposure to a range of organic and inorganic pollutants, including heavy metals, PAHs, and PCBs (see [57,58]). Until now, however, relatively few studies have investigated links between endocrine disruptors and stress hypothalamo-pituitary-interrenal/adrenal (HPI/HPA) axis. Corticosteroid hormones in vertebrates are critical for metabolism, growth, reproduction, immunity, and ion homeostasis, and are an important part of the coping mechanisms involved in the stress responses [59]. Furthermore, chemical activation of the HPA axis by PCBs or through interactions with the GCRs [60] can have adverse effects on a number of different systems, thereby expanding the number of potential targets for EDCs [10]. The underlying mechanisms of the neurocrine stress response and how precisely this affects the fitness of the individual (via reduced growth, immunosuppresion and reproductive failure, etc.) and potentially the population level is not well understood (reviewed by [57]).

Although there is considerable information on the early molecular events involved in hormone response, there is very little knowledge concerning the relationship between those molecular events and adverse health effects such as reproductive toxicity, behavior, and cancer. Immune function, long known to be sensitive to steroids, has also been identified as an EDC target [61]. Examples of chemicals interfering with immune function via endocrine interactions have been described for numerous compounds, including androgens [62], estrogens, organotins, and dioxins [10,61]. EDC exposure may also reduce the production of immune-related proteins in fish, which makes them more susceptible to disease. A recent study demonstrated that a PCB mixture (A1248) modulates both immune function and endocrine physiology in brown bullhead [63]. The results suggest that EDCs may make fish more susceptible to disease by blocking production of hepcidin and other immune-related proteins that help protect fish against disease-causing bacteria, viruses, and parasites [64].

Sex hormones play a critical role in both developmental and adult expression of behavior through actions on the brain. These compounds interact with brain neurochemistry to mediate many social behaviors in vertebrates. Even small deficits in brain function could render the animal less able to escape predation, catch fast-moving prey, attract a mate, and rear offspring. A rapidly increasing body of scientific research is revealing that a large number of EDCs (e.g., dichlorodiphenyltrichloroethane (DDT), PCBs, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), tributyltin [TBT], BPA, methylmercury, many pesticides) can have profound effects on neuroendocrine-regulated behavior in wildlife (and humans), and suggests that such altered behavior may detrimentally affect survival (reviewed by [65]). Recent examples include altered reproductive behavior of the male three-spined stickleback (Gasterosteus aculeatus) caused by the organophosphorous pesticide fenitrothion [66] and impaired courtship and aggressive behavior of the male sand goby (Pomatoschistus minutus) after exposure to EE2 [67]; for more examples, see Section 2.3.

It is important to note that, for some EDCs, the parent compound may have no ED activity whereas the metabolites of the same chemical may be biologically active. For example, methylsulfonyl(MeSO2)-PCB metabolites (which are persistent and bioaccumulative contaminants) are potentally antiestrogenic [68]. Another example is 17β-trenbolone (a synthetic androgen), which is a metabolite of a steroidal growth promoter administered to cattle that can cause reproductive effects in fish [69]. And there are other types of EDCs that affect particular endocrine targets (see also Table 2.1). The highly complex nature of hormonal systems and the many points at which disruption can occur is well demonstrated by a recent in vitro toxicity testing of 27 individual BFRs [70]. The results revealed a scala of ED potencies, some of which had not or had only marginally been described before (AR antagonism, PR antagonism, estradiol sulfotransferase [E2SULT] inhibition, and potentiation of T3-mediated effects). For some BFRs, the potency to induce AR antagonism, E2SULT inhibition, and transthyretin receptor (TTR) competition was higher than for natural ligands or clinical drugs used as positive controls. A number of these BFRs have shown ED effects in vivo, including altered thyroid hormone homeostasis and effects on neurodevelopment [71].

Low-dose effects are commonly associated with the endocrine system. These low-dose effects refer to biological changes that may occur at much lower doses than would normally be expected to have an effect or at doses insensitive to traditional testing methods. A wide range of adverse effects has been reported in experimental animals exposed to low doses of BPA exposed both during development and in adulthood. These findings have been related to the potential involvement of EDCs in a range of human disease processes, such as the increase in prostate and breast cancer, hypospadias a decline in semen quality in men, and various metabolic and neurological disorders [72–77]. The available evidence in the case of BPA also illustrates the issue of a very long latency for effects that may not become apparent until long after EDC exposure during development has occurred. These developmental effects may be irreversible and can occur due to low-dose exposure during brief sensitive periods in development, even though no BPA may be detected when the damage or disease is expressed [75].

2.1.7 Endocrine Disrupters in the Environment

Both natural and synthetic environmental EDCs enter into the different environmental compartments (atmosphere, freshwaters, seawater, soils, and marine sediments) through active application, industrial and domestic waste water discharges, incineration, and/or livestock runoff. The aquatic environment may act as a sink for many contaminants that originate from wastewater, air deposits, runoff, and other sources and could therefore pose a high risk to aquatic organisms and fish-eating top predators. Common EDCs in domestic and some industrial effluents and their receiving surface waters are estrogenic hormones (E2 and E), the synthetic EE2, and other pharmaceuticals including glucocorticosteriods (GCs) and personal care products [51,78,79]. The majority of EDCs are POPs and other bioaccumulating chemicals. They are distributed around the globe through atmospheric transport and can contaminate areas far removed from the original site of contamination. Persistent synthetic EDCs have been detected in all environmental media, although concentrations of some legacy compounds, such as PCBs, DDT, and TBT, have declined markedly in some regions, because they are no longer produced or used in those countries. Many of these chemicals exist within complex mixtures (wastewater effluents) and are mobile in water.

The relative importance of direct uptake from the water and uptake from food will depend on the characteristics of the chemical. If the chemical is persistent, and particularly if it is also lipophilic, food chain effects can be expected to predominate as they can move through food chains and represent a threat to top predators in the aquatic and terrestrial ecosystem, such as birds and mammals. For most aquatic organisms, hydrophilic chemicals are readily taken up via the gills, digestive tract, and skin. Furthermore, the eggs of most aquatic animals are deposited into water, and thus the developing embryos may be directly exposed to EDCs and other toxicants at susceptible stages in their development. Benthic invertebrates may be exposed to EDCs through direct contact and ingestion of sediment/soil particles and pore water or by eating contaminated food. For other terrestrial (land-living) wildlife, the major route of exposure is via the diet [80].

Indeed, significant concentrations of legacy and newly emerging POPs with known or suspected ED proporties and other EDCs are increasingly reported in especially (aquatic) top predators at locations remote from human activity (such as the Pacific Ocean and the Arctic) and might perhaps pose the most serious threat of EDCs to wildlife populations and biodiversity (reviewed by [81,82]). Recent reports that document these threats from putative or known EDCs to marine top predators, such as polar bears, sperm whale, dolphins, tuna, albatross and other bird species, include organochlorines [82,83], TBT [84,85], toxaphenes [86], PBDEs [86–89], perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) [90–93], and phthalate esters [94].

On basis of the information of the many facets of endocrine disruption in wildlife (and humans) just presented, EDCs in the environment are a matter of great concern because:

Hormonal systems can be disrupted by numerous different anthropogenic chemicals through various mechanisms.
EDCs may be effective at very low concentrations, and their effects depend on timing of exposure as well as level of exposure.
Disruption by EDCs may have widespread implications for many organisms, as hormonal regulation of biological functions is common to both vertebrates and invertebrates.
The ubiquity of exposure: both naturally occurring and man-made hydrophilic and bioaccumulative substances can be EDs and are widely distributed over the globe.
Mixtures can produce additive responses at individually negligible concentrations.
The persistence of effects: the effects of exposure to EDCs can be observed long after the actual exposure has ceased. For this reason, the effects may be delayed and subtle and often difficult to discern before they are discovered in the field.
Beside reproductive effects, there is scope for a wide range of potential effects on the immune and neurological systems as well as epigenetic effects.
Our knowledge about the precise endocrine mechanisms and how EDCs interact with and influence endocrine, immune, neural, and other systems is still fragmentary.

2.2 EFFECTS OF EDCS ON WILDLIFE

Here we present an overview of the documented reproductive and nonreproductive effects of EDCs in different vertebrate and invertebrate taxa. We focus on recent literature and new findings, and describe commonalities in observed effects among various wildlife taxa.

2.2.1 Mammals

Numerous field, and some (semi-)experimental studies have shown that aquatic mammals, particularly those high in the food chain, such as pinnipeds, odontocete cetaceans, otters, and polar bears, may accumulate high levels of some EDCs in their fatty tissue, in particular organochlorines and other POPs, and are sensitive to the toxicological effects of these EDCs [7,95].

Although it is generally accepted that persistent contaminants (PCBs, DDT) have played a major role in the population declines of seals populations in northwest Europe in the second half of the twentieth century, the evidence is not conclusive. Field studies on Baltic gray (Halichoerus grypus) and ringed seals (Phoca hispida baltica) and semi-field studies on harbor seals (Phoca vitulina) by Dutch research groups have attributed impairment of both reproduction and immune function to organic contaminant (OC) contamination, notably PCBs, in the food chain. Reproductive effects resulted in population declines, whereas suppression of immune function likely contributed to the mass mortalities due to morbillivirus infections. These historic cases have been described and evaluated at length elsewhere (reviewed by [1,7]). In brief, Baltic seals exhibited sterility and a suite of reproductive and nonreproductive disorders including skull lesions, uterine stenosis, occlusions, uterine smooth muscle tumors (leiomyomas), and adrenocortical hyperplasia (Baltic seal disease syndrome). This syndrome has been associated with high levels of PCB and DDT and their metabolites, notably PCB- and 2,2-bis(p-chlorophenyl)-1,1-dichloroethylene (DDE)-methyl solfones and 1,1-dichloro-2,2-bis(p-chlorophenyl)ethane (DDD), which are known to affect the function of the HPG axis and adrenal axes [1,7]. Several of these EDCs are capable of disrupting both glucocorticosteroid hormone synthesis and receptor-mediated action (reviewed by [7]). The available evidence supports an etiological role of PCB for parts of the disease syndrome, particularly for lesions connected to the reproductive failure among the seals but also involvement of an ED component of physiological stress (reviewed by [7]). In a two-year feeding experiment, female harbor seals fed fish from the polluted Wadden Sea displayed a lower reproductive success than seals fed less contaminated fish from the Atlantic Ocean [96]. Reduced levels of E2, retinol, and thyroid hormones in plasma were found in the group with the highest PCB uptake [96–98]. In the same study, implantation failure was found to be associated with lower levels of E2 [99]. Plausible explanations for the observed effects include: PCB-induced reduction in E2 levels due to alterations in enzyme metabolism and interference by PCB or DDE and their metabolites with receptors in target tissues [7]. However despite the strong correlation of the various reproductive effects in seals with OC exposure, there is still an incomplete understanding of the specific compounds responsible for the reproductive and pathological effects and their mechanism of action(s) [1]. In a second long-term feeding study, it was shown that ambient levels of environmental contaminants, notably PCBs, are immunotoxic to harbor seals [100]. Various immune function parameters were suppressed in the contaminant-fed group including natural killer (NK) cell activity and proliferative lymphocyte responses after stimulation, suggesting an impaired T cell function. These functions are known to be important in the clearance of virus infections [7]. The results obtained in captive seals fed contaminated fish are consistent with the effects observed in laboratory animals exposed to Ah-receptor binding PCBs, polychlorinated dibenzodioxin (PCDDs), and polychlorinated dibenzofuran (PCDFs) [7]. The results of these studies make it likely that chemical-induced immunosuppression could have contributed to the virus-associated mass mortalities among seals inhabiting northwestern Europe in 1988–1989 [7] and possibly also in 2002. Another group of authors found indications for a link between thyroid hormone levels and exposure to PBDEs in gray seals from the United Kingdom during their first year of life [101]. The molecular mechanisms involved in contaminant immunosuppression and susceptibility related to those determining viral susceptibility are, however, still unresolved [102].

There are a several cases in which recent declines in endangered mammalian populations are unexplained and where exposure to environmental EDCs may be involved. Examples are California sea otter declines in the United States and Steller sea lion (Eumetopias jubatus) declines in Alaska [103]. Kannan and coworkers [104] found PCBs, DDTs, and butyltins to be major contaminants in sea otters and their prey collected from California (United States). They noted that California sea otters suffer from various fatal infections and showed that the 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) equivalents of non- and mono-ortho PCBs in sea otters and certain prey species were at or above the theoretical threshold for toxic effects, indicating a possible role of EDCs in the disease events. Also, elevated tissue levels of butyltins, mercury, PCBs, DDTs, chlordanes, and hexachlorobenzene have been reported in Alaskan Steller sea lions. However, the impacts of these exposures on the continued decline of this species remain unknown because causal effects have not been established [103].

A number of studies indicate ED-related effects in polar bears (Ursus maritimus). Masculinization (pseudohermaphroditism) has been reported in polar bears from Spitsbergen, Norway [105]. In a heavily contaminated population of polar bears in Svalbard, Norway, hormonal changes have been linked with the concentrations of PCBs and DDEs in the blood. These hormonal changes included higher blood progesterone levels in more heavily contaminated female bears, but similar estrogen levels compared to females with low burden [106]. Further, more heavily contaminated male bears had lower testosterone levels, and bears (regardless of age and sex) with higher contaminant burdens exhibited lower blood levels of thyroid hormones and cortisol [107–109]. Statistically significant correlations between contaminant body burden and hormone concentrations in individual bears strongly suggest a dose-response relationship [110]. Another study reported a significant negative association between high blood levels of PCBs and serum immunoglobulins as well as cell-mediated immunity in polar bears [111]. It is clear that there is convincing evidence for a role for EDCs in reproductive disorders in polar bears, although the possible population impacts are still unknown.

Numerous other cases refer to mass mortalities by infectious diseases, poor reproductive performance, immunosuppression, thyroid abnormalities, and other nonreproductive disorders in a range of marine mammals, including seals, dolphins and whales. Such effects have to some extent been associated with the presence of POPs (e.g., organochlorine compounds, BFRs, and certain metabolites) and other ED and/or immunotoxic compounds in the body fat [112]. An increasing disease susceptibility in different whale and dolphin populations has led to speculation about a possible negative influence of contaminants on the immune system [113]. In most of these cases, however, it was not possible to confirm a cause-and-effect relationship between a specific chemical or group of chemicals and individual or population level effects.

Lesions suggestive of endocrine disruption are also reported in the Beluga whale (Delphinapterus leucas) population from St. Lawrence estuary in Canada. The lesions include unique thyroid lesions (adenomatous hyperplasia), a case of true hermaphroditism, and one case of male pseudohermaphroditism [114,115].

Surprisingly, there is only little evidence for effects of EDCs in freshwater mammals. Toxicity data and levels of PCBs observed in wildlife populations of mink (Mustrela vison) and otters (Lutra canadensis) in the Great Lakes Basin (United States) were reviewed by Wren [116]. The author concluded that there was little doubt that contaminants had effects on mink and otters from the Great Lakes Basin but that available data were insufficient to provide final proof of cause-and-effect linkages between chemical compounds and population status. Data on contaminants in the diet of mink and otter were not available. The European otter (Lutra lutra) population declined dramatically between the 1960s and the 1980s caused by a combination of habitat destruction and exposure to PCBs, leading to impaired reproduction. Non-ortho-substituted PCBs are generally held responsible for adverse reproductive health effects observed in these species [117,118]. To date, most otter populations show signs of recovery due to a ban on PCBs and as well as conservation measures, such as reintroduction programs.

Several studies report reproductive dysfunctions and disorders in endangered terrestrial top predators and suggest associations with EDC exposure. These include cryptorchidism (90 percent of the male population), a high prevalence of sperm abnormalities, and cases of sterility in Florida panther (Puma concolor coryi), possibly due to mercury, p,p-DDE and PCBs. Similar to reports of masculinization in polar bear, features of masculinization in black and brown bears with unknown etiology have been reported. Data on these cases are limited and the etiology of the observed effects remain unresolved (reviewed by [1,7]).

2.2.2 Birds

Birds, like mammals, are primarily exposed to chemical substances through food. Many lipophilic chemicals, such as PCBs, polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), and chlorinated pesticides, are readily taken up and excreted in the yolk of egg-laying birds. This efficient route of elimination from the oviparous female may result in a pronounced exposure of the avian embryo to toxic chemicals from its early stages of development. Convincing evidence for EDC effects on wild bird populations is exemplified but limited to the well-documented case of DDE-induced eggshell thinning that caused severe population declines in a number of raptor species in Europe and North America (see also Section 2.1.2). A well-accepted mechanism for the eggshell thinning is that DDE blocks the cellular signal that stimulates the eggshell gland to deposit calcium in the shell [119,120]. Other reproductive effects have been reported in many species, especially aquatic and terrestrial birds of prey. Although organochlorine compounds clearly cause serious reproductive problems in a large variety of wild bird species (ospreys, gulls, cormorants, herons, terns, etc.), there is little evidence that sexual development is permanently affected (reviewed by [1,7,110,121,122]). Some evidence links PCB, dioxin, and p,p-DDE contamination to thyroid hormone disruption in wild birds, but results differ between species and between life stages [110]. For example, herring gulls with high PCB body burdens seem to suffer hypothyroidism as chicks but not as adults [123]. Herring gulls in the Great Lakes region may also have compromised GC status. Negative correlations were found between plasma levels of T4 and T4:T3 ratio and blood levels of OCs, especially hexachlorobenzene (HCB) and oxychlordane in free-ranging male glaucous gulls (Larus hyperboreus) breeding in the Barents Sea [124]. Higher burdens of PCBs, PCDDs, and PCDFs in the yolk sacs of embryos were significantly linked to lower plasma corticosterone concentrations and suppressed gluconeogenic and lipogenic enzymes [125]. No data directly link any of these sublethal ED effects to fertility or behavior or survival effects that could impact health or populations.

The studies just mentioned do not include the potential effects of new emerging EDCs, such as BFRs on birds. In recent years, a series of toxicological studies on PBDEs in captive American kestrels (Falco sparverius), a small bird of prey, have become available. The findings of these studies showed that: (1) in ovo and developmental exposure of American kestrel nestlings to a technical Penta-BDE, DE-71 and a combination of pure pentabromoBDE congeners can induce changes in thyroid, vitamin A, glutathione homeostasis, oxidative stress, and immunomodulatory changes in kestrel chicks at environmentally relevant concentrations [126–128]; (2) dietary exposure of adult kestrels to environmentally realistic concentrations can induce changes in reproductive courtship behaviors [129]; (3) additionally, the exposure to DE-71 causes decreased pipping and hatching success in kestrels [130] and can also alter kestrel eggshell thickness and affect reproductive success [131]. These findings might have serious implications for reproductive success and adversely affect the health and populations of wild birds.

Two original studies have examined the effects of EDCs on birds, including starlings and tree swallows foraging near wastewater treatment plants (WWTP) installations. A recent U.K. study examined the effects of endocrine disrupters present in macrofauna from sewage filter beds on the starling (Sturnus vulgaris) [132]. The authors identified the various EDCs present in invertebrate prey and assessed the intake rate of birds observed foraging at these sites. The results show that male sparrows fed with worms spiked with a mixture of EDCs (E2, BPA, dioctyl phthalate, dibutyl phthalate), in concentrations similar to those measured in worms in the filter beds, developed longer and more complex songs that make them more attractive for females but at the same time have a lowered immune response. The ecological consequences of these marked effects on behavior, brain, and immune system for the sparrow populations are, however, still unknown. In another study, a Canadian research group showed that tree swallows (Tachycineta bicolor) that feed on insects and forage near WWTP sedimentation ponds have smaller litters and a reduced number of fledglings compared to a reference location [133]. These two studies show that wastewater treatment sites can be a source of EDCs for birds that forage at these locations. However, no clear effects were found on circulating sex steroids and reproductive performance in tree swallows that feed on the insects at downstream sites of pulp mills (including a very complex mixture of natural and anthropogenic compounds with estrogen- and dioxin-like compounds) in western Canada [134]. This suggests that dietary exposures to EDCs in pulp mill effluents at these sites were insufficient to elicit responses. However, a recent report by another research group on tree swallows breeding near a mercury-contaminated river showed altered plasma corticosterone and thyroid hormone concentrations in nestlings that are important for development, metabolism, and coping with stress [135]. These findings indicate that insectivorous terrestrial vertebrates may also be at risk from EDCs.

2.2.3 Reptiles and Amphibians

Reproductive and nonreproductive effects have been observed in aquatic wild amphibians and reptile populations, especially from North America. One well-documented case is that of impaired reproductive function of the American alligators in Lake Apopka in Florida, which was contaminated by municipal runoff and agricultural chemicals as well as pesticides from a major spill containing dicofol and DDT in 1980. Immediately after the spill, a 90 percent decline in the number of juvenile alligators present was observed; female alligators had twice the concentration of plasma estradiol and also exhibited abnormal ovarian morphology, including polyovular follicles. Alligator eggs showed poor viability and poor survival under laboratory conditions. Male alligators exhibited poorly organized testes, abnormally small phalli, and a sharp reduction in plasma testosterone levels [1,136–138]. Many hypotheses have been proposed to explain the ED effects in male and female alligators. It is, however, conceivable that multiple endocrine mechanisms, including estrogen receptor-mediated and non–receptor-mediated mechanisms are involved. Recent laboratory and field studies indicate that alteration of aromatase enzyme activity (resulting in an altered conversion of androgens to estrogens) and also the thyroid-gonad axis may be involved in the reported reproductive effects [138].

Laboratory studies have indicated that DDT and several other estrogen contaminants may be estrogenic in other reptilian and amphibian species. Examples include o,p-DDT, which induced estrogen-mediated hepatic synthesis of the yolk precursor VTG in male red-eared turtles (Trachemys scripta) [139] and environmentally relevant concentrations of atrazine, which induced hermaphroditism at 0.1 parts per billion (ppb) in American leopard frogs (Rana pipiens) [140]. In snapping turtles (Chelydra serpentina), however, only limited evidence for feminizing effects of atrazine on gonadal development and no evidence of altered thyroid morphology has been reported [141]. Mild signs of feminized sex characteristics have been observed in snapping turtle populations from a site heavily contaminated with PCBs and DDT metabolites in the Great Lakes/St. Lawrence River, but this apparently has not led to fertility problems and adverse impacts on the population density of these reptiles [141]. Larval exposure to environmental concentrations of EE2, a potent estrogenic pharmaceutical and environmental pollutant, has been shown to cause female-biased sex ratios in two frog species (Xenopus tropicalis and Rana temporaria) [142].

Amphibians are subject to multiple exposures during different stages in their life cycle and therefore may be particularly at risk. Numerous laboratory and field studies have suggested that agricultural contaminants are associated with amphibian reproductive abnormalities and population declines (reviewed by [143]). Perhaps the best field evidence for a causal relationship between agricultural contaminants (i.e., atrazine and feminization of frogs) is from recent studies by Hayes and coworkers [140]. They conducted a laboratory experiment with leopard frogs and then carried out a field survey across the U.S. Midwest [140]. The laboratory experiment demonstrated a higher prevalence of gonadal abnormalites and hermaphroditism in frog tadpoles exposed to the herbicide atrazine as low as 0.1 ppb. The results show that male leopard frogs were extremely sensitive to atrazine exposure during metamorphosis from tadpole to adult. The field results revealed widespread gonadal abnormalities in regions where atrazine contamination is within the range shown by the laboratory studies to disrupt development. The studies provide circumstantial evidence that the effects observed in wild leopard frog populations are caused by atrazine, but other EDCs and processes may also be involved. Recent field and mesocosm studies have shown that atrazine and two other pesticides (malathion and esfenvalerate) may potentiate parasitic infestations in amphibians (in addition to the impact of these chemicals on reproductive development) [144,145], suggesting that EDC-induced immunosuppression may be implicated.

Further evidence exists to support the hypothesis that EDCs contribute to amphibian population declines. A retrospective analysis of the observations of the northern cricket frog (Acris crepitanis) collected in Illinois (U.S.) from the period 1852 to 2001 showed that the proportion of intersex individuals in these populations and the decline of cricket frogs were congruent [146]. In another field study from the United States, the occurrence of gonadal abnormalties and measures of gonadal function of the toad (Bufo marinus) were examined across a range of sites that vary in degree of agricultural intensity [147]. Associations were found between the severity of the observed intersexuality (due to feminization and demasculinization of male toads) and the maximum number of gonadal abnormalities, on one hand, and agricultural land use typically polluted with pesticides, on the other. These intersex toads also showed reduced blood testosterone concentrations and feminized or demasculinized secondary sexual traits. It should be noted that this study did not perform analysis of chemical exposure levels, and no true reference site with little human impact was included. In another impressive study performed at the Experimental Lakes Area in Canada, the impact of low EE2 was studied on gonad development and hatching success in native amphibians [148]. Low ng/L concentrations of EE2 were selected based on levels commonly found in sewage effluents and occasionally in surface waters [149,150]. Egg masses were reared in situ in the EE2-amended lake and in two reference lakes in 2001 and 2002. Hatching success was reduced significantly in green frogs (Rana clamitans) but not in mink frogs (Rana septentrionalis) exposed to EE2. EE2-exposed mink frog tadpoles showed intersex in the field populations. The effects were strongest in the third year of exposure, so only after prolonged exposure of EE2 to the parents. These studies in wild populations of toads, together with laboratory studies mentioned earlier, indicate that exposure to EE2 can impact gonad development and hatch success in native amphibians. Whether these morphological effects in tadpoles impact toad populations, however, was not further investigated in this study.

Numerous laboratory studies indicate that an array of environmental EDCs may alter HPT axis activity, development, and reproduction in amphibians. For example, environmental contaminants such as perchlorate, NP, methoxychlor, and DDE are known to inhibit normal thyroid activity, while some other contaminants, including the herbicide acetolchlor, appear to enhance thyroid activity (reviewed by [151]). However, at present, there is an absence of field observations of contaminant-induced HPT alterations in wild amphibians and knowledge about if or how they affect organismal and population health [152].

2.2.4 Fish

In vertebrates, gonadal abnormalities and reproductive impairment in fish are among the most commonly reported ED effects. Reproductive disturbances in fish—including reduced fertility, feminization of males, and masculinization of females—as a result of endocrine disruption have been observed in the last few decades in the vicinity of pollution sources (effluents from sewage treatment and paper mills) in Canada, the United States, Europe, and elsewhere [1,7,58,81,153].

Endocrine disruption of reproductive function resulting in masculinization of fish from exposure to effluents from paper and pulp mills has been demonstrated in numerous species from rivers and lakes in the United States, Canada, and Europe [1,54,58,154]. Observed effects include reproductive disturbances in perch (Perca fluviatilis) in the mid-1980s, which were attributed to unknown chlorinated organic chemicals released from chlorine bleaching process (no longer in use). In Canada, extensive research has focused on white sucker (Catostomus commersoni) populations exposed to bleached kraft mill effluents in Ontario. White sucker fish exposed to the effluent from a bleached kraft pulp mill exhibited changes in reproductive development, including delayed sexual maturity, reduced gonadal growth, and alterations in plasma sex steroid hormone levels. Several studies have shown that TCDD and related compounds present in pulp mill effluents exhibit dioxin receptor-mediated antiestrogenic activity (reviewed by [1]). Although the causal chemicals responsible have not yet been determined, there is compelling evidence that chemicals in bleached kraft mill effuents are responsible for changes in endocrine function and reproductive performance of fish (see [1,54]). Studies with caged fish provide evidence of a temporal link between exposure and changes in sex steroid levels. Studies examining wild or caged fish downstream of other pulp mills in Canada and Sweden provide convincing evidence that chemicals in the effluent contribute to adverse reproductive responses and associated reproductive effects. The field findings are supported by laboratory experiments that provide evidence of adverse effects, mediated through the AhR and sex steroid receptors, and altered steroid hormone function. Although not typically seen in wild fish exposed to pulp mill effluents, feminization has been reported from exposures studies of fish to pulp mill effluents (see [54]). In fact, many laboratory exposures of fish to pulp mill effluents have found that signs of masculinization of females and feminization of males occur simultaneously [54,155,156]. This phenomenon could be caused by compounds in the effluents having both androgenic and estrogenic properties [157]. It has been suggested that fish convert some proportion of androgen-related compounds into estrogens via the aromatase enzyme [54]. At present, the concentrations of chemicals demonstrating activity individually in laboratory studies do not completely explain the responses observed in fish, making it difficult to pinpoint the mechanisms involved. A wide variety of chemicals have been identified in association with pulp mill effluents, including PCDDs, PCDFs, alkylphenol ethoxylates (APEOs), phytoestrogens, and retene, but it has been difficult to isolate the causative agents (reviewed by [158]). The potential for population effects of kraft mill effluents seemingly varies among fish species. So far, no data indicate decreased abundance of wild fish populations in effluent receiving areas [54].

Estrogenic chemicals and their feminizing effects of wild fish populations by WWTPs effluents and other effluents have been widely investigated. Initially discovered in roach (Rutilus rutilus) from U.K. river systems receiving sewage effluent outflows [159], subsequent studies revealed additional reproductive disruptions in wild fish populations. These discoveries were based on the use of the egg yolk precursor protein VTG as a sensitive biomarker for estrogen exposure in male fish [160]. These studies prove that many freshwater species are now experiencing estrogenic effects that appear to be primarily related to natural (and, to some extent, synthetic) substances in sewage and other effluents throughout the world [161–168]. Alongside with induction of VTG in males, altered steroid concentrations, high incidence of oocyte atresia, retardation of gonadal development, feminized external sex organs, decreasing fertility, and frequent intersex have been reported [58]. Caged fish and laboratory-based exposures confirm that sewage effluent is responsible for the observed increases in VTG [1,18,58] and reproductive failure [169]. Fewer effects are detectable in female individuals [18,170], and there are differences in susceptibility between species to estrogenic chemicals. For example, in contrast to roach, sexual disruption seems uncommon in pike (Esox lucius), a fish at the top of the food chain predating on fish species such as roach [171].

Although not yet fully understood, there is evidence indicating that wild fish (roach) populations with intersex features (characterized by simultaneous occurrence of male and female germ cells in one gonad) are compromised in their reproductive capacity and thus that the phenomenon has potential consequences for fish populations, at least on a local scale [6,9,172,173]. Studies on wild populations of roach inhabiting U.K. rivers have shown that exposure to estrogenic effluents emanating from WWTPs causes altered sexual development that can result in reduced fertility [172]. The occurrence of intersex individuals in fish populations is affecting their reproductive success and possibly their population stability. Studies assessing gamete quality of wild intersex roach have found moderately to severely feminized male fish that have reduced sperm quality and quantity (on average 50 percent less), and were less able to release their milt compared with normal males [172]. Controlled exposures of roach to estrogenic effluents have demonstrated that early life stages are especially sensitive to feminizing effects [174,175].

The substances causing these effects are likely a range of natural and synthetic steroidal estrogens, such as E1, E2, and EE2 [176,177], with minor contributions from other estrogenic chemicals found in WWTP effluents, such as bisphenols and phthalates, nonylphenols and their ethoxylates, and carboxylates [149,150,178–182]. In addition, equine estrogens used pharmaceutically in hormone replacement therapy [183] and natural estrogens (estradiol and its metabolites estriol and estrone) and other products contained in animal waste [170] might contribute to the feminization observed in wild fish populations in the United Kingdom and elsewhere. However, increasing emissions of new emerging compounds may also affect the action of hormonal pathways other than the estrogenic hormonal axis. Recently (as yet unknown) contaminants with anti-androgenic properties have been detected in sewage effluents [184,185] and been able to induce biological effects in fish [186]. Because estrogens and xenoestrogens are present as mixtures they can be additive in inducing biological effects [187,188]. It was suggested, using model calculations, that anti-androgenic chemicals of unknown identities are widespread in U.K. effluents and receiving waters and that, in addition to the steroidal estrogens, these constituents of WWTP effluents are likely to play a major role in causing endocrine disruption in wild fish. In addition to sex hormone receptors, other receptor-mediated pathways in wild fish may also be affected. Recent studies show that synthetic GCs, which are used in large amounts as anti-inflammatory drugs by humans, are present in various wastewaters and receiving river waters [79,189]. These compounds can bind to the fish GR, and therefore their combined occurrence and total concentration of GCs in the environment may produce adverse effects to fish and other aquatic organisms [190].

So far, there are no field surveys in which a specific intersex condition in wild fish has been causally linked with exposure to a specific compound [58]. Instead, it is becoming evident that the feminization of wild fish has a multicausal etiology involving contributions from at least both steroidal estrogens and xenoestrogens and from other (as yet unknown) contaminants with anti-androgenic properties [186]. Also some causes of intersex remain unknown. For example, frequent and unexplained gonadal abnormalities including intersex features have recently been observed in whitefish (Coregonus lavaretus) in a relatively pristine lake in Switzerland with no evidence revealed for the presence of exogenic EDCs [191].

An increasing number of studies have observed estrogenic effects in marine fish, including large pelagic predators, evident in male fish in areas away from point sources [192–202]. The causes of these phenomena are not yet fully understood [18,42], but bioaccumulation of unknown substances, possibly through feeding, is a possible explanation for estrogenic exposure for at least some of these species.

A most important but difficult question to answer is the ecological relevance of the observed field observations (reviewed by [9,18,110]). To bridge results from laboratory exposures and field studies, a large-scale study in the field was recently conducted in which a lake was treated with environmentally realistic concentrations of EE2 over a period of three years and the fathead minnow population in the lake was monitored [203]. The results demonstrated that chronic exposure of fathead minnows to 5–6 ng/L of EE2 led to feminization of males through the production of VTG mRNA and proteins, impacts on gonadal development (intersex in males and altered oogenesis in females), and a fathead minnow population collapse in the lake. Incidentally, the few fathead minnows surviving several years after termination of exposure to EE2 showed recovery and again created a viable population. The Canadian study was a breakthrough because of it unique scale and effort and indisputably demonstrated that long-term exposure to relatively low environmental levels of estrogens has a dramatic impact on the population level. Several indirect effects by exogenous estrogens (such as EE2) were observed, such as the decline in numbers of trout feeding on fathead minnows and overall effects on fish population status could be derived. How do the results of this study translate to real field conditions in river systems downstream of sewage outflows? The Canadian study is probably a worst case with isolated fish populations dosed with EE2 concentration maintained artificially at a constant level. The exposure concentrations of 5–6 ng/L EE2 used in the Canadian experiment are in a similar range as those observed in several European and U.S. waters. For example, U.K. freshwater systems have been shown to reach 3.4 ng/L [204]. In effluents from Dutch WWTPs, concentrations were 2.6 ng/L measured, and in Dutch surface waters, maximum measured concentrations were 0.4 ng/L, so a factor of 10 lower. This concentration is near the detection limit but is still harmful to fish based on laboratory exposure. The results of this unique whole lake experiment demonstrate that environmental estrogen exposure in wild fish does have adverse effects on the health and viability of wild fish populations.

Early life stages of fish are particularly sensitive to EDCs due to critical endocrine-dependent developmental processes, such as sexual differentiation [175,205]. However, field studies of EDCs early life stages of fish are very limited. Various salmonid species in the North American Great Lakes and the Baltic Sea have been affected by early mortality syndrome (EMS), which affects the survival of early life stages of these species and results in adverse population level impacts [7,206]. Although associations of EMS with certain PCDDs, PCDFs, and PCBs were established, and possible mechanic linkages have been proposed, such as AhR agonism, altered thyroid hormone and retinol levels, and vitamin B1 deficiency, at present there seems to be no firm evidence to support that EDCs or other chemicals are involved in the etiology of this syndrome (reviewed by [1]). A basic problem is to fully understand the environmental factors influencing the occurrence of this disease.

Epizootics of thyroid hyperplasia and hypertrophy (affecting 100 percent of the population) have been reported in various species of salmon of the Great Lakes (see reviews of [1,58]). There is, however, no firm evidence linking thyroid hyperplasia observed in Great Lakes salmon with any specific chemical contamination [1]. Thyroid-disrupting effects in fish include a significant reduction in retinoid levels in both liver and plasma in flounder in polluted mesocosms, providing a clear indication that retinoid levels are affected by long-term exposure to contaminants [207]. Thyroid abnormalities in wild fish were also reported in mummichogs (Fundulus heteroclitus) from a polluted site and connected to a range of contaminants (especially mercury and petroleum hydrocarbons) in the United States [208] and altered thyroid hormone status in teleosts inhabiting San Francisco Bay associated with contaminants such as PCBs and to a lesser degree chlorinated pesticides, including DDT and chlordanes [209]. Together, these observations suggest that in fish, as in most other vertebrates, thyroid function appears to be sensitive to contaminant exposure, but efforts to investigate the occurrence of thyroid hormone–related effects in wild fish are limited.

An intriguing case is that of the declining eel (Anguilla anguilla) populations in Europe. Eels are unusual in that their fat content is an order of magnitude higher than that of other fish. Their levels of lipophilic contaminants generally reflect the elevated fat content, being five- to tenfold higher than in other fish and invertebrates, depending on the contaminant and the species. Several studies suggest that the decline in eel populations may be at least in part (in addition to overharvesting, habitat loss/degradation, oceanographic conditions, and parasites) due to the exposure to chemical compounds, including dioxin-like compounds [210–212]. A negative correlation exists between embryo survival time and TCDD toxic equivalent (TEQ) levels in the gonads implying TEQ-induced teratogenic effects. The elevated lipid content is important as an energy reserve and is regulated through steroidal–endocrine systems, although these fat reserves appear to be declining overall [213]. Also in this case, the role of EDCs is not well understood. Van Ginneken and coworkers [212] also indicated that transoceanic spawning migration is altered by PCBs.

2.2.5 Invertebrates

Perhaps the best example of population level effects by specific ED contaminants is masculinization (imposex) in female marine gastropods by TBT, a biocide formerly used in anti-fouling paints. A large number of studies have shown that the presence of very low concentrations of TBT (ng/L) will induce imposex or intersex in a range of gastropod species (e.g., dog whelks and netted dog whelks: for a review, see [214–216]). These contaminant effects have in the past caused (local) populations to decline, but nowadays these effects have been alleviated as a consequence of policy measures, such as progressive banning of the substance in anti-fouling formulations [217,218]. TBT-associated imposex is, however, still reported for gastropods in the Mediterranean and Black seas; in and near harbors, marinas, or coastal bays; as well as in waters off the coast of the United Kingdom [218].

TBT compounds are known to induce two different masculinization phenomena: imposex and intersex. Imposex is the occurrence of an entire or partial male organ in a female individual, while the primary female sexual organs are completely intact. The male organ is, as it were, superimposed. In the case of intersexuality in snails, however, some female organs have been transformed into a male organ. The final result of imposex and intersex development is the sterilization of females. Results of laboratory research confirm that this is a dose-dependent effect of TBT and levels as low as 1 ng/L TBT can cause female snails to grow male sex organs [214]. Although the possible mechanism of action of TBT was initially ascribed to inhibition of aromatase activity, altered metabolism of testosterone, and disruption of neuropeptide signaling [214,219], new studies indicate that TBT causes imposex in snails by acting as a retinoid X receptor (RXR) agonist [220–222].

It has been generally accepted that imposex is induced almost typically by TBT used in anti-fouling paints, which is based on the results of laboratory experiments using the dogwhelk (Nucella lapillus) and other species [223]. This is, however, not entirely the case. Laboratory studies have revealed that triphenyltin (TPT) can also promote the occurrence of imposex in some species [224]. Imposex has been found in Japanese waters, where the intensity of shipping traffic is low but where TPT concentrations are, nevertheless, high due to the presence of agriculture [225]. The discovery of imposex in specimens that lived before the TBT era demonstrates that chemical and/or nonchemical factors (e.g., changes in environmental conditions, infestation by parasites) other than TBT or TPT could alter the endocrine control of imposex development [226]. In addition, synergistic effects of organotin mixtures (TBT-Cl and TPT-Cl) have been reported [227].

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