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Alligators and Endocrine Disrupting Contaminants: A Current Perspective 1 [1]

['Guillette', 'Louis J.', 'Department Of Zoology', 'University Of Florida', 'Gainesville', 'Florida Usa', 'Crain', 'D. Andrew', 'Department Of Biology', 'Maryville College']

Date: 2000-06-01

Abstract

Many xenobiotic compounds introduced into the environment by human activity have been shown to adversely affect wildlife. Reproductive disorders in wildlife include altered fertility, reduced viability of offspring, impaired hormone secretion or activity and modified reproductive anatomy. It has been hypothesized that many of these alterations in reproductive function are due to the endocrine disruptive effects of various environmental contaminants. The endocrine system exhibits an organizational effect on the developing embryo. Thus, a disruption of the normal hormonal signals can permanently modify the organization and future function of the reproductive system. We have examined the reproductive and developmental endocrinology of several populations of American alligator (Alligator mississippiensis) living in contaminated and reference lakes and used this species as a sentinel species in field studies. We have observed that neonatal and juvenile alligators living in pesticide-contaminated lakes have altered plasma hormone concentrations, reproductive tract anatomy and hepatic functioning. Experimental studies exposing developing embryos to various persistent and nonpersistent pesticides, have produced alterations in gonadal steroidogenesis, secondary sex characteristics and gonadal anatomy. These experimental studies have begun to provide the causal relationships between embryonic pesticide exposure and reproductive abnormalities that have been lacking in pure field studies of wild populations. An understanding of the developmental consequences of endocrine disruption in wildlife can lead to new indicators of exposure and a better understanding of the most sensitive life stages and the consequences of exposure during these periods.

INTRODUCTION

Endocrine-altering actions of various chemical contaminants have been a major focus of recent research and policy discussions (Kavlock et al., 1996; Ankley et al., 1998). Chemical contaminants of man-made origin have been observed to mimic hormones, act as hormone antagonists or alter the synthesis and/or degradation of hormones (Table 1: for review, see Crain and Guillette, 1997). The most frequently reported endocrine action of environmental contaminants is estrogen mimicry (see McLachlan and Arnold, 1996). For example, the ability of a specific contaminant to induce cellular proliferation of estrogen-sensitive cells (e.g., uterine or breast cells) or its ability to bind with an estrogen receptor has been used as a definition of estrogenicity (Soto et al., 1995). Likewise, other contaminants have exhibited anti-androgenic actions in the developing male reproductive system (Gray et al., 1996). A number of synthetic chemicals and naturally occurring plant compounds have been defined as endocrine disrupting contaminants (EDCs) (see Guillette et al., 1996a; Gray et al., 1996; Soto et al., 1995; McLachlan and Arnold, 1996).

Endocrine alterations are inherently complex, as they are not limited to a particular organ (Fig. 1A) or molecular mechanism (Fig. 1B). For instance, apparent estrogenicity of a compound could be caused by (1) an increase in gonadal estrogen production, (2) a decrease in gonadal androgen production (thus, increasing the estrogen/androgen ratio), (3) an increase in the production of gonadotropin from the anterior pituitary or gonadotropin releasing hormone from the hypothalamus, (4) a decrease in hepatic enzymatic degradation/conjugation of estrogen, (5) an increase in the concentration of serum sex hormone binding proteins, limiting free hormone in the serum, (6) a decrease in cytosolic binding proteins (CBPs) that potentially limit free estrogen in the cell, or (7) agonistic binding of the compound to an estrogen receptor. Such complexity has limited the development of effective methods to screen for endocrine system disrupting compounds.

Much of the current literature on endocrine disrupting contaminants (EDCs) has focused concerns on possible detrimental alterations due to embryonic exposure (see Knobil et al., 1999; Guillette and Crain, 2000). In part, this may be due to the apparent extreme sensitivity of the developing embryo to chemical signals (Bern, 1992). Radical modification of embryonic structure and function, and thus adult form and function, can be induced by epigenetic factors. At this time, it is still unclear what the long-term effects of these embryonic modifications are on the health and reproductive potential of adults. Consequently, the population-level influences of such embryonic modifications are difficult to predict and are generally unknown.

In the following sections, we will briefly review data from our laboratories that are indicative of endocrine disruption in reptiles, specifically in the American alligator. For more complete reviews of endocrine disruption in wildlife and humans, see the rapidly growing literature in this area (see Colborn et al., 1993; Crain and Guillette, 1997; Gray et al., 1996; Guillette et al., 1996a; Toppari et al., 1996; McLachlan and Arnold, 1996; Nimrod and Benson, 1996; Sumpter and Jobling, 1995; Rolland et al., 1997; Knobil et al., 1999; Guillette and Crain, 2000).

ALLIGATORS AND CONTAMINANTS

Field Observations

Studies from our laboratory have shown that alligators living in a central Florida (USA) lake, Lake Apopka, exhibit a number of alterations of the reproductive and endocrine systems. Many of these modifications appear to be developmental defects which are detectable at hatching and which persist throughout juvenile life stages (see discussion below). We do not know if they continue into adulthood. Lake Apopka is a large, hypereutrophic and heavily polluted lake northwest of Orlando, FL (Fig. 2). In 1980, it was the site of a spill of the pesticide dicofol and has received extensive agricultural pesticide and nutrient runoff during the last 40 years. In the five years following the pesticide spill, juvenile recruitment plummeted on Lake Apopka due to decreased clutch viability and increased juvenile mortality (Woodward et al., 1993). The juvenile population remained depressed until the early 1990s (see Woodward et al., 1993; Rice et al., 1996). Recruitment increased the juvenile population in the 1990s, although pre-1980 population levels have not been observed (Woodward et al., 1993; Rice et al., 1996). The rise in juvenile recruitment reflects the rise in clutch viability. Clutch viability, the number of eggs that hatch versus the number laid, remained near or below 20% from 1983–1991 (Woodward et al., 1993), whereas other lakes in Florida such as lakes Okeechobee, Griffin and Jessup averaged 53.8%, 43.1%, and 43.3%, respectively (Masson, 1995). Recent studies indicate that egg viability on Lake Apopka has risen to 46–53% in 1994–1995 (Rice et al., 1996). In contrast, clutch viability rates for Lake Woodruff National Wildlife Refuge (NWR) averaged 87.5% between 1982–1990 (Woodward, 1996) and 76–79% for 1994–1995 (Rice et al., 1996). During the same time periods, clutch viability on Orange Lake, Alachua County, FL, was estimated at 87.5% (Woodward et al., 1992) and approximately 70% during recent years (Woodward et al., unpublished data). Thus, alligators from Lake Apopka, as well as many other freshwater lake systems in Florida, exhibit reduced clutch viability when compared to at least two other lakes—defined as reference sites—in the same broad watershed.

Although egg mortality declined and juvenile recruitment rose during the 1990s, a number of sublethal problems have been reported with the alligators living in Lake Apopka. Examinations of the reproductive and endocrine systems of hatchling and juvenile alligators from this lake, have demonstrated alterations in plasma estradiol-17β, testosterone, dihydrotestosterone and thyroxine concentrations (see Table 2) as well as morphological abnormalities of the testis and ovary (Guillette et al., 1994). The alterations in plasma hormone concentrations have been observed repeatedly over the last nine years (see Guillette et al., 1994; 1996b; 1997; 1999b; Pickford 1995; Pickford et al., 2000; Crain et al., 1998; Rooney, 1998).

Alterations in plasma testosterone concentrations throughout neonatal and juvenile life suggested that anatomical structures dependent on this hormone for growth and differentiation could also be altered. We reported previously that juvenile males from Lake Apopka had reduced phallus size coincident with lower plasma testosterone levels, when compared to a reference population living in Lake Woodruff NWR (Guillette et al., 1996b). Furthermore, there was a poor correlation between plasma testosterone concentrations and phallus size in juvenile male alligators from Lake Apopka, whereas a strong relationship was observed between these variables in males from Lake Woodruff (Guillette et al., 1996b, 1999b). We reevaluated the data collected on phallus size relative to body size in juvenile alligators from lakes Apopka (n = 165) and Woodruff (n = 219) between 1992 and 1999. These data indicate that the significant difference, we first observed with a smaller data set in the early 1990s, persists (see Fig. 3). We were not able to reevaluate the relationships between plasma testosterone and phallus size with this larger data set because these animals were collected during different summer months and plasma testosterone concentrations exhibit monthly variations whereas phallus size does not (Rooney, 1998).

It could be hypothesized that the altered plasma testosterone concentrations observed in the males from Lake Apopka were the result of stress and the release of adrenal steroids. Previous studies have shown in reptiles, specifically male alligators, that stress can depress plasma concentrations of testosterone (Lance and Elsey, 1986) or other sex steroids (for reviews, see Guillette et al., 1995b; Wingfield et al., 1998). Given this hypothesis, we examined the acute stress response of alligators living in Lake Apopka versus those living in a reference lake, Lake Woodruff NWR. In these studies, stress involved capture of wild animals and immediate confinement in a cloth bag for two hours after an initial blood sample was drawn. Juvenile alligators from lakes Apopka and Woodruff had similar plasma corticosterone concentrations within a minute of capture and after a two-hour acute (capture and confinement) stress test (Guillette et al., 1997). Males and females from both lakes displayed a dramatic 40-fold rise in plasma corticosterone concentrations during the two hours of capture and confinement. Further examination of the stress response in these two populations has confirmed that initial plasma concentrations of corticosterone are similar although the pattern of response observed varied between the two populations over a 20 hr period following capture (Rooney, 1998). However, it is unlikely that the altered plasma testosterone concentrations observed in males from Lake Apopka are due primarily to acute stress.

The relationship between body size and plasma androgen concentrations could also be affected by body condition. Animals that show different growth patterns or lower nutritional states could have reduced gonadal steroidogenesis. Initial studies indicate that there is no difference in the relationship between body mass and body size between males or females from lakes Apopka or Woodruff (Fig. 4). Current studies are examining the relationship between juvenile growth-patterns, age and body size.

It is also important to note that reports of altered endocrine parameters in juvenile alligators are available for other Florida wetlands, which are not associated with significant pesticide spills or point source contamination, such as Lake Okeechobee (Crain et al., 1998) and Lake Griffin, FL, USA (Guillette et al., 1999b). Abnormalities in juvenile alligators obtained from these lakes include: (a) reduced phallus size in males, (b) reduced plasma androgens in males, and (c) abnormal plasma thyroxine concentrations in juveniles of both sexes. Lake Griffin is part of the St. Johns River drainage system as are lakes Apopka and Woodruff, whereas Lake Okeechobee represents a distinct system that includes the Everglades. Lake Okeechobee is one of the largest freshwater lakes in the USA, outside of the Laurentian Great Lakes (see Fig. 2).

What induces the endocrine alterations described above? We have hypothesized that embryonic exposure to contaminants capable of acting as endocrine disruptors could induce organizational changes in the developing organism (Guillette et al., 1995a). In order for this hypothesis to have a factual basis, embryonic, neonatal and juvenile alligators must be exposed to types and concentrations of contaminants that are biologically relevant.

A number of contaminants identified in alligator eggs (Heinz et al., 1991; Giroux, 1998) and serum (Guillette et al., 1999a) exhibit an affinity for estrogen (ER) and/or progesterone (PR) receptors obtained from the alligator oviduct (Vonier et al., 1996; Guillette et al., unpublished data). These data indicate that many of the contaminants found in the embryonic or juvenile environment have the potential to be endocrine disrupting contaminants (EDCs). Further, eggs and juveniles from Lake Apopka have higher concentrations of a number of these EDCs when compared to similar samples obtained from alligators living in Lake Woodruff NWR (Fig. 5). Importantly, these chemicals, when combined, exhibit additivity or greater than additivity in ER competitive bindings assays (Vonier et al., 1996). Affinity for a receptor does not guarantee that a contaminant has a steroid mimicking effect, as it could equally act as a hormone antagonist (see (Gray et al., 1996; Kelce et al., 1995). Experimental testing of these compounds in vivo for endocrine disrupting ability is required.

Bioavailability of these compounds is also an important consideration. Although various contaminants exhibit an affinity for receptors, do they reach concentrations in the cell or nucleus that would represent biologically significant concentrations? Several studies suggest that serum and cytoplasmic binding proteins, that normally help regulate hormonal concentrations in the plasma or cytoplasm, show little affinity for EDCs (vom Saal et al., 1995; Arnold et al., 1996; Crain et al., 1998). These data suggest that unlike steroid hormones, whose free concentration in the plasma and cytoplasm is regulated by binding proteins, many contaminants are not regulated. Thus, if the contaminant can cross the cell membrane, all of the chemical in the blood is available to the cell (see Fig. 1B). We have recently demonstrated that not all EDCs interact in the same manner with serum and cytoplasmic binding proteins from the alligator, suggesting that a general conclusion that all EDCs are bioavailable to the same degree can not be made (Crain et al., 1998). This is not a new concept and has previously been reported for pharmaceutical agents such as the synthetic estrogen, diethylstilbestrol (Newbold, and McLachlan, 1985).

EXPERIMENTAL STUDIES

Crocodilians can provide an important model for testing embryonic estrogenic action, as they exhibit environmental sex determination (ESD) where estrogens play a fundamental role in gonadal differentiation. Unlike mammalian and avian species with a dominant genetic basis for sex determination, some reptiles, amphibians and fish respond to environmental cues such as temperature. In crocodilians, including alligators, incubation temperature strongly influences sex determination (Lang and Andrews, 1994; Lance, 1997). However, the effects of temperature on sex determination can be overridden by exposure of the developing reptilian embryo to extremely low concentrations (40 ng/kg) of the natural estrogen, estradiol-17β (Sheehan et al., 1999). Such exposure will induce sex reversal in a male to female direction. Additionally, ovarian differentiation can be altered by embryonic exposure to a inhibitor of aromatase, the enzyme essential for estrogen synthesis (Lance and Bogart, 1992; Wibbels and Crews, 1994). Given the estrogen-sensitive nature of sex determination in alligators, one can examine the estrogenicity of various environmental contaminants by testing for sex reversal.

Recent studies have shown that various pesticides or pesticide metabolites can override the temperature-sensitive sex determination mechanisms in alligator or turtle embryos, demonstrating that these compounds could act in a manner similar to natural estrogens (Bergeron et al., 1994; Matter et al., 1998; Willingham and Crews, 1999). Contaminants capable of altering sex (male to female) in alligator embryos include o,p′-DDE, p,p′-DDE, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), indole-3-carbinol, trans-nonachlor, and p,p′-DDD (see Table 3). As noted on Table 1, p,p′-DDE, a commonly bioaccumulated metabolite of the pesticide DDT, exhibits a variety of actions depending on the species and endpoint examined. Kelce et al. (1995) demonstrated that p,p′-DDE was a potent anti-androgen in vitro and in vivo in mammalian based systems. Similar molecular or organismic studies have not been performed for other vertebrate species. Both isoforms of DDD, also metabolites of DDT, are biologically active, having antagonistic activity on adrenal steroidogenesis in some species (Brown et al., 1973). The compound trans-nonachlor is a component of technical grade chlordane, in the past a pesticide used extensively in the USA for the treatment of termites but currently not in use. Each of these compounds are important contaminants in biological systems as they readily bioaccumulate and biomagnify in the food chain.

The concentrations reported to cause sex reversal are well within the range of concentrations measured in alligator eggs from Lake Apopka. That is, trans-nonachlor and p,p′-DDD cause sex reversal at a dose as low as 100 ppb (Crain, 1997; Rooney, 1998). In a previous study, Heinz and coworkers (Heinz et al., 1991) observed that alligator eggs collected from Lake Apopka in 1984 and 1985 had greatly elevated concentrations of p,p′-DDE: 5.8 ppm wet weight (1984: n = 3 eggs; range = 3.4–7.6 ppm) and 3.5 ppm wet weight (1985: n = 23 eggs; range = 0.89–29 ppm). These concentrations are similar to those observed by Giroux (1998) ten years later (n = 29 eggs; mean ± SE = 4.10 ppm ± 1.27). In addition to p,p′-DDE, alligator eggs (n = 23) collected in 1985 from Lake Apopka had detectable levels of p,p′-DDD (ND–1.8 ppm), dieldrin (0.02–1.0 ppm), and cis-chlordane (ND − 0.25 ppm) (Heinz et al., 1991). These concentrations are elevated compared to eggs collected on several other lakes. Not surprisingly, we have observed that these chemicals, as well as trans-nonachlor, mirex and endrin are present at ppb (µg/kg) concentration in the serum of juvenile alligators from Lake Apopka. If the eggs from Lake Apopka have elevated levels of compounds that can act estrogenically, one could hypothesize that the population would exhibit a skewed sex ratio, with a preference for females. We have not observed such a phenomenon. In fact, examining all of the juvenile alligators captured during our field work over the period 1992–1999 (n = 791), we have observed a male-biased sex ratio on both Lake Apopka (179M:111F) and Lake Woodruff NWR (309M:192F). Can this disparity in observations be explained?

In contrast to the DDT metabolites and other compounds discussed above, the herbicides atrazine and 2,4-D or phytoestrogen coumestrol are not estrogenic—do not induce male-to-female sex reversal—in experimental egg dosing studies (Crain et al., 1997; Matter et al., 1998; Guillette et al., unpublished data). Atrazine, however, at ppm doses can induce elevated testicular expression of the steroidogenic enzyme aromatase in alligator males if exposure occurs in ovo. Atrazine, like p,p′-DDE, exhibits a low affinity for the alligator estrogen receptor (Vonier et al., 1997). 2,4-D does not influence aromatase activity in alligators at the treatment doses used (Crain et al., 1997).

Could mixtures of endocrine disrupting contaminants produce complex interactions at the molecular level? We have observed that mixtures of just two pesticides or pesticide metabolites induce different endpoints than the compounds alone. For example, mixtures of p,p′-DDE and trans-Nonachlor do not induce sex reversal (male-to-female) as occurs with trans-Nonachlor alone (Guillette et al., unpublished data). That is, p,p′-DDE does not cause sex reversal on its own, but can block trans-Nonachlor-induced sex reversal when eggs are incubated at 33°C and treated just prior to the period of sex determination. However, p,p′-DDE exhibits mixed results in alligators from other laboratories. It caused sex reversal at high doses (1–10 mg/kg) in one experiment, synergized with its isoform o,p′-DDE to produce 100% sex reversal in another experiment, and in a third experiment acted as a partial anti-estrogen in the alligator egg when combined with ethinylestradiol (Matter et al., 1998). These data indicate a possible anti-estrogenic or estrogenic role for p,p′-DDE in alligators, depending on treatment temperature and its interaction with other chemicals. As indicated, receptor binding studies revealed that p,p′-DDE does show an affinity for the alligator estrogen receptor (aER). Interestingly, the characterization of the alligator oviductal estrogen receptor showed that this receptor has an affinity for the natural androgen, dihydrotestosterone but not testosterone, unlike rodent and human ER (Vonier et al., 1997). p,p′-DDE shows affinity for the androgen receptor (AR) of rodents and humans but no affinity for ERα. We have no data for the alligator AR, but our experimental data support the hypothesis that p,p′-DDE is a mixed-function ligand, having hormonal and anti-hormonal action depending on the environment (i.e., temperature, hormonal milieu) in which it is found. In this case, the response is complicated by the fact that the contaminant could also interact with more than one receptor class (ERs and ARs). These observations support the concerns of many researchers that in vitro screening studies, using only a few receptor types derived from human or traditional laboratory animals, could be very limited or misleading in predicting effects in the diversity of species exposed in various ecosystems.

CONCLUSIONS

A variety of environmental contaminants have the potential to act as endocrine disruptors in wildlife. In the American alligator, several populations from Florida are suspected to be experiencing reproductive and developmental impairment due to the presence of EDCs in the environment. Endpoints examined to date include clutch viability, plasma hormone concentrations, gonadal steroidogenesis and aromatase activity, gonadal morphology and phallus morphology/morphometrics. We have presented field and laboratory-based experimental data suggesting that contaminants can alter the endocrine and reproductive systems of wildlife by various mechanisms, including mechanisms other than hormone mimicry. Specifically, EDCs can act as hormone antagonists, alter steroidogenesis, and alter hepatic degradation of hormones. Moreover, these data indicate, as do many other recent studies, that a focus on “estrogenic” chemicals is inappropriately restrictive, as endocrine disrupting contaminants appear to interact with a number of other hormonal signals (e.g., androgens, progestins, thyroid hormones) and endocrine altering mechanisms (e.g., up or down regulation of steroidogenic enzymes, steroid hormone metabolism) (Crain and Guillette, 1997; Gray et al., 1996). Further complicating the matter is the fact that environmental exposure usually consists of a combination of chemicals unique to that particular area or ecosystem. It is still unclear as to whether any given combination of suspected EDCs will act in a non additive, additive or synergistic manner. To date, few studies have examined transgenerational effects of EDCs. That is, what are the long term consequences of embryonic exposure to EDCs in terms of reproductive fitness across generations? It is also important to note that species variation does occur and direct linkages between wildlife and human abnormalities may or may not occur. Understanding the normal functioning of the endocrine system of a diversity of species is essential if we are to determine the impact of ecosystem contamination by endocrine disrupting contaminants. Studies of endocrine disruption in wildlife species, must insure that (1) appropriate endpoints are examined for each species, (2) effects of EDCs are monitored across generations, (3) endpoints representative of responses to ecologically-relevant mixtures are examined, and (4) the focus is on embryonic, fetal and neonatal life stages.

Open in new tabDownload slide Fig. 1. A. Organs that are targets of endocrine disrupting compounds include the (a) brain, through production of regulating hormones such as gonadotropins (Gn), (b) the liver, through altering hepatic degradation or conjugation of steroids and the production of sex hormone binding proteins (SBP), and (c) the gonad, through altering steroids production. B. The specific molecular mechanisms of disruption are many, but include altering concentrations of cytosolic binding protein (CBP), and competitively binding receptors

Open in new tabDownload slide Fig. 2. Map of Florida showing the study locations. Note that all of the lakes examined are within the greater St. Johns River drainage system except Lake Okeechobee

Open in new tabDownload slide Fig. 3. The relationship between snout-vent length and phallus tip length in juvenile male alligators from lakes Apopka and Woodruff, FL. The relationships, examined using correlation statistics, are significantly different between populations (F = 4.429; P < 0.0001). Apopka: Phallus tip = −0.76 + 0.157*SVL; R2 = 0.56. Woodruff: Phallus tip = −1.97 + 0.212*SVL; R2 = 0.74

Open in new tabDownload slide Fig. 4. The relationship between snout–vent length and body mass in male and female juvenile American alligators from lakes Apopka and Woodruff, FL. No differences were observed between sexes or lakes using correlation analysis

Open in new tabDownload slide Fig. 5. Mean serum contaminant concentrations in male and female juvenile alligators from Orange Lake, Lake Woodruff and Lake Apopka, Florida, USA. Data from Guillette et al. 1999a

The field work reported above is the product of a collaborative effort among a number of researchers from several agencies, including the Florida Fish and Wildlife Conservation Commission, National Biological Service of the USGS and the University of Florida. We thank all of our colleagues of the Florida Alligator Research Team for their many years of assistance. We also thank the many undergraduate students that have helped with fieldwork, animal care and production of the histological preparations required by this work. Research reported here was funded in part by grants from the St. Johns River Water Management District, U.S. EPA and NBS/USGS through cooperative agreement #11-16-0009-1544 RWO#137 with the University of Florida and grants from the EPA (#CR821437 & #R824760-01-0) and the NIEHS (#PR471470) to LJG.

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