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Effects of endocrine-disrupting contaminants on reproduction in the American alligator, Alligator mississippiensis

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Effects of endocrine-disrupting contaminants on reproduction in the American alligator, Alligator mississippiensis
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Crain, David Andrew, 1970-
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English
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vii, 153 leaves : ill. ; 29 cm.

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Alligators ( jstor )
Animals ( jstor )
Carrier proteins ( jstor )
Contaminants ( jstor )
Eggs ( jstor )
Enzymes ( jstor )
Female animals ( jstor )
Hormones ( jstor )
Plasmas ( jstor )
Steroids ( jstor )
American alligator -- Physiology ( lcsh )
American alligator -- Reproduction ( lcsh )
Dissertations, Academic -- Zoology -- UF ( lcsh )
Endocrine glands -- Effect of chemicals on ( lcsh )
Veterinary endocrinology -- Florida ( lcsh )
Zoology thesis, Ph. D ( lcsh )
Lake Apopka ( local )
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bibliography ( marcgt )
non-fiction ( marcgt )

Notes

Thesis:
Thesis (Ph. D.)--University of Florida, 1997.
Bibliography:
Includes bibliographical references (leaves 132-152).
General Note:
Typescript.
General Note:
Vita.
Statement of Responsibility:
by David Andrew Crain.

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EFFECTS OF ENDOCRINE-DISRUPTING CONTAMINANTS ON REPRODUCTION
IN THE AMERICAN ALLIGATOR, ALLIGATOR MISSISSIPPIENSIS












By

DAVID ANDREW CRAIN


A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY

UNIVERSITY OF FLORIDA


1997

























This is dedicated to my wife Holly,
Who gives me endless love, peace, and joy,
And to my son Jared,
Who gives me hope for the future.











ACKNOWLEDGMENTS


There are numerous people to whom I owe thanks for assisting me in my research. First and foremost, I want to thank my friend and mentor Dr. Louis Guillette. Without his constant encouragement and assistance, none of this work would have been possible. I also want to thank my other committee members for their input: Dr. Karen Bjorndal, Dr. David Evans, Dr. Larry McEdward, and Dr. Dan Sharp helped mold my ideas into sound studies. Cathy Cox, Ed Orlando, Dan Pickford, and Andy Rooney were fellow graduate students who both assisted with various laboratory and field projects and helped develop my research ideas. I was fortunate to be able to work with the following undergraduates who helped with my laboratory work: Bart Edmiston, Amy Pickle, Megan Pew, Michelle Scargle, Daniel Spiteri, Shadi Tolymat, and Christi Waldi. All of these undergraduates are of the highest caliber, and I am indebted to them for their hard work.

Collection of wild alligators and eggs was made possible through the collaborative assistance of Allan Woodward of the Florida Game and Freshwater Fish Commission and Franklin Percival of the Florida Cooperative Fish and Wildlife Research Unit. I am also indebted to the many others who assisted in the field studies.

Funding for my research was kindly provided through a graduate student fellowship from the Environmental Protection Agency, a research grant from the University of Florida Division of Sponsored Research, and Greg Masson of the U.S. Fish and Wildlife Service. Also, the University of Florida Department of Zoology provided support that made this work possible.


iii














TABLE OF CONTENTS





page


ACKN OW LED GM EN TS .......................................................................................... iii

A B S T R A C T ................................................................................................................... v i

CHAPTERS

I INTROD UCTION .............................................................................................. . 1

T h e P ro b le m .............................................................................................................. 1
T h e P u rp o se ...............................................................................................................2
The Rationale...................................................................................................... 3


2 ENDOCRINE-DISRUPTING CONTAMINANTS AND REPRODUCTION
IN VERTEBRATE W ILD LIFE ............................................................................ 5

Introduction........................................................................................................ 5
An Evolutionary Perspective ................................................................................. 5
Theory of D isruption: Organization vs. Activation .............................................. 10
Alteration of Reproductive Tissues ...................................................................... 14
C o n c lu sio n s ............................................................................................................. 3 3

3 SEX-STEROID AND THYROID HORMONE CONCENTRATIONS IN
JUVENILE ALLIGATORS (ALLIGATOR MISSISSIPPIENSIS) FROM
CONTAMINATED AND REFERENCE LAKES IN FLORIDA .........................36

In tro d u c tio n ............................................................................................................. 3 6
M aterials and M ethods......................................................................................... 39
R e su lts .....................................................................................................................4 6
D isc u ssio n ...............................................................................................................5 8


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4 TESTOSTERONE SYNTHESIS IN EMBRYONIC AND JUVENILE
ALLIGATORS EXPOSED TO ENDOCRINE-ALTERING
ENVIRONMENTAL CONTAMINANTS........................................................... 65

In tro d u c tio n .............................................................................................................6 5
M aterials and M ethods........................................................................................ . 67
R e su lts .....................................................................................................................7 2
D isc u ssio n ...............................................................................................................7 4


5 ALTERATIONS IN STEROIDOGENESIS IN ALLIGATORS (ALLIGATOR
MISSISSIPPIENSIS) EXPOSED NATURALLY AND EXPERIMENTALLY
TO ENVIRONMENTAL CONTAMINANTS ......................................................... 85

In tro d u ctio n ............................................................................................................. 8 5
M aterials and M ethods ........................................................................................ . 87
R e su lts ................................................................................................................. .... 9 3
D iscussion .......................................................................................... 99


6 CELLULAR BIOAVAILABILITY OF NATURAL HORMONES AND
ENVIRONMENTAL CONTAMINANTS AS A FUNCTION OF SERUM
AND CYTOSOLIC BINDING FACTORS................................. 105

Introduction..................................................................... 105
M aterials and M ethods................................................... 108
Results................................... ........ ......................... 113
Discussion ............................................................. 119


7 SUMMARY AND CONCLUSIONS ..................... .............. 126

The Rationale Revisited ......................................... 126
The Purpose Revisited .......................................... 127
The Problem Revisited ......................................... 128

LIST OF REFERENCES.......................................... 132

BIO G RAPH ICAL SKETCH .................................................................... 153


v













Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy

EFFECTS OF ENDOCRINE-DISRUPTING CONTAMINANTS ON REPRODUCTION IN THE AMERICAN ALLIGATOR, ALLIGATOR MISSISSIPPIENSIS By

David Andrew Crain

August, 1997


Chairman: Louis J. Guillette, Jr.
Major Department: Zoology

Wildlife and humans are exposed to numerous chemicals in the environment that can alter the function of the endocrine system. These endocrine-disrupting contaminants (EDCs) change the normal functioning of reproduction in many wildlife species. This dissertation contributes to the current knowledge of contaminant-induced endocrine disruption by reviewing the phenomenon in wildlife species, describing endocrine disruption in several wild populations of alligators, and examining the mechanisms of this endocrine disruption.

Plasma concentrations of estradiol- 170 (E2), testosterone (T), triiodothryonine

(T3), and thyroxine (T4) in juvenile alligators from two contaminated lakes and one reference lake in Florida are presented. Males from the two contaminated lakes (Lakes Apopka and Okeechobee) showed no correlation between T4 and total length, whereas males from the reference lake (Lake Woodruff) showed a strong correlation. Male


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alligators from the contaminated lakes had significantly lower T concentrations compared to males from the reference lake, and there was a poor relationship between T and total length in Apopka males (r2=0.007, p=0.75). These results are consistent with other studies indicating that alligators living in Lakes Apopka and Okeechobee experience endocrine disruption.

Mechanisms underlying the alterations in steroid hormones were studied in a series of experiments. First, gonadal T synthesis was examined in wild alligators from a contaminated lake (Lake Apopka) and a reference lake (Lake Woodruff). Whereas there was no difference in T production in animals of the two lakes, male alligators from Lake Woodruff that were exposed in ovo to p,p '-DDD had elevated T synthesis. Testes from these exposed animals appeared developmentally accelerated. Second, steroidogenesis was studied by examining aromatase activity in gonads from hatchlings that were exposed in ovo to two modern-use herbicides. Alligators that were exposed to atrazine had higher aromatase activity compared to controls. This study also showed that gonadal aromatase activity is higher in animals from Lake Woodruff than those from Lake Apopka. Third, evidence is presented that few EDCs interact with steroid binding proteins, thus increasing the cellular exposure to EDCs relative to endogenous hormones. The dissertation concludes by presenting a model summarizing the effects of EDCs on steroid hormone dynamics and emphasizing future research needs.


vii













CHAPTER 1
INTRODUCTION


The Problem


The study of adverse effects of xenobiotics-defined as toxicology-dates back to the earliest humans, who used animal venoms and plant extracts for hunting. Scholars of ancient Greece, such as Socrates (470-399 BC) and Theopharstus (370-286 BC), studied lethal dosages and began classifying poisons. But it was not until the Age of Enlightenment that toxicology was viewed as a science and a discipline. Philippus Aureolus Theophrastus Bombastus von Hohenheim-Paracelsus (1493-1541) is viewed as the father of modern toxicology due to his premise that (1) experimentation is essential to determine the response of a chemical, (2) a distinction should be made between the therapeutic and toxic properties of a chemical, and (3) a chemical's properties are sometimes, but not always, distinguishable only by dose. Paracelsus' wisdom is evident in his most famous quote:

All substances are poisons; there is none
which is not a poison. The right dose
differentiates a poison from a remedy.
Circa 1533

Modern toxicology primarily has focused on lethal endpoints as a result of animal and human exposure to large quantities of numerous toxic compounds. Two of the most common tests used to determine the toxicity of a chemical are the LD5o (lethal dose required to kill 50% of the experimental units) and the Ames test (used to determine the


I





2


dose that stimulates mutagenesis, and thus cancer). While knowledge pertaining to these tests is critical, other sublethal endpoints are necessary for evaluating the toxicity of a particular xenobiotic.

Whereas toxicologists view the individual as their experimental unit, ecologists and conservationists view a population of animals as the experimental unit. With this "population perspective," reproduction of the individual is as important a variable as the survivorship of that individual. In essence, survival is not the seminal variable dictating survival of a population; reproduction is that variable. Therefore, successful reproduction, and not death or carcinogenicity exclusively, should be used as an endpoint in the assessment of the toxicity of a xenobiotic.

Reproduction has long been a consideration in toxicity studies, but the last 5-10 years have seen an exponential increase in the number of studies examining the adverse effects of compounds on reproduction. During this time, it has been discovered that many xenobiotics adversely affect reproduction by altering the endocrine system (Colborn et al., 1993). This has led to a new subdiscipline that merges the fields of toxicology and endocrinology-the study of endocrine-disrupting contaminants (EDCs). As the endocrine system is involved in the regulation of virtually all biological phenomena, there is a high probability that EDCs also alter functions other than reproduction (such as growth and maintenance).

The Purpose


The purpose of this dissertation is to better understand the scope of endocrine

disruption in vertebrate wildlife and the ways that environmental contaminants can disrupt





3


the endocrine system. Specifically, the research described herein examines endocrine disruption in the American alligator (Alligator mississippiensis). Several recent studies have noted structural and functional abnormalities of the reproductive system in alligators exposed to EDCs (Guillette et al., 1994; Guillette et al., 1995b; Guillette et al., 1996b), but the mechanisms of endocrine disruption in these animals remain poorly understood. It is hoped that this dissertation will help elucidate the phenomena of endocrine disruption.

The Rationale


In order to ascertain the magnitude and scope of contaminant-induced endocrine disruption in vertebrates, a review will be provided in Chapter 2 that examines endocrine disruption in vertebrate wildlife. Also in this chapter, concepts that are necessary for understanding endocrine disruption will be presented. In Chapter 3, data will be presented from wild alligator populations, two of which are exposed to a variety of environmental contaminants. In Chapters 4, 5, and 6, the specific mechanisms through which environmental contaminants can cause endocrine disruption in the American alligator will be examined. The theoretical framework for testing these potential mechanisms is presented in Fig 1-1. A contaminant potentially can influence the endocrine system by causing an alteration at any point in the cycle of steroid hormone dynamics. The potential for contaminants to alter steroid production will be explored in Chapters 4 and 5, and contaminant bioavailability will be considered in Chapter 6. Finally in Chapter 7, conclusions and future perspectives will be given.







Steroid Production


Steroid


Biotransformation


Steroid
Availability


Hormone Excretion


Steroid Action


Figure I-I A simplistic model of steroid hormone dynamics. After a steroid hormone is produced, its bioavailability is
regulated by circulating intracellular binding proteins. Then the steroid elicits action by binding to a specific
receptor. The steroid is then either excreted in the urine after hepatic conjugation reactions or biotransformed into
another steroid, which will begin the cycle of bioavailability, action, and excretion/biotransformation.













CHAPTER 2
ENDOCRINE-DISRUPTING CONTAMINANTS AND REPRODUCTION IN VERTEBRATE WILDLIFE

Introduction


The fields of toxicology, endocrinology, and reproductive physiology recently have combined resources to study the effects of endocrine-disrupting contaminants (EDCs) in wildlife populations. EDCs include a wide variety of chemicals that are only related by the ability to disrupt normal function of an animal's endocrine system (McLachlan, 1993). Although studies documenting endocrine disruption by contaminants have been conducted for many years, only recently have studies systematically explored the effects and mechanisms of EDCs. This recent synthesis has led to the hypothesis that anthropogenic EDCs are associated with an increase in abnormalities of the reproductive system and a decrease in reproductive success in vertebrates (Colborn and Clement, 1992). This brief review considers the phenomenon of contaminant disruption of wildlife reproduction at several levels: evolutionary, tissue, and mechanistic. Only through such an integrative perspective can an accurate representation be achieved and solutions gained.

An Evolutionary Perspective


Animals are constantly exposed to many foreign chemicals in their diet, and these compounds can decrease the survival and reproductive capacity of individuals. Given that evolution favors maximal reproductive success, the evolution of mechanisms to eliminate


Note: This chapter is published in Reviews in Toxicology (Crain and Guillette, 1997a).
5





6


the deleterious effects of xenobiotics is expected. One such mechanism that has evolved in animals is the phase I and phase II biotransformation pathways (Brouwer, 1991). Phase I and II processes can render xenobiotics more lipophilic through the addition of hydroxyl moieties (phase I) or conjugation to polar endogenous molecules (phase II). Most phase I reactions are carried out by a particular class of enzymes, the cytochrome P450 enzymes of the liver. These enzymes increase the hydrophilicity of the compound by adding a hydroxyl moiety to it. The induction of hepatic P450 enzymes after contaminant exposure is a very sensitive and phylogenetically conserved mechanism, and as such, P450 induction is commonly used to monitor contaminant exposure in wildlife species (Rattner et al., 1989).

While many P450 enzymes are involved in decontamination, other P450 enzymes are involved in steroidogenesis-the production and conversion of steroid hormones. Thus, a family of enzymes that are induced by contaminant exposure to detoxify compounds could also cause an alteration in the hormonal environment of an animal. This leads to an evolutionary dilemma concerning the two fundamental constraints that dictate natural selection: the constraint of reproduction and the constraint of survival (Jacob, 1977). Survival requires that xenobiotics be detoxified and eliminated but, at the same time, reproduction can not be compromised. Therefore, to support survival and reproduction, P450 enzyme induction must be specific to detoxification and not alter steroidogenesis. The specificity of P450 induction is unknown, and this should be an area of future research. It is possible that many of the reproductive abnormalities seen in vertebrates during recent decades are due partially to the alteration of normal steroidogenesis as a result of increased exposure to a wider range of contaminants.





7


Many plants contain endocrine-disrupting compounds such as antithyroidal goitrogens (Yamada et al., 1974), phytoestrogens (Hughes, 1988), and androgen disrupters (Gray et al., 1996) that can alter reproduction after ingestion. The innovation of such endocrine-altering compounds by plants is predicted by evolutionary theory that suggests that plants will respond to predation by herbivores (Ehrlich and Raven, 1964; Jansen, 1980). Reproduction of wildlife species is adversely affected by plants containing endocrine-disrupting compounds (Leopold et al., 1976; Berger et al., 1977; Howell and Denton, 1989), but such plant-animal interactions have evolved through time to yield an evolutionary stable strategy for the wildlife and plants. For instance, ruffed grouse utilize aspen buds as a major food source, but the grouse have had to adapt to the endocrinedisrupting effects of a component of the aspen buds, coniferyl benzoate. When ingested, coniferyl benzoate is metabolized into ferulic acid which causes anti-reproductive effects through altering estrogen and prolactin function (Jakubas et al., 1993). The grouse avoid such endocrine-altering effects by both selectively feeding on buds having low concentrations of coniferyl benzoate (Jakubas et al., 1989) and utilizing aspen buds less frequently when coniferyl benzoate levels are high (Jakubas and Gullion, 1991). In addition to such behavioral alterations, adaptation to plant endocrine disrupters can involve chronological or physiological adjustments. Wildlife can adapt to the compound by altering the timing of reproduction to avoid exposure during critical reproductive stages. Selection would favor such chronological avoidance, since animals that avoid the phytotoxicants would have increased reproductive success. Wildlife can also physiologically adapt to plant-derived endocrine disrupters. For instance, after ingestion the phytoestrogen genistein induces many of the same effects as 17p-estradiol (Levy et al.,





8


1995), but genistein also stimulates production of sex-hormone binding globulin (SHBG) (Mousavi and Aldercreutz, 1993) and suppresses aromatase activity (Adlercreutz et al., 1993), both of which reduce the amount of bioavailable estrogen. The role of SI-BG in controlling the bioavailability of genistein itself is unknown, but SHBG could be protective if it binds genistein as it does 170-estradiol.

It is clear that coevolution has provided a means for animals to adapt to

phytochemical mimicry of reproductive hormones, but such adaptations have not evolved for animals exposed to anthropogenic EDCs. From a classical gradualist perspective, adaptation to an environmental condition (here, EDC exposure) requires a predictable exposure over a relatively long period of time-generations. Whereas wildlife have coevolved with plants for hundreds of millions of years, exposure of wildlife to large numbers of anthropogenic EDCs is limited to the past two centuries-since the onset of the industrial revolution. Indeed, the largest exposure in number of compounds and concentrations of these compounds has occurred in the last 50 years. Relatively few generations have been produced in this short amount of time, precluding extensive adaptation for many long-lived vertebrates. Given generations, constant or predictable exposure of a population to a contaminant provides an opportunity for adaptation. By chance, some individuals will be better able to survive and reproduce during the exposure. These animals produce progeny with similar characteristics, and eventually the population will consist of animals able to survive and successfully reproduce during the contaminant exposure. Although exposure to some EDCs could be predictable (e.g., constant exposure of a fish population to sewage effluent), most wildlife populations are exposed to EDCs in an unpredictable manner, making chronological and physiological adaptation





9


difficult. Moreover, the combination of chemicals that wildlife are exposed to can vary dramatically by location, season, and life stage.

It has been suggested that exposure to dietary phytoestrogens is far greater than exposure to industrial estrogenic compounds and, thus, that the contribution of industrial estrogenic compounds to reproductive dysfunction is nominal (Safe, 1995). Although endocrine disruption is likely to occur in embryos exposed to high concentrations of maternally ingested phytoestrogens, laboratory data suggest that the potency of ingested phytoestrogens is nominal when compared with other modes of exposure (such as direct in utero or in ovo exposure) (Cain et al., 1987; Lien and Cain, 1987). Guillette et al. (1 996a) recently argued that the in vivo estrogenicity of an "ecoestrogen" (an environmental contaminant that acts as an estrogen agonist) is determined by many factors that differ between natural phytoestrogens and synthetic estrogens, namely (a) the binding affinity of the compound to the estrogen receptor, (b) the accumulation of the compound in the environment and the body, (c) the degradation or metabolism of the compound in the environment and the body, and (d) the availability of the ecoestrogen to the target cell. When these variables are considered, the potential "potency" of many EDCs may be substantial. Additionally, each of these variables can be contaminant- and species-specific, and a significant degree of phylogenetic variation can be seen in the response to EDCs. For example, Kelce et al. (1995) found that p,p'-DDE acted as an androgen antagonist in rodents whereas Soto et al. (1995) found it to be estrogenic in human MCF-7 cells.





10


Theory of Disruption: Organization vs. Activation


The effects of EDCs on reproduction can be classified as either organizational or activational. A contaminant that permanently modifies the morphology or function of a tissue as the result of exposure during a particularly sensitive period of development is said to have an organizational effect, whereas if an EDC temporarily alters the function of a normally organized tissue, it has induced an activational effect (Guillette et al., 1995a). Figure 2-1 illustrates the typical periods when organizational and activational effects are induced in a vertebrate. In general, alterations occurring from gamete production through juvenile development are permanent and organizational in nature, whereas insults during mature stages are transitory and activational. Although the dichotomy between organizational and activational effects is imperfect (Arnold and Breedlove, 1985), this concept provides an appropriate framework for the discussion of contaminant-induced endocrine effects.

As an embryo develops, numerous inductive processes are initiated through

extremely complex networks and cascades (Jacobson, 1966). The evolutionary result is that the overall developmental process will resist significant modification (Raff and Kaufmnan, 1991). However, exposure of embryos to compounds that mimic or block critical developmental signals can readily alter normal developmental processes and, thus, the organization of structures. Embryos are particularly sensitive to organizational disruption by EDCs due to (a) high rates of cellular division, differentiation, and metabolism, (b) critical windows of sensitivity, (c) the relative amount of contaminant





I I


available to cells, and (d) mobilization of maternal bioaccumulated contaminants (Guillette et al., 1995a).

There is a great deal of variation in the organizational responses of various phyla to EDC exposure. For instance, consider the organizational effects of neonatal polychlorinated biphenyl (PCB) exposure in two animals that have different modes of sex determination. Sex in some species of turtles (and many other reptiles) is determined by the temperature of embryonic development (Pieau, 1996), but the development of the embryonic gonad can be altered by steroid exposure. Administration of exogenous estradiol can cause feminization of individuals incubated at male-producing temperatures (Crews et al., 1991) and, thus, estrogens and chemicals that act as estrogen agonists have the potential to alter sexual differentiation. Indeed, Bergeron et al. (1994) discovered that exposure of developing turtles (Trachemys scripta) to less than 9 ppm of some PCBs could override normal temperature-dependent male sex determination by producing turtles with ovaries and Mullerian ducts. Additionally, some PCB mixtures at less than 1 ppm synergize to produce ovarian development in T scripla embryos incubated at a maleproducing temperature (Crews et al., 1995a). Conversely, in a species with genetic sex determination, the rat, exposure of embryos to a mixture of PCBs (Aroclor 1242 and Aroclor 1254) actually increases adult testis weight and sperm production and does not cause sex reversal (Cooke et al., 1996). These masculinizing effects are thought to be mediated through the induction of hypothyroidism, which leads to increased Sertoli cell proliferation. Such masculinizing effects were not noted in T. scripta exposed to PCBs.


























Figure 2-1. Superimposed on the life cycle of a sexually reproducing, diploid organism are the periods when organizational and activational disruption predominately occur.










GAMETES


0/





ADULT




Activation


JUVENILE


Ell


Organizational
ZYGOTE


w 0
2N

(EMBRYO



6\e





14


In many instances, it is not clear whether endocrine alterations are the result of organizational or activational disruption. Consider the altered reproductive behaviors noted in herring gulls exposed to a mix of organochlorines (Fox et al., 1978; Gilman et al., 1978). The exposed gulls show a reduction in the defense of territories, a behavior mediated by the endocrine system. Wingfield (1987) noted that testosterone concentrations are related to the intensity of aggression in most birds. Thus, aggressive behavior could change if the regulatory effects of testosterone are altered. EDCs that act in an antiandrogenic manner could cause aggressive behavior to subside. DDE (1,1dichloro-2,2 bis(p-chlorophenyl)ethylene), a major breakdown product of DDT (1,1,1 trichloro-2,2 bis(p-chlorophenyl)ethane), is a widespread organochlorine contaminant known to bioaccumulate and biomagnify readily (Clement International Corporation, 1994). Since p,p'-DDE acts as an EDC via antiandrogenic mechanisms in rodents (Kelce et al., 1995), one could hypothesize that DDT exposure explains the alterations in herring gull behavior. It is not known if such behavioral aberrations are organizational or activational in origin.

Alteration of Reproductive Tissues


From the time of conception until death, hormones affect the morphology and physiology of an individual. Likewise, any contaminant that alters the dynamics of hormones can cause morphological and physiological alterations. This section examines the structural and functional alterations induced by EDCs in the reproductive system, the liver, and the thyroid, as they relate to the development and function of vertebrate





15


reproduction. Table 2-1 presents a representative list of reproductive abnormalities caused by exposure of wildlife to EDCs.

Reproductive System-Gonads

The gonads serve both to provide gametes for reproduction and to produce steroid hormones that mediate reproductive physiology and behavior. Either of these functions can be altered activationally or organizationally by exposure to EDCs.

Within a female's ovaries, several organizational, morphological changes are indicative of exposure to EDCs. In a normal female, the ovarian follicles contain one oocyte (the future ovum or egg) having a single nucleus. Mice experimentally treated preor postnatally with diethylstilbestrol (DES), a potent synthetic estrogen, exhibit polyovular follicles (Iguchi, 1992). The existence of such polyovular follicles has been used as a marker of endocrine disruption in American alligators (Alligator mississippiensis) (Guillette et al., 1994). Alligators from Lake Apopka, a contaminated lake in central Florida, exhibit polyovular follicles and polynuclear oocytes, suggesting organizational disruption as a result of exposure to estrogenic contaminants. Laboratory data support the hypothesis that such abnormalities are organizational in origin. Preliminary studies suggest that experimental exposure of neonatal alligators to p,p'-DDE can induce the development of polyovular follicles (Pickford, 1995).

In many vertebrates, the developing gonad is composed of a cortical layer

surrounding an inner medullary region. In males, the cortex degenerates to a single cell layer and the medullary region expands, whereas in females, the medullary region degenerates and the cells in the cortex proliferate. The retention of a gonadal cortex in males and the existence of primordial germ cells in this cortex are indicative of endocrine





16


disruption (Fry and Toone, 1981; Guillette et al., 1994). Fry and Toone (Fry and Toone, 1981) found that embryonic exposure of gulls (Larus calfornicus) to low concentrations (from 2 ppm) of o,p'-DDT and methoxychlor caused males to develop clusters of primordial germ cells in the cortex of the testes. Although it is not known if the altered gull testes are capable of producing viable sperm, many EDCs have the potential to organizationally alter the function of the testes. Lye et al. (1997) found that 53% of male flounder (Platichthysflesus) surveyed in a population exposed to sewage treatment effluent had testicular abnormalities including thick interstitial tissue, testicular cysts, and aberrations from the normal elongate structure. Alteration in sperm production can also be a sign of disruption, as decreased spermatogenesis is one of the most sensitive signs of reproductive toxicity in male mammals (Peterson et al., 1993). When mother rats are treated with 4-octylphenol and butyl benzyl phthalate (both common EDCs), male offspring have significantly reduced testicular size and sperm production (Sharpe et al., 1995).

Recently, the reproductive impairment of male Florida panthers (Felis concolor coryi) has been attributed, in part, to exposure to EDCs (Facemire et al., 1995). Male panthers exhibit sterility, cryptorchidism (undescended testes), low sperm concentration, poor sperm motility, and a high proportion of abnormal sperm (Roelke, 1990). Studies examining semen quality determined that the abnormalities were not due to differences in steroid hormones but more likely were mediated at the testicular level (Barone et al., 1994). Although it has been speculated that these reproductive abnormalities are due to inbreeding (Roelke et al., 1993), Facemire et al. (1995) note high concentrations of DDE





17


and PCBs in panther fat and hypothesized that the alterations in the male reproductive system could be due, at least in part, to EDC-induced organizational disruption.

The hypothesis that p,p '-DDE can induce abnormal development of the male reproductive tract is supported by laboratory evidence. Experimental exposure of fetal male rats to the antiandrogenic metabolites of the fungicide vinclozolin produces characteristics similar to those seen in the panthers, including cryptorchidism and atrophic seminal vesicles (Gray et al., 1993b; Kelce et al., 1994). Becausep,p'-DDE appears to act through mechanisms similar to vinclozolin (Kelce et al., 1995), the hypothesis forwarded by Facemire et al. (1995) is plausible and should be examined further. The antiandrogenic effects ofp,p '-DDE have also been implicated in altering the sexual characteristics of male American alligators. Androgens are responsible for the formation of sexual genitalia of male reptiles (Raynaud and Pieau, 1985), and wild alligators exposed to p,p'-DDE and other EDCs have significantly smaller penis size in relation to body size when compared to control animals (Guillette et al., 1996b). Reproductive System-Ducts

EDCs can also cause abnormal formation and function of reproductive ducts. In birds, only the left oviduct and ovary are functional in females of most species, a trait uncommon among vertebrates. The right ovary is rudimentary, and can be induced to develop into a mature (although non-functional) ovary upon hormonal manipulation (van Teinhoven, 1983). This observation leads to the prediction that birds exposed to estrogenic contaminants during gonadal organization could have both right and left ovaries and/or right and left oviducts. Indeed, Fry and Toone (1981) found that exposure of female embryonic gulls (Larus californicus) to op '-DDT (in concentrations





18


comparable to those found in the wild - 2 to 100 ppm) induced development of the right oviduct, which is usually rudimentary. Additionally, male breeding birds were extremely rare in the colony with heavily contaminated individuals, presumably due to the feminization or "neutering" effects of DDT exposure (Fry, 1995).

Perhaps the most well known example of reproductive "endocrine disruption" is the effects of organochlorines (namely DDT) on eggshell thickness in birds (Elliot et al., 1988; Burger et al., 1995). The mechanism of this eggshell thinning is still debated but appears to be associated with contaminant-induced alterations of the enzymatic functioning of the oviduct. Ducks exposed to p,p'-DDE have reduced prostaglandin synthesis in the eggshell gland mucosa, increased calcium content in eggshell gland mucosa, and decreased calcium in the shell gland lumen (Lundholm, 1994). Thus, the eggshell-thinning effects of DDT are consistent with the inhibition of prostaglandin synthesis in eggshell gland mucosa by p,p'-DDE (Lundholm, 1994).

Mammalian wildlife also exhibit reproductive tract alterations as a result of exposure to EDCs. Wild mink populations have declined in the Southeast, and these declines have been attributed to PCB-induced reproductive dysfunction (Osowski et al., 1995). Indeed, the reproductive tract of female mink is highly sensitive to organochlorineinduced impairment, as Patnode and Curtis (1994) found that 3,3',4,4',5,6' hexachlorobiphenyl (a coplaner PCB) caused significant upregulation of uterine estrogen receptors and, most notably, significantly increased the progesterone receptor dissociation constant.

Reproductive impairment in seals has also been attributed to PCB exposure.

Female Baltic seals have exhibited low reproductive output in recent years, and this has





19


been attributed largely to the occurrence of uterine occlusions (Helle, 1989). Such occlusions are thought to be caused by PCB exposure, as peak PCB levels (up to 100 ppm in blubber) in the Baltic seals preceded a rapid increase in the frequency of uterine occlusions (Helle, 1989). The detrimental impact of PCBs on seal reproduction has been documented experimentally in harbor seals (Reijnders, 1986). Reijnders (1986) found that a diet of PCB-contaminated food caused female harbor seals to resorb their embryos. This has led researchers to hypothesize that PCB-induced uterine occlusions result after an initial interruption of pregnancy around the time of implantation (Reijnders, 1984; Helle, 1989), a phenomenon dependent on the correct hormonal environment. Liver

The liver is an important organ associated with reproduction in that it synthesizes proteins, such as vitellogenin and sex-hormone binding globulin, necessary for normal hormonal homeostasis in the plasma. This organ also converts steroids to more hydrophilic excretory products. A recently observed, much publicized, hepatic alteration following exposure to EDCs is the abnormal synthesis of vitellogenin (Vg) in nonmammalian male vertebrates. Vitellogenin is an estrogen-dependent protein that is the major component of yolk in eggs of oviparous vertebrates (van Teinhoven, 1983). Females normally produce Vg during the reproductive stages of egg assimilation, but Vg can also be induced by exogenous estrogens such as estradiol- 170 or estrogenic contaminants in males and non-reproductive females. For instance, adult male frogs (Xenopus laevis) and turtles (Trachemys scripta) produce significant levels of Vg when treated with the environmental estrogen o,p'-DDT (Palmer and Palmer, 1995). The consequences of male Vg secretion are unknown, but the presence of Vg in males is





20


correlated with a reduction in testis size in male trout (Oncorhynchus mykiss) exposed to estrogenic contaminants (Jobling et al., 1996). Vitellogenin has also been detected in juvenile male and female guppies (Poecilia reticulata) exposed to 3hexachlorocyclohexane, a persistent environmental contaminant (Wester et al., 1985). Here, the production of Vg was associated with a dose-dependent increase in hepatocellular basophilia, indicating that 0-hexachlorocyclohexane induced both structural and functional damage. Similar hepatic structural damage is noted in winter flounder (Pleuronectes americanus) exposed to pulp and paper mill effluent (Khan et al., 1994).

Vitellogenin production by cultured hepatocytes has been used as an indicator of the estrogenic potency of chemicals. Six different phytoestrogens (biochanin A, coumestrol, daidzein, equol, formononetin, or genistein) stimulate Vg production in rainbow trout hepatocytes, the phytoestrogens having approximately one-thousandth the potency of 17p-estradiol (Pelissero et al., 1993). Additionally, a number of EDCs stimulate in vitro Vg synthesis. Several studies indicate that alkylphenolic compounds are estrogenic based on the ability to stimulate Vg synthesis from fish hepatocytes (Jobling and Sumpter, 1993; White et al., 1994; Jobling et al., 1996). Alkylphenols are environmentally persistent and the parent compounds, alkylphenol polyethoxylates, are widely used as nonionic surfactants in detergents, paints, herbicides, pesticides, and many other formulated products. Over 300,000 tons of alkylphenol polyethoxylates are produced worldwide annually, making this the second largest group of nonionic surfactants in commercial production (Chemical Manufactures Association, 1994). It is estimated that 60% of these manufactured compounds end up in the aquatic environment after sewage treatment (Giger et al., 1987). Therefore, the occurrence of Vg in the





21


plasma of male rainbow trout (Oncorhynchus mykiss) (Purdom et al., 1994), carp (Cyprinus carpio) (Folmar et al., 1996), and flounder (Platichthysflesus) (Lye et al., 1997) following exposure to sewage effluent could be mediated, in part, by alkylphenolic compounds such as nonylphenol and octylphenol.

Vitellogenin appears to be a good in vitro and in vivo indicator of exposure to

estrogenic contaminants in many instances. However, the absence of Vg does not indicate the absence of endocrine disruption, as many compounds could act through mechanisms other than the estrogen receptor (see Cellular and Molecular Mechanisms of Disruption).

Vitellogenin could also provide a route of exposure for metal accumulation in

oocytes. Recent work on red drum (Sciaenops ocellatus) has discovered that Vg in vivo readily binds calcium, magnesium, zinc, iron and copper (Ghosh and Thomas, 1995). Furthermore, Ghosh and Thomas (1995) found that cadmium-vitellogenin injections resulted in cadmium incorporation into the ovaries. Thus, vitellogenic animals could be particularly sensitive to metal exposure. The possible role of Vg in transporting other xenobiotics into the cytoplasm of eggs is unknown. Thyroid

The thyroid is often thought of as exclusively associated with the control of

metabolism, but thyroid hormones are involved in numerous reproductive events, varying from stimulation of gonadal maturation in fish (Leatherland, 1994) and amphibians (van Teinhoven, 1983) to regulation of metamorphosis in amphibians or migration behavior in birds (van Teinhoven, 1983). Like steroid hormones, thyroid hormones are regulated by the hypothalamic-pituitary axis. The hypothalamus secretes thyrotropin-releasing hormone, which stimulates thyroid-stimulating hormone secretion from the anterior





22


pituitary (adenohypophysis). Thyroid-stimulating hormone then travels via the circulatory system to the thyroid where iodine is actively transported into the gland and used to synthesize thyroxine (T4) and triiodothyronine (T3).

Hypothyroid animals have low serum thyroid hormone concentrations which result in spontaneous abortions, stillbirths, or congenital defects (Burrow, 1986). Animals said to be hyperthyroid have excessive thyroid hormone production which can induce amenorrhea (loss of reproductive cyclicity) and reduced fertility (Burrow, 1986). Hyperthyroid animals have altered sex-hormone profiles as a result of increased levels of sex-hormone binding globulin (Tulchinsky and Chopra, 1973) and increased concentrations of circulating estrogens (Akande and Hockaday, 1972). Because sexhormone binding globulin (SHBG) has a greater affinity for testosterone than 170estradiol, an increase in SHBG causes relatively more 170-estradiol to be free and active.

Low reproductive success of herring gulls (Larus argentatus) in Lake Ontario was originally attributed solely to eggshell thinning, but further investigation revealed that other pollutant-induced factors contributed to the reported reproductive failure (Mineau et al., 1984). Among these factors, thyroid dysfunction was credited with altering normal reproductive behaviors such as egg incubation. Surveys of Lake Ontario herring gulls revealed several signs of hypothyroidism: decreased thyroid colloid content, increased thyroid epithelial cell height, decreased thyroid follicular diameter, and decreased circulating T3 and T4 concentrations (Rattner et al., 1984). Moccia et al. (1986) obtained similar results for herring gulls from other colonies in the Great Lakes basin, and both Rattner et al. (1984) and Moccia et al. (1986) attributed the hypothyroidism to a forage fish-borne goitrogenic etiology rather than iodine deficiency. In ovo hypothyroidism has





23


been noted in cormorants (Phalacrocorax carbo) exposed to PCBs, dibenzo-p-dioxins (PCDDs), and dibenzofurans (PCDFs), and this state could have a role in the observed low breeding success of cormorant colonies in the Netherlands (van den Berg et al., 1994). Embryos and juveniles could be particularly sensitive to contaminant-induced thyroid dysfunction, as developmental exposure to low levels of PCBs can cause organizational disruption of the thyroid (Gray et al., 1993a; Goldey et al., 1995).

Some of the effects of EDCs on the thyroid are activational. Brouwer et al. (1989) found that common seals (Phoca vitulina) fed PCB-contaminated fish had significantly lower T3 and T4 concentrations compared to seals fed non-contaminated fish. Similarly, fish (Oreochromis mossambicus) exposed for 20 days to 1,2,3,4,5,6hexachlorocyclohexane (BHC) displayed goiter formation, decreased colloid, and atrophied follicles, but these characteristics normalized when BHC-exposed fish were subsequently exposed to clean water (Pandey and Bhattacharya, 1991). The effects of particular pesticides on thyroid function can be complex. For instance in the freshwater catfish (Clarias batrachus), endosulfan induces a decrease in T3 and an increase in T4, malathion causes decreased T3 and no change in T4, and carbaryl provokes an increase in T3 and a decrease in T4 (Sinha et al., 1991). What direct effects these alterations in plasma T3 and T4 have on reproductive activity have not been reported. Cellular and Molecular Mechanisms of Disruption

There are numerous mechanisms through which contaminants can alter the

hormonal milieu in an animal. If altered, each mechanism can produce a unique disruption profile. For example, consider two well defined case studies of animals exposed to EDCs: the white sucker fish (Calostomes commersoni) of the Great Lakes and the American





24


alligators (Alligator mississippiensis) in Florida. White sucker fish exposed to bleached kraft mill effluent (BKME) have reduced circulating concentrations of 17-estradiol (E2) and testosterone (T) when compared to control fish (McMaster et al., 1991). These decreased hormone concentrations are correlated with a lower reproductive success of BKME-exposed fish. McMaster (1995) found that the decreased E2 and T concentrations could be attributed partially to decreased steroidogenic enzyme activity (aromatase) and a reduction in the cholesterol substrate used as the precursor of steroids.

Alligators from Lake Apopka, Florida, also show decreased reproductive success apparently as a function of endocrine disruption, but the hormonal profile and mechanisms of disruption differ from those of white sucker fish. Lake Apopka is contaminated with dicofol, DDT, DDD (1, 1-dichloro-2,2-bis(p-chlorophenyl)ethane), and p,p'-DDE (U.S. EPA, unpublished report), and alligator eggs from Lake Apopka contain significant residues of toxaphene, dieldrin, p,p '-DDE, p,p '-DDD, trans-nonachlor, and PCBs (Heinz et al., 1991). Reproductive success is dramatically lower on Lake Apopka compared to other Florida lakes (Woodward et al., 1989; Woodward et al., 1993; Masson, 1995). Plasma samples from juvenile Lake Apopka alligators reveal an altered steroid profile: males have lower T and females have higher E2 when compared with control animals (Guillette et al., 1994). In vitro culture of gonads found that testes from Lake Apopka males secreted more E2 but did not differ from controls in T synthesis (Guillette et al., 1995b). Additionally, female ovaries generated significantly less E2 compared to controls. Therefore, altered gonadal steroidogenesis does not explain the plasma steroid abnormalities in Apopka alligators, and the contaminant-induced alteration differs from that of the Great Lakes white sucker fish.





25


To explain the variation in particular case studies, we must examine the

mechanisms through which contaminants can alter the endocrine system. Contaminants could disrupt normal endocrine function by altering (a) the hypothalamic-pituitary axis of endocrine control, (b) the activity of steroidogenic enzymes, (c) the function of steroid binding molecules such as sex-hormone binding globulin, (d) the activity of hormone receptors by acting as a hormone agonist or antagonist, or (e) the hepatic clearance rate of steroids. The relationship between these functions is depicted in Figure 2-2.

In considering the ways that contaminants alter an animal's hormonal environment, first consider affects on the neuroendocrine control of steroid hormone secretion. The control of reproduction is an extremely complex integration of external stimuli into internal physiological signals. Basically, external stimuli are transduced through numerous neural pathways in the brain, and one of these pathways assimilates information in the hypothalamus. Here hypothalamic neurons secrete the neurohormone gonadotropinreleasing hormone (GnRH) into a vessel (hypothalamic-pituitary portal) that carries GnRH directly to the anterior pituitary gland. The anterior pituitary gland synthesizes gonadotropins (follicle stimulating hormone, luteinizing hormone) which are glycoprotein hormones that are released into the circulatory system. The gonadotropins travel to the gonads where they stimulate gametogenesis and sex-specific steroid hormone secretions that result in reproductive maturation (in the case of juveniles) or reproductive stimulation (in the case of adults). In a normal animal, feedback circuits control the secretion of gonadotropic hormones.
























Figure 2-2. Several components of the reproductive endocrine system with specific points where endocrine disruption can occur: (a) hypothalamic-pituitary axis, (b) gonadaladrenal steroidogenesis, (c) steroid receptors, (d) sex-hormone binding globulin production and interactions, and (e) hepatic steroid metabolism.





27


HYPOTHALMUS O GnRH ANTERIOR PITUITARY SBP-Steroid


SBP CIRCULATION

LIVER


Excretory Product Sex Steroid





GONAD
Intracellular Steroid Receptor

0
Cytoplasmic SBP





28


Toxicants that alter gonadotropin release will change an animal's steroid-hormone profile. For instance, chlordimeform (an acaricide recently banned in the United States but used in other parts of the world) is thought to influence endocrine regulation adversely by interfering with the activity of the hypothalamus (Goldman et al., 1990). Goldman et al. (1990) found that chlordimeform treatment in male rats decreased serum gonadotropins and decreased the hypothalamic GnRH response to norepinephrine stimulation. Because brain norepinephrine increases after a single dose of chlordimeform (Bailey et al., 1982), the endocrine-disrupting effects of chlordimeform are attributed to desensitization of the hypothalamic adrenergic receptors (Goldman et al., 1990). Many other EDCs could alter reproductive endocrinology by changing the hypothalamic and pituitary control of steroid production, but there is no data available for such alterations in wildlife species.

Steroid-hormone dynamics are dictated by a number of enzymes that convert

steroids. The steroidogenic pathway begins with cholesterol and, through the actions of numerous steroidogenic enzymes, progresses through progestins, androgens, and estrogens (see Figure 2-3). Induction or repression of the genes encoding steroidogenic enzymes can result in altered hormone production. Therefore, some of the endocrine disruption noted in wildlife could be due to altered transcription of genes encoding steroidogenic enzymes. This hypothesis is supported by several lines of evidence: first, the endocrine-disrupting effects of many contaminants appear to be mediated independent of steroid receptor binding (McLachlan, 1993) and, second, transcriptional control of steroidogenic enzymes is extremely labile during embryonic and neonatal periods (Gustafsson, 1994). Few studies have examined the activity and transcription rate of steroidogenic enzymes after EDC exposure, but it is clear that many compounds have the





29


potential to alter steroidogenesis in wildlife. For example in the American alligator, the actions of the enzyme aromatase (P450,.m; the enzyme responsible for converting androgens to estrogens) were blocked in vivo after embryonic exposure to several aromatase inhibitors (Lance and Bogart, 1992). The treated alligators had inhibited ovarian development, but were not masculinized.


Cholesterol




IF3HSD P45Pc21 P450c11
17HPregnenolProgesten 7 Deoxycorivosterom - Cortcosteron
Iso erases
P450c17 P450c17

IF 30-HSD IFP450021 P450c11
17-OH Pmrteeom--j0 17-OH Progesteror +11 Deoxycodisot - 10 Cortisol
Isomnerases
P450c17 P450c17
IF 3 -HS I 17 etqP450 aromn
Dehydroepiandrostemi Androstenedior Testoste 17p-Fstradiol
Isomerases Reductase
Figure 2-3: Steroidogenic pathway illustrating the P450 enzymes (bold) associated with the production of specific steroid hormones. Based on Miller (1988).


After steroid hormones are produced and secreted, they circulate throughout the body attached to carrier molecules. The most prominent of these molecules is sexhormone binding globulin (SHBG) which is produced by the liver. SHBG protects the steroid hormone from hepatic degradation and helps direct the hormone to target organs (Hammond and Boccinfuso, 1995). When attached to SHBG, steroid hormones are





30


biologically inactive. Sex-steroid binding proteins are found in animals from all vertebrate classes, but little consideration has been given to the effects of EDCs on SHBG presence and function. One effect of antiestrogenic drugs is to increase the circulating level of SHBG. Droloxifene, an antiestrogenic drug used to treat breast cancer, increases the plasma concentration of SHBG (Geisler et al., 1995). An increase in SHBG concentration causes more native hormone to be bound in a biologically inactive form, thereby decreasing the amount of steroid hormone that reaches the inside of a cell. Little attention has been given to the binding of environmental EDCs to SHBG and other plasma proteins. If certain EDCs do not bind to SHBG, these contaminants would have an elevated availability to target cells compared to native steroids. For instance, even though the affinity of the xenoestrogen o,p'-DDT for the estrogen receptor is 1000-fold lower than E2, the estrogenicity of o,p'-DDT would be enhanced if it did not bind to SHBG. Indeed, Arnold et al. (1996c) found that purified human SHBG selectively decreased the transcriptional activity of E2 compared to the xenoestrogens o,p '-DDT and octylphenol. The addition of alligator or human serum mimicked the results of purified SHBG, suggesting that SHBG and other proteins in serum selectively bind E2 but do not bind certain EDCs to a similar extent (Arnold et al., 1996c). Therefore, the binding of EDCs to plasma proteins such as SHBG is a major determinant of the EDC's bioavailability and potency in target cells.

Contaminants could also disrupt endocrine function by binding directly to hormone receptors. Steroid hormones are derivatives of cholesterol and easily diffuse through cellular membranes. Once inside the nucleus, steroid hormones bind to a hormone-specific receptor. The hormone-receptor complex binds directly to DNA, activating the





31


transcription of hormone-induced genes. Many environmental contaminants can bind directly to hormone receptors, stimulating (hormone agonist) or blocking (hormone antagonist) the expression of hormone-induced genes (McLachlan, 1993). A recent study indicates that combinations of relatively weak environmental estrogens can have a synergistic effect on the activation of the estrogen receptor (Arnold et al., 1996b). These complex interactions between multiple environmental contaminants and steroid receptors are not yet understood and are currently an area of interest and importance.

Although steroid hormones often are converted and "recycled" into other steroid hormones (see Figure 2-3), steroids are also excreted after hepatic conversion to hydrophilic compounds (primarily through Phase II conjugation reactions). For instance, men excrete approximately 50 pg of testosterone conjugates per day in the urine (Lipsett, 1986). Based on this principal, urine and fecal samples are used routinely for analysis of reproductive status in numerous animal species. An alteration in normal steroid-hormone excretion rates will alter the hormonal profile of an animal, and many environmental contaminants cause such alterations. It has long been recognized that the liver is the primary site of xenobiotic metabolism and steroid metabolism. Pioneering work by Conney and Klutch (1963) suggested that the administration of many drugs could alter the metabolism of in vivo steroids by altering hepatic enzyme systems. Welch et al. (1967) found that insecticides could also alter steroid metabolism in the rat liver. Specifically, organic phosphorothionate insecticides, such as parathion and malathion, inhibit liver microsomal hydroxylation (and, thus, excretion) of testosterone, whereas halogenated hydrocarbon insecticides, such as DDT and chlordane, stimulate the hydroxylation of steroids. In the case of chlordane, Levin et al. (1968) found that this decreased the





32


stimulatory effects of androgens on seminal vesicle weight by enhancing androgen metabolism. Welch et al. (1971) found similar metabolism alterations in females. Rats and mice exposed to a number of halogenated hydrocarbon insecticides (chlordane, dieldrin, heptachlor, lindane, p,p'-DDE, p,p '-DDD, or toxaphene) had stimulated hepatic metabolism of estrone and a decrease in uterine wet weight. These reactions to contaminants appear to be phylogenetically conserved, as Peakall (1967) noted that DDT and dieldrin exposure increased the excretion of testosterone's and progesterone's polar metabolites in birds.

Many of the hepatic alterations induced by environmental contaminants could be sex specific. For example, research in rats indicates that embryonic steroid exposure establishes sexual dimorphism of steroid-metabolizing enzymes in the liver (Lucier et al., 1982; Lucier et al., 1985). Male and female rodents exhibit different patterns of hepatic steroid metabolism, and these patterns are organized during development by exposure to androgens (Gustafsson, 1994). For instance, gonadectomy of males results in a femalepattern of steroid degradation, whereas administration of testosterone to a gonadectomized male results in male-pattern steroid degradation (Jansson et al., 1985). It is clear that sex steroids affect liver metabolism, such that exogenous androgens can masculinize a female liver and exogenous estrogens can feminize a male liver (Gustafsson, 1994). It is therefore plausible that exposure to EDCs, many of which mimic sex hormones, could alter normal hepatic sexual dimorphism.





33


Conclusions


In considering the effects of EDCs on wildlife reproduction, we have explored topics ranging from evolutionary considerations to molecular mechanisms of disruption. After briefly considering the topic at these scales, one theme emerges - many anthropogenic compounds released into the environment have adverse effects on the development and function of the reproductive system of wildlife species. Future studies should focus on the specific mechanisms of disruption, giving consideration to "mechanisms" from gene to ecosystem. Only with a better understanding of these mechanisms can progress be made in the discipline of endocrine toxicology.










Table 2-1. Representative reproductive abnormalities for wildlife exposed to EDCs. Alterations are categorized based on whether the abnormality was induced by experimental exposure (Experimental Manipulation) or detected from animals in the wild (Descriptive Observation).


Species


Experimental Manipulation Descriptive Observation


EDC


Ref


Fish
Catfish



Lake whitefish
White sucker Rainbow trout

Mosquitofish
Sole


t T3, 4 T4
4 T3
tT3, 4-T4 4-Vg


t ER' and Vg mRNA
1 Vg in males, 4 testis growth masculinization of females


Flounder


Freshwater perch Atlantic croaker


oocyte atresia
4 ovarian and plasma E2

ovarian atresia,- follicle growth
4 germinal vesicle breakdown


Endosulfan Malathion Carbaryl
Endosulfan 4-E2, 4T BKME
- E2, 4 T, 4 gonad size, T liver size BKME nonylphenol, PCBs alkylphenolics2 masculinization of females kraft mill effluent
4 E2, - vitellogenesis Puget Sound,?
4 ovarian production of E2 Puget Sound, ?
4 fertilization success San Francisco Bay, ?
1 Vg in males; malformed testes Sewage effluent Carbofuran metacid-50, carbaryl 4 gonad size (compared to control site) PAHs, PCBs, metals oil, napthalene o,p '-DDD, Kepone


Amphibians
African clawed frog Vg synthesis in males Reptiles
Red-eared turtle Sex reversal; male to female Vg synthesis in males


o,p '-DDT PCBs o,p '-DDT


(Sinha et al., 1991) (Sinha et al., 1991) (Sinha et al., 1991) (Chakravorty et al., 1992) (Munkittrick et al., 1992a) (Munkittrick et al., 1994b) (Flouriot et al., 1995) (Jobling et al., 1995) (Davis and Bortone, 1992) (Johnson et al., 1988) (Johnson et al., 1993) (Spies and Rice, 1988) (Lye et al., 1997) (Sukumar and Karpagaganapathy, (Choudhury et al., 1993) (Hontela et al., 1995) (Thomas and Budiantara, 1995) (Ghosh and Thomas, 1995)

(Palmer and Palmer, 1995)

(Bergeron et al., 1994) (Palmer and Palmer, 1995)









Table 2-1 (continued) Species Experimental Manipulation


Descriptive Observation


American alligator


Birds
Gull Gull
Marine birds
Parakeet
Munia Quail Quail
White king pigeon Mammals
Dall's porpoise
Beluga whale

Sea lion
Oldfield mice
River otter


males have 4 T, females have 4 E2 polyovular follicles, polynuclear oocytes testis produces T E2; ovary I E2


development of right oviduct

4 testes weight, ec4 degeneration
4 # st6 with good germ cells atrophic testes, abnormal sperm t oviducal secretary granules t hepatic steroid metabolism


4 fertility


fewer males in colony eggshell thinning


4T
4 follicular activity, mammary carcinoma mammary gland lesions (36% of females) premature births

4 baculum length and weight


Lake Apopka' Lake Apopka' Lake Apopka3


DDT
DDT
DDT
quinalphos5 quinalphos kepone kepone DDT, dieldrin


p,p'-DDE, PCBs DDT, mirex, PCPs many EDCs' organochlorines PCBs (Aroclor 1254) dioxin-like cpds


Guillette et al., (Guillette et al., (Guillette et al.,


1994) 1994) 1995b)


(Fry and Toone, 1981) (Fry and Toone, 1981) (Burger et al., 1995) (Maitra and Sarkar, (Maitra and Sarkar, (Eroschenko and (Eroschenko and (Peakall, 1967)

(Subramanian et al., (Beland et al., 1992) (Beland et al., 1993) (Delong et al., 1973; (McCoy et al., 1995) (Henny et al., 1996)


Florida panther abnormal sperm, 4 sperm, cryptorchidism PCBs, p,p '-DDT, (Facemire et al., 1995)
ER = estrogen receptor
4-nonylphenol, nonylphenoxycarboxylic acid, 4-tert-octylphenol, and nonylphenoldiethoxylate Numerous EDCs have been found in Lake Apopka animal tissues and eggs including: of toxaphene, dieldrin, p,p '-'-DDE, p,p '-'-DDD, trans-nonachlor, and
PCBs
4EC = epithelial cell
5Quinalphos is a commonly used organophosphorus pesticide. 6ST = seminiferous tubules
The following EDCs were found to be higher in beluga samples obtained from the St. Lawrence estuary compared to arctic belugas: mercury, lead, PCBs,
DDT, Mirex, benzo[(x]pyrene metabolites, dioxin equivalents, furans, and PAH metabolites.


Decinie beraio D Upf


EDC


Ref













CHAPTER 3
SEX-STEROID AND THYROID HORMONE CONCENTRATIONS IN JUVENILE ALLIGATORS (ALLIGATOR MISSISSIPPIENSIS) FROM CONTAMINATED AND REFERENCE LAKES IN FLORIDA

Introduction


Growth and reproduction are two of the fundamental variables that dictate the life history strategy of all animals. In a simplistic model, animals must survive and grow until they reach an optimal size that maximizes reproductive output (Stearns, 1992). Both growth and reproduction are under hormonal control and, as such, maintaining normal hormone concentrations is critical. There are numerous hormones that are involved in the complex regulation of somatic growth and development. Among these, the thyroid hormones appear to have a pivotal role in the growth of all vertebrates (Norris, 1997). The thyroid hormones thyroxine (T4) and triiodothyronine (T3) are necessary for normal differentiation, maturation, and growth of many systems including the central nervous system and skeletal systems (McNabb and King, 1993). Thyroid hormones also have a cooperative role in regulating the reproductive activities of vertebrates, but sex hormones are the most potent regulators of reproductive cyclicity in vertebrates (Norris, 1997). Testosterone (T) and estradiol- 17p (E2) are particularly important in regulating the development and function of reproductive activity and behavior in both sexes.

Several recent case studies have noted altered reproductive activity as a result of exposure to environmental contaminants that change the hormonal regulation of growth Note: This chapter is published in Environmental Toxicology and Chemistry (Crain et al., 1997 a)
36





37


and reproduction. Numerous studies have been conducted on white sucker (Catostomus commersoni) to determine the causes and mechanisms of reduced reproductive fitness of these fish. Depressed circulating steroid levels are noted in white suckers exposed to pulp mill effluent (McMaster et al., 1991; Munkittrick et al., 1991; Gagnon et al., 1994a), and these reductions correlate with reduced gonadal development, reduced expression of secondary sexual characteristics, delayed maturity, decreased egg size, and reduced fecundity with age (Gagnon et al., 1994b; Munkittrick et al., 1994a; Munkittrick et al., 1994b). A similar correlation between contaminant exposure and abnormal reproductive function has been noted for alligators (Alligator mississippiensis) from Lake Apopka, Florida, USA. The alligator population on Lake Apopka precipitously declined from 30 juveniles/km to 4 juveniles/km between 1980 and 1983, and the mean clutch viability (eggs hatched / eggs laid) dropped from 54% in 1983 to 13% in 1986 (Woodward et al., 1993). These declines followed a 1980 spill of a pesticide mixture that was primarily composed of dicofol but also had as much as 15% DDT, DDD, and DDE (Environmental Protection Agency, 1994). Recent studies have indicated that plasma steroid hormone concentrations are abnormal in the alligators of Lake Apopka (Guillette et al., 1994; Guillette et al., 1996b; Guillette et al., 1997), and these abnormal steroid concentrations correlate with altered gonadal morphology (Guillette et al., 1994) and smaller phallus size (Guillette et al., 1996b; Guillette et al., 1997) in juvenile alligators. The hormonal abnormalities in the juvenile alligators of Lake Apopka include depressed T in males (Guillette et al., 1994; Guillette et al., 1996b; Guillette et al., 1997) and elevated E2 concentrations in females (Guillette et al., 1994). Whereas males and females have both T and E2 circulating in their blood, it is the relative ratio of the two steroids that dictates





38


reproduction (Bogart, 1987). For instance, females have more E2 relative to T and this is thought to affect ovarian differentiation as well as female reproductive function. Many environmental chemicals can alter normal steroid concentrations by changing hypothalamic-pituitary control of steroid synthesis, gonadal steroid synthesis, hepatic steroid conversion and excretion rates, and receptor-mediated responses (see CHAPTER

2 for review).

This study examines plasma concentrations of the sex-steroids E2 and T and the

thyroid hormones T3 and T4 in three populations of juvenile American alligators in Florida. Two of these populations are in historically contaminated lakes, Lake Apopka and Lake Okeechobee. The third population is in the relatively pristine environment of Lake Woodruff National Wildlife Refuge. Previous studies have noted altered circulating concentrations of steroid hormones in the alligators of Lake Apopka. The purposes of this study are to: (1) reevaluate circulating steroid hormones in Lake Apopka alligators, relative to the reference lake and another contaminated lake, (2) analyze, for the first time, circulating thyroid hormone concentrations in alligators from these three lakes, and (3) evaluate the steroid and thyroid hormone concentrations relative to body size. If alligators exposed to environmental contaminants are affected by endocrine disrupters, this may be evident in the circulating concentrations of sex steroids or thyroid hormones.





39


Materials and Methods


Study Sites and Sample Collection


The three lakes in this study were chosen based on historical and ongoing episodes of exposure to environmental contaminants. Lake Woodruff (lat. 29*06'N, long. 81*25'W; sampled on April 25, 1995) is in Lake Woodruff National Wildlife Refuge and is considered a relatively pristine, natural environment. Lake Apopka (lat 28040'N, long 81*38'; sampled on May 2, 1995) is 1.5 miles downstream from an EPA Superfund site where an industrial spill of dicofol and DDT occurred in 1980 (Environmental Protection Agency, 1994). Lake Apopka has also received substantial agricultural and municipal runoff (Sengal and Pollman, 1991; Schelske and Brezonik, 1992) and several studies have documented reproductive endocrine disruption in the alligators of Lake Apopka (Guillette et al., 1994; Guillette et al., 1996b). Lake Okeechobee (lat. 26'56'N, long. 80049'W; sampled on May 3, 1995) receives significant agricultural runoff that has led to eutrophication (Federico et al., 1981), and Lake Okeechobee is known to be contaminated with mercury (Jurczyk, 1993; Sundlof et al., 1994) and pesticides (Turnbull et al., 1989), but endocrine-disruption has not been studied in the alligators of Lake Okeechobee.

Juvenile American alligators (Alligator mississippiensis) from 60-140 cm total

length were hand captured from an airboat at night. Animals of this size are 2-6 years of age (Woodward et al., 1992). To minimize temporal effects of sampling, all samples were collected during a 9-day time span. A blood sample (1-2 ml) was collected in a heparinized vacutainer from the post-cranial sinus within 30 min of capture. A previous study found that capture stress has no influence on the acute plasma steroid concentrations





40


in alligators (Guillette et al., 1997). Individuals were sexed using the phallic criteria previously defined (Allsteadt and Lang, 1995; Guillette et al., 1996b), and body measurements (snout-vent length and total length) were recorded. Animals were released in the vicinity of their capture. Blood samples were stored on ice for 10-12 hrs until centrifugation at 1500 g for 15 min. Plasma was stored at -72*C until hormone analysis.



Steroid Hormone Radioimmunoassays


E2 and T were analyzed using radioimmunoassays previously validated for alligator plasma (Folmar et al., 1996; Guillette et al., 1997). Briefly, plasma (100 p) was extracted 2x with diethyl ether (5 ml). The ether extracts were dried under constant air stream for 15 min. The dried samples were resuspended with borate buffer (100 pl; 0.5 M; pH=8.0). To reduce non-specific binding, 100 pl of borate buffer with bovine serum albumin (Fraction V; Fisher Scientific) at a final assay concentration of 0. 15% for T and 0.19% for E2 was added to each tube. Antibody was then added (200 p1; final concentration of 1:55,000 for E2, 1:25,000 for T; Endocrine Sciences). Finally, 100 P1 of radiolabelled steroid was added (12,000 cpm per 100 p; [2,4,6,7,16,17-3H]Oestradiol @ I mCi/mi; [1,2,6,7-3H]Testosterone @ 1 mCi/ml; both from Amersham International, Arlington Heights, IL). For standard tubes, either T or E2 was added at 0, 1.56, 3.13, 6.25, 12.5, 25, 50, 100, 200, 400, and 800 pg/tube. Tubes were vortexed for 1 min and incubated overnight at 4'C. All standards and samples were prepared in duplicate.

Bound-free separation was accomplished by adding 500 pl 5% charcoal - 0.5%

dextran and immediately centrifuging at 1500 g at 4'C for 30 min. 500 p was then added





41


to 5 ml scintillation cocktail and the tubes were counted on a Beckman scintillation counter.

The steroid RIAs were validated using both plasma dilutions and internal

standards. For internal standards, 100 pLl of steroid-free plasma (steroids removed by incubating plasma with activated charcoal, 10% w/v, and collecting the supernatant after centrifugation at 10,000 g for 15 min) was spiked with steroid (6.25, 12.5, 25, 50, 100, 200, 400, 800 pg for T; 3.125, 6.25, 25, 50, 100, 200 pg for E2). The internal standards were extracted and assayed as described above. For the T plasma dilutions, 3, 6, 12, 25, 50, 75, and 100 pl of plasma pool were placed in separate tubes, and the volumes were brought to 100 d with steroid-free plasma. For the E2 plasma dilutions, 25, 50, 75, and 100 pl of the plasma pool were pipetted into separate tubes and the volumes were brought to 100 pl with steroid-free plasma. The plasma dilutions were extracted and assayed as described above. All standards, internal standards, and plasma dilutions were performed in duplicate. Parallelism of the internal standard curve and plasma dilution curve with the standard curve was tested using the test for homogeneity of slopes (SuperAnova, Abacus Concepts, Berkeley CA). Figures 3-1 and 3-2 present the validation curves for the steroid hormones.




42


120

-n- aaxad Ome 100 1 -A- Rasmn[Dluicns
+o Irtenal Sardards 80
CD A

0

40






1 10 100 1000
Tesosmerne (pj Figure 3-1 Radioimmunoassay validation for the measurement of testosterone in alligator plasma. The slopes of the internal standards and plasma dilutions were not significantly different from the standards (p=O.184).





43


x
0


1201008060

4020-


n


1 10 100 1000

Esradd (pg)

Figure 3-2 Radioimmunoassay validation for the measurement of estradiol in alligator plasma. The slopes of the internal standards and plasma dilutions were not significantly different from the standards (p=0.73 1). Thyroid Hormone Radioimmunoassavs


Unextracted plasma (100 pl) was used for determination of circulating

concentrations of T3 and T4. Assay buffer (100 pl of 0.2 M borate buffer; pH=8.0) was added to each tube, followed by the addition of 200 pl of BSA/y-globulin/ANS buffer (1% bovine serum albumin, 1.25 mg/ml y-globulin, 2 mg/ml 8-alinino-1-napthalene-sulfonic acid; all from Sigma Chemical Co., St. Louis, MO). Antibody (100 pl) was added to give


-a- Sandard Oie
-m+ Pmasr DIldons
-o- Irtemd Standards





44


a final concentration of 1:8,000 for T3 antisera and 1:2,000 for T4 antisera (Endocrine Sciences, Calabasas Hills, CA). Finally, 100 ptl of iodinated T3 (40,000 cpm/tube) or T4 (50,000 cpm/tube) was added to the tubes (both from New England Nuclear; L-[125I]thyroxine @ 1250 pCi/4g; L-3,4,3'-[125I}-triiodothyronine @ 1200 pCi/pg). Standards were prepared in 100 pl of assay buffer and substituted for this constituent in the assay described above. For T4, 0, 25, 50, 100, 200, 400, 800, 1600, 3200, 6400, and 12800 pg/tube were prepared, whereas 0, 6.25, 12.5, 25, 50, 100, 200, 400, 800, 1600, and 3200 were prepared for the T3 assay. Tubes were vortexed and incubated at 37'C. After 2 hrs, the tubes were incubated at room temperature for 1.5 hr. Bound-free separation was accomplished by adding 1.5 ml of 60% saturated ammonium sulfate to each tube, vortexing, and centrifuging at 1500 g for 30 min. The supernatant (containing the free hormone) was discarded and the pellet resuspended in a 9:11 mixture of saturated ammonium sulfate and assay buffer with 0.5% BSA. The tubes were then vortexed, centrifuged at 1500g for 30 min, and the supernatant discarded. The pellets were counted on a Beckman gamma counter.

The thyroid hormone RIAs were validated using plasma dilutions. Plasma (25, 50, 75, and 100 pl) was aliquated from a plasma pool, and the dilutions were extracted and assayed as described above. Parallelism of the internal standard curve and plasma dilution curve with the standard curve was tested using the test for homogeneity of slopes (SuperAnova, Abacus Concepts, Berkeley CA). Figures 3-3 and 3-4 present the validation curves for the steroid hormones.





45


100

80 ---Standard Curve
-A- Plasma Dilutions
0
60
x

_M 40

20

0
1 10 100 1000 10000

Triiodothyronine (pg)

Figure 3-3 Radioimmunoassay validation for the measurement of triiodothyronine (T3) in alligator plasma. The slope of the plasma dilutions was not significantly different from that of the standards (p=0.71 1).


100
-+-Standard Curve 80 - --&-Plasma Dilutions
0
S60
x

m 40

20

0
1 10 100 1000 10000
Thyroxine (pg)

Figure 3-4 Radioimmunoassay validation for the measurement of thyroxine (T4) in alligator plasma. The slope of the plasma dilutions was not significantly different from that of the standards (p=0.454).





46


Statistical Analysis


Hormone concentrations were determined using commercially available software (Microplate manager III; Biorad, Hercules, CA, USA). All statistical tests were performed with SuperAnova (Abacus Concepts; Berkeley, CA, USA). Mean body size was not significantly different for males (p=O. 19) or females (p=O.55) from the three lakes. Thus, an analysis of variance (ANOVA) was conducted to determine whether differences in hormone concentrations existed among alligators of the 3 lakes. Prior to this analysis, hormone concentrations were log transformed to obtain homogeneity of variance. Fisher's PLSD was used as a post-hoc test. To examine the relationship of body size to hormone concentrations, linear regression analysis was conducted for each sex on each lake.

Results


Mean Hormone Concentrations


All the animals sampled in this study represent one life history stage -juvenile. Therefore, we chose to first analyze the data by comparing hormone concentrations in alligators from the three lakes (see Table 1). There was no difference in mean E2 concentrations among males (F=1.62; df=2, 43; p=0.21) or females (F=0.50; df=2, 40; p=0.61) from the three lakes. Mean T concentrations were not different among females from the three lakes (F=1 .36; df-2, 40; p=0.27), but T concentrations were significantly different among males (F=4.96; df-2, 44; p=0.01). Woodruff males had significantly more T compared to males from either Apopka (p=0.009) or Okeechobee males (p=0.01).





47


For triiodothyronine (T3) concentrations, there was no difference among females (F=0.74; df-2, 40; p=0.48) or males (F=1.10; df-2, 43; p=0.34) from the 3 lakes. Thyroxine (T4) concentrations were significantly different among lakes for males (F=5.67; df=2, 43; p=0.007) but not for females (F=3.16; df=2, 39; p=0.053). For males, T4 concentrations were significantly lower in Woodruff animals compared to Okeechobee alligators (p=0.002).


Table 3-1. Mean plasma concentrations (+ ISE) of the thyroid hormones T3 and T4 (ng/ml) and the steroid hormones E2 and T (pg/ml) in male and female alligators from the three lakes.
Woodruff Apopka Okeechobee

male female male female male female

T3 2.1 0.4 1.6 0.2 1.7 0.3 2.3 0.4 2.2 0.3 2.0 0.3

T4 13.9 2.3 13.5 2.2 17.3 1.9 18.8 3.1 21.8 2.0 18.9 2.2

E2 14.7 3.9 25.4 4.8 21.6 5.1 23.9 7.1 12.0 2.6 28.7 4.3

T 962.6 283.9 86.4 10.1 162.7 71.8 86.5 10.1 218.8 108.0 66.6 11.2



Body Size and Hormone Concentrations


The relationship between body size and hormone concentrations was examined using regression analysis (see Table 3-2). For the steroid hormones, there was no apparent relationship between E2 and body size in males from Woodruff and Okeechobee, but a relationship was apparent in males from Apopka (see Figure 3-5). However, this relationship disappears (r2=0.03 1, p=0.25) if the one Apopka male that is larger than 105 cm is removed from the analysis. T was positively related to body size in males from





48


Okeechobee and Woodruff, whereas there was no such relationship in Apopka males (see Figure 3-6). Among females, a clear positive relationship existed between E2 and body size for Woodruff and Okeechobee animals, but females from Apopka had no apparent relationship between E2 and body size. T was positively related to body size in females from Woodruff, but T had no such relationship to body size in females from Apopka or Okeechobee.

For the thyroid hormones, there was a negative relationship between both T3 and T4 and body size for Woodruff males and females (see Figures 3-7 and 3-8). Additionally, T3was negatively related to body size in males from Apopka and Okeechobee, but females from these lakes showed no clear relationship. For T4, Apopka females showed a clear negative relationship with body size, whereas Apopka males and Okeechobee males and females exhibited no apparent relationship.





49


Table 3-2. Results for the linear regression analysis of hormone concentrations as a function of total body size in male and female alligators from the three lakes.
Woodruff Apopka Okeechobee

male female male female male female

r -0.827 -0.731 -0.493 -0.564 -0.627 -0.472

T3 r2 0.684 0.535 0.243 0.318 0.393 0.223

p <0.001 <0.001 0.039 0.089 0.012 0.088

r -0.662 -0.691 -0.122 -0.709 -0.358 -0.521

T4 r2 0.438 0.477 0.015 0.502 0.128 0.271

p 0.005 0.001 0.660 0.022 0.189 0.067

r 0.1 0.643 0.680 0.303 0.505 0.573

E2 r2 0.010 0.413 0.462 0.092 0.303 0.328

p 0.701 0.003 0.004 0.541 0.569 0.032

r 0.521 0.553 0.083 0.474 0.594 0.063

T r2 0.271 0.306 0.007 0.225 0.353 0.004

p 0.032 0.014 0.752 0.166 0.025 0.821





50


Figure 3-5. Relationship between estradiol-170 concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff. A clear relationship between body size and E2 concentration was detected in Okeechobee and Woodruff females, but not in Apopka females. The significant relationship between body size and E2 concentration in Lake Apopka males disappears if the one individual greater than 105 cm is removed from the analysis.





51


100* 8060

4020-


60


100


(


100


0


80


I a


100


120


140


I


60 80 100 120 140

, ..


60


80


100


120


140


Length (cm)


Lake Apopka A females -


A0


males .. I
-- -
0 0 0

0 0 0-& 0
0 C A A
-0i,6


U-I


755025-


0


Lake Okeechobee








0 A C0


7550


254


.-


Lake Woodruff
0



A0

0 0AO 0 000 0&


I I I


O


0





52


Figure 3-6. Relationship between T concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff Males from Woodruff and Okeechobee exhibited a relationship between body size and T concentration, but males from Apopka showed no such relationship.





53


-1 I


Lake Apopka


* females

0 males


C cJ2


3


2


1


0 -I


80


100


120


0


60


4


Lake Okeechobee




0


000--
0 - - .00 6 A()


60


80


100


120


1


Lake Woodruff


0 0 0


0 ..-6000
0 0 0M


60


80


100


120


Length (cm)


4


-II -


140


3


2


1


0-


bb



0
0


40 40


4

3

2

1

0


V
C
'-4
V
C ci:~
V
H


I.


----


1





54


Figure 3-7. Relationship between triiodothyronine concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff. Both males and females from Woodruff exhibited a negative relationship between T3 concentration and body size, whereas there was no apparent relationship for females from Apopka and Okeechobee.






55


Q' 6
-0 o Lake Apopka
bL0 5A a females o 4
-z 0- 0 m ales - -0
2
0 ,b

0_ 0
~~ 1 0 00


60 80 100 120 140




5- Lake Okeechobee

4- 1
OA
3- 00

00
2

00
0-,&



60 80 100 120 140




A 5-Lake Woodruff
3 0
& 0 04

60 80 100 120 140
6- 0


4-0



0 A
00
00
H60 80 100 120 140


Length (cm)





56


Figure 3-8. Relationship between thyroxine concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff. A clear relationship between T4 and body size existed for Woodruff males and females, but this relationship was not seen in males from Apopka or females from Okeechobee and Apopka.





57


50
Lake Apopka 40- A females

30- 0 males ---oC 0
20- 3 0

0 6 2

0
40 60 80 100 120 140

50
o Lake Okeechobee
4030


0
S10


60 80 100 120 140


50
Lake Woodruff
40- 0

I 3020

0

10
60 80 100 120 140


Length (cm)





58


Discussion


The results indicate differences in concentrations of both sex steroid and thyroid hormones among the alligators of Lake Apopka, Lake Woodruff, and Lake Okeechobee. Whereas there were no differences in E2 concentrations among animals of the three lakes, T concentrations in Lake Apopka and Lake Okeechobee male alligators were significantly lower than T concentrations in Lake Woodruff male alligators. Concentrations of T4 also differed in animals of the 3 lakes, with T4 concentrations being lower in Lake Woodruff males compared to Lake Okeechobee male alligators. In general, the thyroid hormones were inversely related to alligator size, whereas the steroid hormones exhibited a positive relationship to size. Deviations from this pattern were most pronounced in Lake Apopka animals, with a lack of any relationship between T4 or T and size in males, and a lack of any clear relationship between T3 or E2 and size in females. Additionally, body size had no correlation with plasma E2in males of Lakes Okeechobee or Woodruff, whereas Apopka males showed a positive correlation. This positive correlation is attributed to one large Apopka male, and it is unknown if this pattern is common among Apopka males of a similar size. Animals of this size class are rare on Lake Apopka due to the lack of emergent vegetation. Neither T3 nor T4 was correlated with alligator size in Okeechobee females, but there was a clear correlation in Okeechobee males. These results indicate differences in hormone concentrations among lakes, and these differences are dependent upon both the sex and the size of the alligator.

The differences noted between males and females support previous data suggesting that the specific effects of endocrine-disrupting chemicals can be gender specific.





59


Normally, juvenile reptiles exhibit sexual dimorphism in circulating steroid hormone concentrations. Juvenile male green sea turtles (Chelonia mydas) have higher circulating testosterone concentrations compared to females; in fact, this difference is the only means of determining the sex of these immature sea turtles (Wibbels et al., 1987). Similarly, 6month-old male alligators from Lake Woodruff have significantly more T than females, and Woodruff females have significantly more circulating E2 than males (Guillette et al., 1994). Guillette et al. (1994) found, however, that 6-month-old alligators from Lake Apopka did not conform to this pattern. Apopka males had low concentrations of T and Apopka females had higher E2 concentrations compared to animals from Lake Woodruff These results are from alligators reared in a control environment since hatching; thus, they represent a difference in the embryonic organization of the reproductive system (Guillette et al., 1995a). Results of the current study support these previous data, as Apopka males have a significantly lower mean concentration of T and a pattern of higher concentrations of E2 compared to Lake Woodruff males. These data suggest that the previously reported organizational alterations in 6-month-old alligators from Lake Apopka persist through the juvenile years and, in fact, may be more pronounced in larger (> 80 cm) juvenile alligators.

Depression of circulating T concentrations has been noted before in animals exposed to environmental contaminants. Adult male rats treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin exhibit reduced androgen concentrations (Moore et al., 1985), apparently due to inhibition of cholesterol mobilization early in the steroidogenic pathway (Moore et al., 1991). These results are similar to those found in white sucker (Catostomus commersoni) exposed to bleached kraft pulp mill effluent (BKME) in Lake Superior. After exposure to BKME, male white sucker exhibit decreased concentrations





60


of T and 11-ketotestosterone, and females exhibit decreased concentrations of T during both vitellogenesis and spawning (McMaster et al., 1991; Munkittrick et al., 1991). In populations of Dall's porpoises (Phocoenoides dalli) in the North Pacific, increasing concentrations of PCBs and DDE in the blubber are correlated with decreased circulating T levels (Subramanian et al., 1987). These altered androgen concentrations in rats, white sucker fish, and porpoises are likely activational; that is, the function of a normally organized reproductive system is altered due to exposure to an endocrine-disrupting agent (Guillette et al., 1995a). For example, normal goldfish (Carassius auratus) exposed to BKME demonstrate reduced circulating T concentrations within 4 days of exposure (McMaster et al., 1996). Most of the evidence for Lake Apopka animals, and possibly those of Okeechobee, does not suggest activational disruption, but that an organizational change has occurred (Guillette et al., 1995a). Exposure of developing embryos to hormone-disrupting contaminants can alter the normal development of endocrine organs such as gonads and thyroids. For example, several polychlorinated biphenyls (PCBs) cause gonadal sex reversal (from default male to female) in a turtle species with temperature-dependent sex determination (Bergeron et al., 1994). Similar sex reversal from male to female has been noted in embryonic alligators exposed to the natural estrogen estradiol- 170 (Lance and Bogart, 1992; Crain et al., 1997b), the estrogen agonist/antagonist tamoxifen (Lance and Bogart, 1991; Crain et al., 1997b), and several inhibitors of steroidogenic enzymes (Lance and Bogart, 1992). It is unknown if such sexreversed animals exhibit normal reproductive function later in life, and it is possible that the altered steroid and thyroid hormone concentrations noted in the present study are due to such organizational endocrine disruption.





61


Although the exact causative agents of endocrine disruption in Lake Apopka

alligators have not been identified, it is likely that the alterations in steroid hormones are due to embryonic exposure to a number of different compounds. As previously mentioned, Lake Apopka is adjacent to an EPA Superfund site. Due to a 1980 spill at the Tower Chemical Co. site, a pesticide mixture composed of dicofol, DDT, DDD, and DDE entered the lake (Environmental Protection Agency, 1994). Although these compounds are known endocrine disrupters, it is impossible to draw a direct cause-effect relationship between these contaminants and the hormonal disruption in Apopka's alligators because of the historical and current extensive agricultural practices surrounding Lake Apopka. Alligator eggs collected from Lake Apopka in 1984 and 1985 had relatively high concentrations ofpp'-DDE, pp'-DDD, toxaphene, dieldrin, and trans-nonachlor (see Table 3) (Heinz et al., 1991). A recent study showed that two of these compounds (p,pDDD and trans-nonachlor) exhibit binding to the alligator estrogen receptor (Vonier et al., 1996). Additionally, Arnold et al. (1996b) found that the combination of dieldrin and toxaphene actively competes for the human estrogen receptor, whereas each individual compound has no affinity for the receptor. This "synergistic" activation of the estrogen receptor emphasizes the potential for mixtures of contaminants to alter the organization of the reproductive system.

Circulating thyroid hormones were highly correlated to body size in male and female alligators from Lake Woodruff Similar strong relationships between thyroid hormone concentration and body size were noted for Lake Apopka and Okeechobee males for T3 and for Apopka females for T4. However, females from Apopka and Okeechobee showed little relationship between body size and T3, and males from Apopka and both





62


sexes from Okeechobee showed little relationship for T4. In reptiles and mammals, T3 is considered the active thyroid hormone, as it binds avidly to nuclear receptors resulting in the orchestration of metabolism and growth (Eales, 1990). However, T4 may be the most important circulating indicator of thyroid hormone status, as T4 is the primary hormone secreted by the thyroid. At the level of the cell, T4 is converted to T3 by deiodonase enzyme activity (Eales, 1990). Thus, both T4 and T3 concentrations are critical regulators of growth and metabolism, and both thyroid hormones can be differentially produced depending upon season and sex (Gancedo et al., 1995). These thyroid hormone concentrations can be modified by exposure to endocrine disrupters. For instance, brown trout (Salmo trutla) exposed to sublethal levels of aluminum exhibit elevated circulating concentrations of both T3 and T4 (Waring et al., 1996; Waring and Brown, 1997). Aluminum is thought to stimulate T4 release from the thyroid and increase hepatic monodeiodination of T4 to T3, thus resulting in elevated T3 and T4 in the plasma (Waring and Brown, 1997). Therefore, it is plausible that the elevated T4 concentrations in Lake Okeechobee males and the lack of correlation between body size and thyroid hormones in Lake Okeechobee and Lake Apopka animals could be attributed to exposure to endocrine disrupters.

The lack of correlation between the thyroid hormones and size in both Lake

Apopka and Lake Okeechobee animals could reflect altered reproductive potential in these animals, as the thyroid hormones cooperatively regulate the reproductive activities of vertebrates. For instance, thyroidectomy causes complete inhibition of spermatogenesis in the lizard Calotes versicolor and the gecko Coleonyx variegatus, and this inhibition is restored by administration of T4 (Plowman and Lynn, 1973; Haldar-Misra and Thapliyal,





63


1981). However, administration of T4 to normal, non-thyroidectomized geckos causes inhibition of spermatogenesis similar to that in thyroidectomized animals (Plowman and Lynn, 1973). These results suggest that any aberration in the concentration of circulating T4, whether it be increased or decreased, can have a substantial effect on spermatogenesis and, thus, male reproductive success. This corroborates previous suggestions that the gonads can only function normally within a limited range of thyroid activity (Eyeson, 1970). For this reason, contaminants that interfere with the function of the thyroid can be potent disrupters of reproduction. In Lake Apopka alligators, previous studies have documented reproductive abnormalities that are symptomatic of endocrine disruption (Guillette et al., 1994; Guillette et al., 1996b), but the pathways through which environmental chemicals elicit such abnormalities have not been fully elucidated (Guillette et al., 1995a). It is possible that the thyroid/gonad axis is involved, and future studies should examine the relationship between reproductive endocrine disruption, thyroid function, and environmental contaminants.


Table 3-3. Mean concentrations (ppm) of environmental contaminants measured in alligator eggs collected from Lake Apopka during 1984 and 1985.


Contaminant 1984 Concentration (ppm) 1985 Concentration (ppm)

Toxaphene 0.09 2.4

Dieldrin 0.24 0.11

p,p '-DDE 5.8 3.5

p,p'-DDD 0.82 0.37

trans-nonachlor 0. 11 0.15

Source: Data from Heinz et al. (1991) with permission.





64


In summary, this study has shown that (1) circulating testosterone is reduced in male Apopka and Okeechobee alligators relative to males from Lake Woodruff, (2) circulating thyroxine is elevated in Lake Okeechobee males compared to Lake Woodruff males, and (3) in general, body size is correlated with steroid and thyroid hormone concentrations in animals of Lake Woodruff, but this is not always the case in animals of Lake Apopka or Lake Okeechobee. Future studies of endocrine disruption in ectotherms should consider (1) the role of thyroid hormones in contaminant-induced reproductive alterations, (2) the potential of endocrine-disrupting contaminants to alter growth via thyroid hormone disruption, and (3) size-specific responses to endocrine-disrupting chemicals.













CHAPTER 4
TESTOSTERONE SYNTHESIS IN EMBRYONIC AND JUVENILE ALLIGATORS EXPOSED TO ENDOCRINE-ALTERING ENVIRONMENTAL CONTAMINANTS Introduction


Environmental contaminants can alter the reproduction and growth of animals by interfering with the normal functioning of the endocrine system. The endocrine regulation of reproduction appears particularly susceptible to perturbation by endocrine-disrupting contaminants (EDCs), as many EDCs affect synthesis, availability, and binding of reproductive hormones (Chapter 2). .Contaminant-induced endocrine disruption has been shown to alter the reproduction of fish (McMaster, 1995), amphibians (Palmer and Palmer, 1995), reptiles (Guillette and Crain, 1995), birds (Fry, 1995), and mammals (Subramanian et al., 1987; Beland et al., 1993). These endocrine-altering effects do not appear to be characteristic of any particular class of environmental contaminants, as endocrine alteration can be seen after exposure to agricultural, industrial, and municipal waste compounds (Colborn et al., 1993).

One of the best-described case studies of endocrine disruption in a wildlife species involves the American alligators (Alligator mississippiensis) of Lake Apopka, Florida (USA). Lake Apopka is 1.5 miles downstream from an EPA Superfund site (where a spill of dicofol, DDT, and other compounds occurred in 1980), receives significant agricultural runoff, and was previously used as a municipal sewage reservoir (Schelske and Brezonik, 1992). A study conducted in the late 1980s and early 1990s showed that both the number Note: This chapter is in review in Conparative Biochemisay and Physiology, Part C (Crain and Guillette, 1997b).
65





66


ofjuvenile alligators and the clutch viability are reduced on Lake Apopka compared to other lakes in Florida (Woodward et al., 1993), and alligator eggs taken from Lake Apopka in the mid 1980s had high residues of several organochlorines (Heinz et al., 1991). Juvenile alligators living in Lake Apopka exhibit a number of morphological modifications suggestive of endocrine disruption: (a) males on Apopka have smaller penis size compared to males on other lakes (Guillette et al., 1996b), (b) males on Apopka have poorly organized testes compared to those on other lakes (Guillette et al., 1994), and (c) females on Apopka have abnormal ovaries that exhibit polyovular follicles and polynuclear oocytes (Guillette et al., 1994). In addition to these morphological abnormalities, animals from Apopka have abnormal circulating concentrations of reproductive hormones. When compared to animals from a reference site, Apopka males have decreased circulating testosterone concentrations (Chapter 3) (Guillette et al., 1994; Guillette et al., 1997) and Apopka females have increased circulating estradiol concentrations (Guillette et al., 1994).

There are a number of ways that contaminants alter circulating hormone

concentrations, such as (a) directly affecting gonadal steroid production, (b) indirectly altering steroid production by changing the hypothalamic-pituitary control of gonadal steroid synthesis, (c) changing the normal hepatic excretion rate of steroids, or (d) affecting the amount of circulating "free hormone" by altering sex-steroid binding protein synthesis. The purpose of this study is to test the first of these possibilities: that contaminants can directly alter gonadal steroid production. A descriptive and an experimental study are utilized to assess the effects of contaminants on gonadal testosterone synthesis. This approach is taken to further elucidate the mechanism(s) of endocrine disruption in alligators exposed to EDCs.





67


Materials and Methods


Descriptive Study

Eggs were collected from Lake Woodruff National Wildlife Refuge, Florida, and Lake Apopka, Florida, under permit from the Florida Game and Freshwater Fish Commission. Eggs were incubated at 31 C, a temperature that produces mostly females (Lang and Andrews, 1994). After hatching, neonates were transported to the Santa Fe Teaching Zoo (Gainesville, Florida) where they were housed in an outdoor semi-aquatic enclosure. Animals were fed ad libitum daily with a commercial alligator chow (Burris Mill and Feed, Inc., Franklinton, LA). At 9 months of age, animals were sexed by examination of phallus development (Allsteadt and Lang, 1995), and the animals were transported to our lab for examination. Five males and 5 females were examined for Lake Apopka hatchlings, whereas 3 males and 7 females were examined for Lake Woodruff hatchlings. We were unable to examine more male alligators from Lake Woodruff due to a lack males in this cohort group.

Experimental Study

Eggs were collected from 5 nests at the reference site, Lake Woodruff, and transported to our laboratory. One egg from each nest was opened to determine the developmental stage of the embryo (staging based on criteria defined by Ferguson (1985)). Eggs from each clutch were separated equally among 5 groups (15 eggs per group), and each group was exposed to a particular treatment at developmental stage 21, just prior to the onset of sexual differentiation (Lang and Andrews, 1994). Table 1 summarizes the experimental design. Two of the treatment groups were controls (one group incubated at





68


a male-producing temperature - 33*C, and one group incubated at a female-producing temperature - 300C), one group was an endocrine-disrupting standard (0. 1 ppm estradiol17p), and two groups were exposed to common environmental contaminants (5 ppmp,p'DDE or 5 ppm p,p '-DDD, both metabolites of DDT). Except for the control females, all eggs were incubated at 33*C, a temperature that normally produces 100% males (Chapter 4, Chapter 5) (Lang and Andrews, 1994). Five embryos from each treatment group were sampled at each of the following stages: stage 23 (in the middle of gonadal differentiation), stage 25 (at the end of gonadal differentiation), and hatching (approximately 2 weeks after gonadal differentiation).



Table 4-1. Design for the experimental dosing of alligator eggs. Eggs were assigned to an experimental group, exposed to the treatment, and sampled at either developmental stage 23, developmental stage 25, or hatching. Five animals were sampled for each treatment group at each developmental stage.
Stage 23 Stage 25 Hatching

Control Male 5 5 5

Control Female 5 5 5

p,p'-DDD (5 ppm) 5 5 5

p,p '-DDE (5 ppm) 5 5 5

Estradiol-170 (0.1 ppm) 5 5 5



Tissue Culture

A lethal injection of sodium pentobarbital (0.4 mg/g) was administered into the dorsal postcranial sinus of hatchlings and 9-month olds and into the vitellein vein of the





69


embryos. The gonad/adrenal/mesonephric complex (hereafter referred to as gonad) was removed and weighed. Tissue culturing procedures were modified from McMaster et al. (1995a).

For the descriptive study, the left gonad (termed "stimulated gonad") was placed in 700 pl media (Medium 199 with Hanks salts, L-glut, and 25 mM Hepes; Gibco BRI, Gaithersburg, MD) supplemented with 5 pM forskolin as a cAMP stimulator, 0. 1 mM 3isobutyl 1-methylxanthine (IBMX) as a cAMP protectant, and androstenedione as a steroid precursor. The right gonad (termed "unstimulated gonad") was placed in 700 p1 media supplemented with 0.1 mM IBMX. The gonads were incubated at 32*C for 5 hrs; the media was removed and flash frozen in liquid nitrogen, and fresh media (either supplemented media for the left gonad or non-supplemented media for the right gonad) was added. After 5 more hours, the media was again collected and changed. This last aliquot of media was collected after 15 hrs of incubation. After flash freezing, the media was stored at -720C prior to analysis for steroid hormones.

For the experimental study, the right gonad was removed for histology and the left gonad was placed in 500 pl media supplemented with forskolin, IBMX, and androstenedione as described above. After 5 hours of incubation, the media was collected, frozen in liquid nitrogen, and stored at -72*C. Histological Analysis

Gonads from the experimental animals were examined by histology to document which compounds induced the production of females at a male-producing temperature and to observe tissue-level changes induced by the compounds. The gonad was preserved in





70


Bouin's fixative, serial sectioned at 7 tm following paraffin embedding, and stained with a modification of Harris' trichrome staining procedure (Humason, 1972). Gonads were inspected and scored as testis or ovary by two independent researchers. Histological criteria originally reported by Forbes (Forbes, 1940) and recently reestablished by Guillette et al. (Guillette et al., 1994) were used to determine sex. In brief, criteria for testes included reduced cortex and medullary sex cord proliferation, whereas criteria for ovaries included hypertrophied cortex, medullary reduction, the presence of lacunae in the medulla, and germ cells in the cortex.

Radioimmunoassays

T concentrations were determined using a radioimmunoassay (RIA) previously validated for alligator plasma (Chapter 3(Guillette et al., 1997). Briefly, 70 pl of culture media was extracted 2x with 5 ml diethyl ether. The ether extract was air dried and the following were added to the dried assay tubes: Assay buffer (100 p.l; 0.05M borate buffer, pH=8.0), assay buffer with bovine serum albumen (100 pl; BSA at a final assay concentration of 0.15%), antibody (200 tl; final concentration of 1:25,000; Endocrine Sciences, Calabasas Hills, CA), and radiolabelled tracer (100 p1; 12,000 cpm per 100 p1; 1 mCi/ml; Amersham International, Bukinghamshire, England). Samples were compared to reference standards prepared at 0, 1.56, 3.13, 6.25, 12.5, 25, 50, 100, 200, 400, and 800 pg/tube. All samples and standards were prepared in duplicate. Tubes were incubated overnight at 4'C. Bound-free separation was accomplished by adding 500 pL 5% charcoal

- 0.5% dextran and immediately centrifuging at 1500 g at 4*C for 30 min. Supernatant (500 pl) then was added to 5 ml Scintiverse BD (Fisher Chemical Company) and the tubes





71


were counted on a Beckman scintillation counter. All samples were assayed in duplicate in a single assay, and the intraassay variability averaged 6.78%.

The T RIA was validated for the culture media by comparing media dilutions to

the T standard curve. A pool was made from an aliquot of each sample, and the following volumes of this pool were aliquanted: 0, 10, 25, 40, 55, and 70 pl. All tubes were brought to 70 pl with fresh, uncultured media. These samples were extracted and assayed as described above. Figure 1 presents the results of the RIA validation. Statistical analysis (test for homogeneity of slopes; SuperAnova; Abacus Concepts, Inc., Berkeley, CA) revealed that there was no significant difference (p=0.361) between the slopes of the culture media dilutions and the standard curve.

We attempted to measure estradiol- 173 (E2) in the culture media using similar

extraction and assay techniques to those in the T RIA. While this E2 RIA is successful in measuring plasma concentrations of E2 down to 12 pg/ml (Chapter 3) (Guillette et al., 1997), there was not sufficient E2 in the culture media to produce a validation for the RIA. Therefore, results for gonadal production of E2 are not presented. Statistical Analysis

Hormone concentrations were determined using a commercially available software package (Microplate manager III; Biorad, Hercules, CA). A comparison of T production between Apopka and Woodruff animals was conducted using a repeated measures ANOVA (SuperAnova; Abacus Concepts, Inc., Berkeley, CA). For the results of the Experimental Study, a 2-way ANOVA was used to evaluate the effects of treatment and developmental stage on T production. To obtain homogeneity of variances, T





72


concentrations were log transformed prior to this analysis. Fisher's Protected LSD was used as a post hoc test, and p=0.05 was the accepted level of significance.



100


80 - - Standard Curve

0 -u- Media Dilution



4m -25 1l
0
140 100
55 gI
20 -O0 1





Testosterone (pg)

Figure 4-1. Radioimmunoassay validation for the measurement of testosterone in the culture media. The statistical test for homogeneity of slopes revealed that the media dilution curve was not significantly different from the standard curve (p=O.36 1).


Results


Descriptive Study

Figure 4-2 presents the results of the descriptive study. Exposure to

androstenedione, forskolin, and IBMX ("stimulated gonads") stimulated significantly greater T secretion from ovaries (p<0.0001) and testes (p<0.0001) compared to unstimulated gonads. The interaction of lake and time of incubation was not significant for either testes or ovaries. Time of incubation had a significant influence on ovarian T





73


production in unstimulated (p=0.003) and stimulated (p=0.002) ovaries. However, time of incubation had no effect on testicular T production in either unstimulated (p=0.209) or stimulated (p=0.672). There was no difference in T production between Apopka and Woodruff animals for either stimulated ovaries (p=0.733), unstimulated ovaries (p=0.936), stimulated testes (p=0.395), or unstimulated testes (p=0.799). Experimental Study

We were unable to differentiate between testes and ovaries in stage 23 and stage 25 embryonic alligators. However, clear differences were seen in the hatchling alligators. Histological examination of the hatchling gonads revealed that female alligators were produced at a male-producing temperature in the estradiol- 170 (5 of 5) and p,p '-DDD (2 of 5) treatment groups (Figure 4-3). The remaining treatment groups produced the sex defined by the temperature regime.

Gonads from the p,p '-DDD-treated male hatchlings appear different from control males (Figure 4-4). Testes from animals exposed to 5 ppm p,p '-DDD appear developmentally accelerated, as there is a clearly defined lumen in the seminiferous tubules. Additionally, these testes have poorly organized seminiferous tubules when compared to testes from control males. Exposure to 5 ppm p,p '-DDE produced no observable abnormalities in the hatchling testes.

Figure 4-5 presents mean T concentrations of the alligators treated in ovo. The interaction of treatment group and embryonic stage was not significant (p=0.067), indicating that there is no apparent stage-specific treatment effect. However, both developmental stage (p=0.001) and treatment group (p=0.050) did influence T concentrations. Among stages, stage 23 animals produced more T than stage 25 animals





74


(p=0.017) and hatchlings (p=0.002). Among treatment groups, p,p '-DDD treated animals produced significantly more T than control females (p=0.004) and E2-treated animals (p=0.038).

Discussion


This study detected no difference in testosterone (T) production between juvenile alligators from Lakes Apopka and Woodruff. However, Lake Woodruff alligators treated in ovo with p,p'-DDD (one of the compounds found in Lake Apopka eggs) had significantly higher in vitro gonadal T production compared to control female alligators and E2-treated alligators. Therefore, it appears that exposure to p,p'-DDD is able to alter gonadal steroid production, but this alteration does not explain the hormonal or morphological alterations noted in Lake Apopka alligators.

Several studies have documented reduced circulating T concentrations in juvenile male alligators from Lake Apopka (Chapter 3(Guillette et al., 1997). This reduction appears to be defined in ovo, as 6-month-old Apopka males reared in the lab show reduced plasma T concentrations compared to lab-reared Lake Woodruff males (Guillette et al., 1994). The mechanisms of T reduction in the Apopka males are unknown, but a previous study found no difference in the production of T in ovaries or testes from 6month-old Apopka and Woodruff alligators (Guillette et al., 1995b). These results are consistent with those of this more detailed study that found no difference in gonadal T production between 9-month-old Apopka and Woodruff alligators. If the Apopka animals produced decreased T as a result of direct gonadal suppression of steroidogenesis, then the cAMP-stimulated gonads from Apopka animals would show reduced T synthesis









4
3.5
3
2.5
2
1.5 1
0.5


0


5 10 15
Time (hours)


20 25


Unstimulated Testes

A _. Apopka

-s- Woodruff







0 5 10 15 20 25

Time (hours)


4
3.5 3
2.5
2
1.5 1
0.5
0 1


0 5 10 15 20 25
Time (hours)


C
0

0
U) GD


1.8 1.6
1.4 1.2
1
0.8 0.6
0.4 0.2 0 1


0 5 10 15 20 25
Time (hours)


Figure 4-2. Testosterone production ( 1 SE) of testes (AB) and ovaries (C,D) from 9-month-old alligators from Lake Apopka and Lake Woodruff. A and C represent T production in unstimulated gonads, whereas B and D represent T production in gonads stimulated with a cAMP inducer (forskolin) and a T precursor (androstenedione).


Stimulated Testes
.B


-e
01


0
Ga I.


1.8 1.6
1.4 1.2 1
0.8 0.6
0.4 0.2-


01 01
S
C
0


V1 01

0


Unstimulated Ovaries


Stimulated ovaries


I





























Figure 4-3. Photomicrographs of hatchling alligator ovaries from a control female (A, incubated at 30'C) and a alligator that was incubated at a male-producing temperature (33C) and treated with p,p'-DDD (B). Numerous germ cells are present in the cortex
(C) of both animals, as well as a clearly defined medullary region (M). Both gonads appear to be normal ovaries, even though the p,p'-DDD-treated animal was incubated at a temperature that produces testes. Magnification 25x.

































ILI
4q f'-


-~ ( A-

































Figure 4-4. Photomicrographs of hatchling alligator testes from a control male (A) and a p,p'-DDD treated male (B). Tissue from the control males had well-organized seminiferous tubules with little or no lumen. In contrast, testes from hatchlings treated in ovo with p,p'-DDD have poorly organized testes and pronounced lumen. Magnification I0Ox.





79


mow A
14









TN



Af


ALA





80


Figure 4-5. Testosterone production ( 1 SE) of gonads from embryonic and hatchling alligators in the experimental treatment groups. Control females were incubated at 30'C (female-producing temperature), whereas all other treatment groups were incubated at 33'C (male-producing temperature). Gonads from alligators treated with p,p'-DDD produced significantly more T compared to control females and E2-treated animals. Although p,p'-DDD induced ovarian development in 2 of the 5 treated alligators, analysis revealed that the elevated T concentrations in the p,p'-DDD treatment group can be attributed solely to T production from the testes (p testicular T production = 77.3 ng/g/hr; p ovarian T production = 6.59 ng/g/hr).


30


' 25I-

0) 20a) C 15
0
a)
M 10
0 U)



0-


h A, o


M Stage 23 LI Stage 25 M Hatching


p,p'-DDD p,p'-DDE


Control Female


330C


.


I


Control Male


Estradiol


300C





81


compared to cAMP-stimulated Woodruff gonads. This is not the case and, therefore, the reduced circulating T concentrations previously reported for juvenile Apopka alligators can not be attributed to a direct reduction in gonadal T synthesis. Alternative sites for the androgen reductions in the Apopka males include (1) increased conversion of T to estradiol-17p by the enzyme aromatase, (2) increased excretion of T, (3) decreased availability of free T as a result of increased sex-steroid binding protein production, (4) decreased gonadotropin synthesis from the anterior pituitary. Future studies should explore the effects of endocrine altering contaminants on these potential sites of disruption.

The experimental egg dosing study found that both developmental stage and

treatment group had an influence on gonadal T synthesis. T synthesis is greatest during the period of gonadal differentiation (Stage 23 embryos). This period coincides with the window in development when temperature influences the determination of sex in alligators (Lang and Andrews, 1994). We were unable to detect significant concentrations of estradiol-170 (E2) in the culture media with our E2 radioimmunoassay, but a previous study found that female alligators progressively produce more E2 from stage 23 to hatching (Smith et al., 1995). Conversely, we found that both male and female alligators produce progressively less T from stage 23 to hatching. Therefore, it appears that T is produced in both males and female embryos at the onset of gonadal differentiation (Stage 23) and, in females, aromatase enzyme activity is elevated as development proceeds from Stage 23 to hatching. This causes increased E2 in females but not males (Smith et al., 1995).





82


Hatchling alligators that were exposed in ovo to 5 ppm p,p'-DDD exhibited

increased T production compared to control females and E2-treated hatchlings. In the hatchlings that were exposed in ovo to p,p'-DDD, consideration of both the histology and the T concentrations revealed that the elevated T production can be attributed to only those hatchlings that were males. Further, the testes of these p,p'-DDD treated males appear to be developmentally advanced compared to control males, as the sex cords have clearly defined lumen. Therefore, p,p'-DDD can override the effects of temperature by producing female alligators at a male-producing temperature, but also can induce testicular changes in alligators that are not "sex reversed." These effects are likely mediated through the estrogen receptor, as p,p'-DDD binds to the alligator estrogen receptor (Vonier et al., 1996). Future studies should determine if these testicular abnormalities are prevalent and persistent among males that are exposed to p,p'-DDD as embryos.

Unlike those exposed in ovo to p,p'-DDD, hatchlings that were exposed in ovo to 5 ppm pp '-DDE exhibited no structural or functional gonadal alterations. All p,p'-DDE exposed animals had apparently normal testes and produced normal concentrations of testosterone. This lack of effect may be due to the dynamics ofp,p'-DDE with steroid receptors. Unlike p,p '-DDD, p,p '-DDE exhibits weak affinity to the alligator estrogen receptor (Vonier et al., 1996). In rodents, p'p'-DDE is an antagonist of the androgen receptor both in vitro (Kelce et al., 1995) and in vivo (Kelce et al., 1997), although p,p'DDE may interact with the androgen receptor at a dose approximately 106-fold higher than endogenous testosterone (Gaido et al., 1997). Therefore, it is possible that higher dosages ofp,p'-DDE could cause testicular abnormalities through anti-androgenic mechanisms.





83


Although the present study found no difference in testicular T production between Lake Apopka and Lake Woodruff alligators, it is possible that such a difference would occur or become evident in older alligators. Environmental contaminants can differentially affect plasma androgen concentrations depending on the animal's developmental stage, sex, and reproductive status. The effects of bleached kraft mill effluent (BKME) on androgen concentrations in white sucker fish (Catostomus commersoni) best illustrate this. Gagnon et al. (1 994a) found that BKME exposure causes no change in T concentrations in males, but induces increased T concentrations in females. However, a more detailed study of male white suckers found that while suckers sampled in August were not influenced BKME exposure, BKME caused increased T concentrations in males sampled in September (Munkittrick et al., 1992b). Conversely, BKME-exposed male fish had reduced 1 1-ketotestosterone in August but not in September (Munkittrick et al., 1992b). Although the mechanisms for increased circulating androgens remain unknown, it has been suggested that the decreased androgen concentrations are due to BKME-induced reduction in steroid biosynthesis (McMaster et al., 1991; McMaster et al., 1995b).

This study has shown that in ovo exposure to p,p'-DDD can cause structural and functional changes in the gonads of hatchling alligators. In some individuals, exposure to

5 ppmp,p '-DDD caused the production of females in prospective males. In other individuals exposed to p,p'-DDD, sex reversal was not induced, but testes appeared to be developmentally accelerated and poorly organized. These testes also produced significantly more T compared to ovaries, whereas testes from control animals did not. Although alligators from Lake Apopka are exposed to similar concentrations ofp,p '-DDD (up to 1.8 ppm (Heinz et al., 1991)) during embryonic development, nine-month-old





84


alligators from Lake Apopka do not show such abnormalities. No difference was seen in T synthesis between alligators from Lake Apopka and Lake Woodruff. Future studies should more fully characterize both the endocrine-altering effects ofpp'-DDD and the mechanisms of endocrine disruption in Lake Apopka alligators.














CHAPTER 5
ALTERATIONS IN STEROIDOGENESIS IN ALLIGATORS (ALLIGATOR
MISSISSIPPIENSIS) EXPOSED NATURALLY AND EXPERIMENTALLY TO ENVIRONMENTAL CONTAMINANTS

Introduction


Environmental contaminants alter the reproduction of a number of wildlife species by changing the normal endocrine environment that mediates sexual differentiation and function (Chapter 2). Many of these endocrine alterations are thought to occur by direct interactions between the contaminants and hormone receptors (McLachlan, 1993), but the specific mechanisms by which most environmental contaminants cause endocrine disruption are unknown. Although all vertebrates are potentially susceptible to reproductive disruption by endocrine-disrupting contaminants (EDCs), many ectothermic vertebrates are particularly sensitive due to the processes mediating the organization of the reproductive system (Guillette et al., 1995a). Unlike birds and mammals, many fish, amphibians, and reptiles exhibit environmental sex determination, by which the gender of the undifferentiated embryo is determined by an environmental variable. In many reptiles, the temperature of egg incubation determines the sex of the offspring (Bull, 1980). Exposure of developing reptile embryos to exogenous chemicals can mimic the effects of temperature on sex determination. For example, when red-eared turtle (Trachemys scripta) embryos are incubated at a male-producing temperature and exposed to estradiol170 during the window of developmental sex determination, phenotypically female turtles Note: This chapter is published in Environmental Health Perspectives (Crain et al., 1997b)
85





86


are produced (Wibbels et al., 1991; Wibbels et al., 1993). This estrogen-induced sex reversal appears to be dose dependent (Crews et al., 1991), and suggests that other steroidal agonists, steroidal antagonists, and steroidogenic disrupters could alter normal sexual differentiation. Indeed, Wibbels and Crews (1992) found that steroid hormones are not exclusive in their ability to alter normal sex determination, as many estrogen agonists and steroidogenic modifiers mimic and/or reverse the effects of temperature on the differentiation of primary sex organs in red-eared turtles.

The specific mechanisms by which temperature determines gender are unknown, but it is hypothesized that temperature stimulates or suppresses pivotal steroidogenic enzymes (Wibbels et al., 1994). These enzymes then propagate a cascade of events leading to the organization of a testis or ovary. This hypothesis is supported by work conducted on the steroidogenic enzyme aromatase. Aromatase converts androgens to estrogens by binding the C19 androgen substrate and catalyzing several reactions leading to a phenolic ring characteristic of estrogens (Simpson et al., 1994). Several lines of evidence support the pivotal role of aromatase in temperature-dependent sex determination. First, several studies indicate that aromatase activity is increased in prospective females during periods coinciding with thermosensitivity (Desvages and Pieau, 1992; Chardard et al., 1995; Smith et al., 1995). Second, high doses (50-100 pg per egg) of testosterone cause feminization of T. scripta at a male-producing temperature (Wibbels and Crews, 1992; Crews et al., 1995b). Because testosterone is the precursor to estradiol- 170, this phenomenon is thought to be mediated by the enzyme aromatase. Third, administration of an aromatase inhibitor induces male sex determination in both a female unisexual (parthenogenetic) lizard and a turtle with temperature-dependent sex





87


determination (Wibbels and Crews, 1994). Collectively, these studies suggest that aromatase is an enzyme critical to thermosensitive sex determination and is capable of being modified by extrinsic factors.

In consideration of these studies, I propose that the endocrine-altering effects of some environmental contaminants may be mediated via changes in the expression or activity of the aromatase enzyme. Two studies, one descriptive and one experimental, were conducted to test this hypothesis. First, juvenile alligators from a control lake and a lake historically contaminated with a number of persistent organochlorines were analyzed for plasma steroid hormones and in vitro gonadal-adrenal aromatase activity. Second, embryos from a control lake were exposed to several known hormonal modifiers and two common herbicides, and hatchlings were analyzed for (a) egg chorioallantoic fluid hormones, (b) plasma steroid hormones, and (c) in vitro aromatase activity. Using these studies, I sought to determine whether aromatase function could explain endocrine alterations in alligators exposed to EDCs.

Materials and Methods


Animals and Treatments

For the descriptive study, eggs were collected from 6 nests on Lake Apopka, Florida (contaminated lake) and 6 nests on Lake Woodruff National Wildlife Refuge, Florida (control lake) during the first week of July 1995. Lake Apopka is designated as one of Florida's most polluted lakes (Schelske and Brezonik, 1992) due to extensive agricultural activities around the lake, a sewage treatment facility associated with the city of Winter Garden, FL, and a major pesticide spill from the Tower Chemical Company.


F_





88


The pesticide spill, which occurred in 1980, consisted primarily of dicofol but had significant amounts of DDT, DDE, and DDD in the mixture (U.S. EPA, 1994). Analysis of alligator eggs taken from Lake Apopka in 1984 and 1985 revealed significant residues of toxaphene, dieldrin, p,p '-DDE, pp '-DDD, trans-nonachlor, and PCBs (Heinz et al., 1991). A previous study found evidence of "estrogenic contamination" among the female alligators of Lake Apopka (Guillette et al., 1994) and, as I wanted to minimize the number of eggs taken from Lake Apopka, eggs were incubated only at a female-producing temperature (30'C) (Lang and Andrews, 1994). After hatching, alligators were housed at Sante Fe Teaching Zoo (Santa Fe Community College, Gainesville, FL) in outdoor, semiaquatic enclosures. At nine months of age, the female alligators were transported to the laboratory for collection of tissues.

For the experimental study, eggs were collected from 7 nests on Lake Woodruff, Florida during the first week of July 1995. Eggs were transported to the lab, placed in an incubator at 300C, and one egg from each clutch was opened to stage the embryos. Staging was based on criteria defined by Ferguson (1985). Five days after collection (and prior to the temperature sensitive period when sex determination occurs), eggs were separated into two groups such that half of the eggs from one clutch were incubated at 300C (female producing) and half at 330C (male producing). Eggs were maintained at approximately 90% humidity using sphagnum moss as incubation material. Within each incubation group, eggs from each clutch were distributed among 6 treatment groups of varying dosages (see Table 5-1). One treatment group served as a control and three groups served as endocrine-disrupting standards: estradiol- 170, tamoxifen which acts as an estrogen in embryonic alligators but as an antiestrogen in hatchlings (Lance and Bogart,





89


1991), and vinclozolin which is a potent antiandrogen in rodents (Gray et al., 1993b). The two remaining treatment groups were the modern-use herbicides atrazine and 2,4dichlorphenoxyacetic acid (2,4 D). Treatments were applied topically to the eggshell in 50 pl of 95% ethanol, a technique frequently used to transport compounds inside reptilian eggshells (Crews et al., 1991; Wibbels and Crews, 1992). Using this method, Crews et al. (1991) found that greater than 89% of the applied compound is incorporated into the embryo. The treatments were applied at stage 21 of embryonic development, the beginning of the critical period of gonadal differentiation (Lang and Andrews, 1994).


Table 5-1. Experimental treatments and dosages (in parts-per-million - ppm) that were applied to different groups of eggs. A sample size of 5 eggs was included in each dosetreatment group.
Treatments Effect Doses

Control' None Nothing; Diluent only
Estradiol Natural Estrogen 0.014, 0.14, 1.4, 14 ppm
Vinclozolin Androgen antagonist in rodents 0.14, 1.4, 14 ppm
Tamoxifen Estrogen agonist/antagonist 0.14, 1.4, 14 ppm
2,4-D ?, Common herbicide 0.14, 1.4, 14 ppm

Atrazine ?, Common herbicide 0.14, 1.4, 14 ppm


a Each chemical was solubilized in 95% ethanol prior to topical application on the egg and, thus, two control doses were used-one with 95% ethanol and one without. There was no difference between these controls for any of the variables measured.



Upon pipping, chorioallantoic fluid was collected and frozen at -720C until steroid hormone analysis. Total protein content in the CAF samples was determined using a commercially available Bradford assay kit (Biorad, Hercules, CA), and CAF steroid





90


hormone concentrations are presented as per ptg protein. This was necessary due to differential hydration states of the CAF samples. Aromatase Assay

Hatchlings were individually housed for 10 days prior to tissue collection.

Following the collection of blood from the dorsal post-cranial sinus, a lethal injection of sodium pentobarbital (0.4 mg/g) was administered in the sinus. Animals are anesthetized within thirty seconds using this method. The right gonadal-adrenal-mesonephros (GAM) complex was immediately removed for the aromatase bioassay. Aromatase activity was measured indirectly based on the release of tritium from 1 -3H-androstenedione during aromatization of the substrate into estrogen (Smith and Joss, 1994). Briefly, the tissue was placed in 400 pl culture media (RPMI-1640; Sigma Chemical Co., St. Louis, MO) supplemented with 0.8 mM tritiated androstenedione (DuPont NEN Research Products; # NET-926). After a 6 hr incubation at 320C, 300 pl of the media was transferred to a new tube. Chloroform (1.5 ml) was added, the tube was vortexed, and centrifuged for 15 min at 2000 g. A 200 pl aliquant of the aqueous phase was added to a new tube. 5% charcoal

- 0.5% dextran (200 tl) was added, the tube was vortexed and then immediately centrifuged for 15 min at 2000 g. Scintiverse BD (5 ml) was added to 300 pl supernatant and the tube was counted on a Beckman Scintillation counter. Aromatase activity is proportional to the amount of tritium in the scintillation vial and is calculated as a percentage of the total substrate added. After subtracting the nonspecific tritium release, the DPM of the sample tubes are converted to a percentage of the total DPM added. This percentage is multiplied by the mass of the substrate added. After adjusting for extraction





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loss, the value obtained represents the amount of substrate converted to tritiated water, which is proportional to aromatase activity. Assay sensitivity was defined as twice the mean cpm of blank tubes, which corresponds to 0.15 pmol/g/hr for the average weight GAM (0.032g).

GAMs from three additional control female alligators were used to determine the specificity of the aromatase assay. The left GAM was incubated as above, while the right GAM was exposed to media supplemented with the aromatase inhibitor 4-hydroxy androstenedione (100 pM). Alligators exposed to the aromatase inhibitor had significantly lower GAM aromatase activity (p=0.45 pmol/g/hr) compared to the individuals incubated normally (pt=3.15 pmol/g/hr).

Histology

Histology was conducted to determine histological sex in order to document which compounds induced sex reversal. A complete histopathological examination of the GAMs was beyond the scope of this study. The left GAM was preserved in Bouin's fixative, serial sectioned at 7 pm following paraffin embedding, and stained with a modification of Harris' trichrome staining procedure (Humason, 1972). Gonads were inspected and scored as testis or ovary by two independent researchers. Histological criteria originally reported by Forbes (1940) and recently reestablished by Guillette et al. (1994) were used to determine sex. In brief, criteria for testes included reduced cortex and medullary sex cord proliferation, whereas criteria for ovaries included hypertrophied cortex, medullary reduction, the presence of lacunae in the medulla, and germ cells in the cortex.





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Radioimmunoassays

Estradiol- 171 (E2) and testosterone (T) concentrations were measured in plasma of the 9-month-old descriptive study animals and in plasma and chorioallantoic fluids (CAF) of all hatchlings that provided ample fluids. Radioimmunoassays for E2 and T were performed as previously described (Folmar et al., 1996) with the following modifications in sample extraction. CAF (750 pl) was mixed overnight (15 hrs) with 2 ml of 95% ethanol. The suspension was centrifuged at 1200 g for 20 min. Supernatant (500 PIl) was pipetted in duplicate for each sample and dried under constant air stream. Extraction efficiency averaged 92% for T and 94% for E2 with this method. For plasma extraction, plasma (125 pl) was mixed with 4 ml ethyl ether for 1 min. The aqueous layer was frozen in a dry ice-methanol bath (-25*C), and the ether phase decanted into an assay tube. The aqueous pellet was reextracted with ether, and the ether added to the assay tube. The ether was dried with constant air stream. Extraction efficiency was consistent and averaged 95% for T and 94% for E2. Crossreactivities of the T antisera (T3-125, Endocrine Sciences, Calabasas Hills, CA) to other ligands are as follows: dihydrotestosterone, 44%; A-1-testosterone, 41%; A-1-dihydrotestosterone, 18%; 5 oxandrostan-30,170 diol, 3%; 4-androsten-3p,170-diol, 2.5%; A-4-androstenedione, 2%; 50-androstan-30,170-diol, 1.5%; estradiol, 0.5%; all other ligands <0.2%. Crossreactivities of the E2 antisera (E26-47, Endocrine Sciences, Calabasas Hills, CA) to other ligands are as follows: estrone, 1.3%; estriol, 0.6%; 16-keto-estriol, 0.2%; all other ligands <0.2%. For the plasma RIAs, inter-assay variance was 15.0% for T and 12.6% for E2, and intra-assay variance was 3.6% for T and 3.7% for E2. For the CAF RIAs inter-





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assay variance was 11.8% for T and 16.1% for E2, and intra-assay variance was 4.68% for T and 3.5% for E2.

Statistics

Hormone concentrations were estimated from raw data with the commercially

available Beckman EIA/RIA ImmunoFittm software program (Fullerton, CA). Statistics were performed with the software packages StatView (Abacus Concepts, Inc., Berkeley, CA, 1992) and SuperAnova (Abacus Concepts, Inc., Berkeley, CA, 1989). For the descriptive study, a t-test was used for between-lake comparisons. In the experimental study, a two-factor ANOVA was used to test the effects of treatment and dose on aromatase activities. Where the interaction of treatment and dose was not significant, the factor of dose was removed from the analysis and a one-factor ANOVA was used to test the effects of treatment on hormone concentrations and aromatase activity. Fisher's protected LSD was used as a post-hoc test to discriminate which groups differed significantly.

Results


Descriptive Study - Female Juvenile Alligators

Results of the aromatase enzyme assay are expressed both as fmol/hr and

pmol/g/hr. The former is used in other studies of alligator gonadal aromatase activity (Smith and Joss, 1994; Smith et al., 1995) and is presented here for comparative purposes only. A comparison of the female juvenile alligators found that gonadal-adrenalmesonephros (GAM) aromatase activity was significantly elevated in Lake Woodruff alligators compared to Lake Apopka alligators (see Table 5-2 and Figure 5-1). Mean




Full Text

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EFFECTS OF ENDOCRINE-DISRUPTING CONTAMINANTS ON REPRODUCTION IN THE AMERICAN ALLIGATOR, ALLIGATOR MISSISSIPPIENSIS By DAVID ANDREW CRAIN A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 1997

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This is dedicated to my wife Holly, Who gives me endless love, peace, and And to my son Jared, Who gives me hope for the future.

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ACKNOWLEDGMENTS There are numerous people to whom I owe thanks for assisting me in my research. First and foremost, I want to thank my friend and mentor Dr. Louis Guillette. Without his constant encouragement and assistance, none of this work would have been possible. I also want to thank my other committee members for their input. Dr. Karen Bjorndal, Dr. David Evans, Dr. Larry McEdward, and Dr. Dan Sharp helped mold my ideas into sound studies. Cathy Cox, Ed Orlando, Dan Pickford, and Andy Rooney were fellow graduate students who both assisted with various laboratory and field projects and helped develop my research ideas. I was fortunate to be able to work with the following undergraduates who helped with my laboratory work: Bart Edmiston, Amy Pickle, Megan Pew, Michelle Scargle, Daniel Spiteri, Shadi Tolymat, and Christi Waldi. All of these undergraduates are of the highest caliber, and I am indebted to them for their hard work. Collection of wild alligators and eggs was made possible through the collaborative assistance of Allan Woodward of the Florida Game and Freshwater Fish Commission and Franklin Percival of the Florida Cooperative Fish and Wildlife Research Unit. I am also indebted to the many others who assisted in the field studies. Funding for my research was kindly provided through a graduate student fellowship from the Environmental Protection Agency, a research grant from the University of Florida Division of Sponsored Research, and Greg Masson of the U.S. Fish and Wildlife Service. Also, the University of Florida Department of Zoology provided support that made this work possible. in

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TABLE OF CONTENTS page ACKNOWLEDGMENTS ijj ABSTRACT vi CHAPTERS 1 INTRODUCTION i The Problem I The Purpose 2 The Rationale 3 2 ENDOCRINE-DISRUPTING CONTAMINANTS AND REPRODUCTION IN VERTEBRATE WILDLIFE 5 Introduction 5 An Evolutionary Perspective 5 Theory of Disruption: Organization vs. Activation 10 Alteration of Reproductive Tissues 14 Conclusions 33 3 SEX-STEROID AND THYROID HORMONE CONCENTRATIONS IN JUVENILE ALLIGATORS (ALLIGATOR MISSISSIPPIENSIS) FROM CONTAMINATED AND REFERENCE LAKES IN FLORIDA 36 Introduction 36 Materials and Methods 39 Results 46 Discussion 5g iv

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4 TESTOSTERONE SYNTHESIS IN EMBRYONIC AND JUVENILE ALLIGATORS EXPOSED TO ENDOCRINE-ALTERING ENVIRONMENTAL CONTAMINANTS 65 Introduction 65 Materials and Methods 67 Results 72 Discussion 74 5 ALTERATIONS IN STEROIDOGENESIS IN ALLIGATORS (ALLIGA TOR MISSISSIPPIENSIS) EXPOSED NATURALLY AND EXPERIMENTALLY TO ENVIRONMENTAL CONTAMINANTS 85 Introduction 85 Materials and Methods 87 Results 93 Discussion 99 6 CELLULAR BIOAVAILABILITY OF NATURAL HORMONES AND ENVIRONMENTAL CONTAMINANTS AS A FUNCTION OF SERUM AND CYTOSOLIC BINDING FACTORS 105 Introduction 105 Materials and Methods 108 Results 113 Discussion 119 7 SUMMARY AND CONCLUSIONS 126 The Rationale Revisited 126 The Purpose Revisited 127 The Problem Revisited .128 LIST OF REFERENCES 132 BIOGRAPHICAL SKETCH 153 v

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Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy EFFECTS OF ENDOCRINE-DISRUPTING CONTAMINANTS ON REPRODUCTION IN THE AMERICAN ALLIGATOR, ALLIGATOR MISSISSIPPIENSIS By David Andrew Crain August, 1997 Chairman: Louis J. Guillette, Jr. Major Department: Zoology Wildlife and humans are exposed to numerous chemicals in the environment that can alter the function of the endocrine system. These endocrine-disrupting contaminants (EDCs) change the normal functioning of reproduction in many wildlife species. This dissertation contributes to the current knowledge of contaminant-induced endocrine disruption by reviewing the phenomenon in wildlife species, describing endocrine disruption in several wild populations of alligators, and examining the mechanisms of this endocrine disruption. Plasma concentrations of estradiol170 (E 2 ), testosterone (T), triiodothryonine (T 3 ), and thyroxine (T 4 ) in juvenile alligators from two contaminated lakes and one reference lake in Florida are presented. Males from the two contaminated lakes (Lakes Apopka and Okeechobee) showed no correlation between T 4 and total length, whereas males from the reference lake (Lake Woodruff) showed a strong correlation. Male vi

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alligators from the contaminated lakes had significantly lower T concentrations compared to males from the reference lake, and there was a poor relationship between T and total length in Apopka males (r 2 =0.007, p=0.75). These results are consistent with other studies indicating that alligators living in Lakes Apopka and Okeechobee experience endocrine disruption. Mechanisms underlying the alterations in steroid hormones were studied in a series of experiments. First, gonadal T synthesis was examined in wild alligators from a contaminated lake (Lake Apopka) and a reference lake (Lake Woodruff). Whereas there was no difference in T production in animals of the two lakes, male alligators from Lake Woodruff that were exposed in ovo to p,p -DDD had elevated T synthesis. Testes from these exposed animals appeared developmentally accelerated. Second, steroidogenesis was studied by examining aromatase activity in gonads from hatchlings that were exposed in ovo to two modern-use herbicides. Alligators that were exposed to atrazine had higher aromatase activity compared to controls. This study also showed that gonadal aromatase activity is higher in animals from Lake Woodruff than those from Lake Apopka. Third, evidence is presented that few EDCs interact with steroid binding proteins, thus increasing the cellular exposure to EDCs relative to endogenous hormones. The dissertation concludes by presenting a model summarizing the effects of EDCs on steroid hormone dynamics and emphasizing future research needs. vii

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CHAPTER 1 INTRODUCTION The Problem The study of adverse effects of xenobiotics-defined as toxicology-dates back to the earliest humans, who used animal venoms and plant extracts for hunting. Scholars of ancient Greece, such as Socrates (470-399 BC) and Theopharstus (370-286 BC), studied lethal dosages and began classifying poisons. But it was not until the Age of Enlightenment that toxicology was viewed as a science and a discipline. Philippus Aureolus Theophrastus Bombastus von Hohenheim-Paracelsus (1493-1541) is viewed as the father of modern toxicology due to his premise that (1) experimentation is essential to determine the response of a chemical, (2) a distinction should be made between the therapeutic and toxic properties of a chemical, and (3) a chemical's properties are sometimes, but not always, distinguishable only by dose. Paracelsus' wisdom is evident in his most famous quote: All substances are poisons; there is none which is not a poison. The right dose differentiates a poison from a remedy. Circa 1533 Modern toxicology primarily has focused on lethal endpoints as a result of animal and human exposure to large quantities of numerous toxic compounds. Two of the most common tests used to determine the toxicity of a chemical are the LD 50 (lethal dose required to kill 50% of the experimental units) and the Ames test (used to determine the 1

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2 dose that stimulates mutagenesis, and thus cancer). While knowledge pertaining to these tests is critical, other sublethal endpoints are necessary for evaluating the toxicity of a particular xenobiotic. Whereas toxicologists view the individual as their experimental unit, ecologists and conservationists view a population of animals as the experimental unit. With this "population perspective," reproduction of the individual is as important a variable as the survivorship of that individual. In essence, survival is not the seminal variable dictating survival of a population; reproduction is that variable. Therefore, successful reproduction, and not death or carcinogenicity exclusively, should be used as an endpoint in the assessment of the toxicity of a xenobiotic. Reproduction has long been a consideration in toxicity studies, but the last 5-10 years have seen an exponential increase in the number of studies examining the adverse effects of compounds on reproduction. During this time, it has been discovered that many xenobiotics adversely affect reproduction by altering the endocrine system (Colborn et al., 1993). This has led to a new subdiscipline that merges the fields of toxicology and endocrinology— the study of endocrine-disrupting contaminants (EDCs). As the endocrine system is involved in the regulation of virtually all biological phenomena, there is a high probability that EDCs also alter functions other than reproduction (such as growth and maintenance). The Purpose The purpose of this dissertation is to better understand the scope of endocrine disruption in vertebrate wildlife and the ways that environmental contaminants can disrupt

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3 the endocrine system. Specifically, the research described herein examines endocrine disruption in the American alligator (Alligator mississippiensis). Several recent studies have noted structural and functional abnormalities of the reproductive system in alligators exposed to EDCs (Guillette et al., 1994; Guillette et al., 1995b; Guillette et al, 1996b), but the mechanisms of endocrine disruption in these animals remain poorly understood. It is hoped that this dissertation will help elucidate the phenomena of endocrine disruption. The Rationale In order to ascertain the magnitude and scope of contaminant-induced endocrine disruption in vertebrates, a review will be provided in Chapter 2 that examines endocrine disruption in vertebrate wildlife. Also in this chapter, concepts that are necessary for understanding endocrine disruption will be presented. In Chapter 3, data will be presented from wild alligator populations, two of which are exposed to a variety of environmental contaminants. In Chapters 4, 5, and 6, the specific mechanisms through which environmental contaminants can cause endocrine disruption in the American alligator will be examined. The theoretical framework for testing these potential mechanisms is presented in Fig 1-1. A contaminant potentially can influence the endocrine system by causing an alteration at any point in the cycle of steroid hormone dynamics. The potential for contaminants to alter steroid production will be explored in Chapters 4 and 5, and contaminant bioavailability will be considered in Chapter 6. Finally in Chapter 7, conclusions and future perspectives will be given.

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4 • i— t

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CHAPTER 2 ENDOCRINE-DISRUPTING CONTAMINANTS AND REPRODUCTION IN VERTEBRATE WILDLIFE Introduction The fields of toxicology, endocrinology, and reproductive physiology recently have combined resources to study the effects of endocrine-disrupting contaminants (EDCs) in wildlife populations. EDCs include a wide variety of chemicals that are only related by the ability to disrupt normal function of an animal's endocrine system (McLachlan, 1993). Although studies documenting endocrine disruption by contaminants have been conducted for many years, only recently have studies systematically explored the effects and mechanisms of EDCs. This recent synthesis has led to the hypothesis that anthropogenic EDCs are associated with an increase in abnormalities of the reproductive system and a decrease in reproductive success in vertebrates (Colborn and Clement, 1992). This brief review considers the phenomenon of contaminant disruption of wildlife reproduction at several levels: evolutionary, tissue, and mechanistic. Only through such an integrative perspective can an accurate representation be achieved and solutions gained. An Evolutionary Perspective Animals are constantly exposed to many foreign chemicals in their diet, and these compounds can decrease the survival and reproductive capacity of individuals. Given that evolution favors maximal reproductive success, the evolution of mechanisms to eliminate Note: This chapter is published in Reviews in Toxicology (Crain and Guillette, 1997a). 5

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6 the deleterious effects of xenobiotics is expected. One such mechanism that has evolved in animals is the phase I and phase II biotransformation pathways (Brouwer, 1991). Phase I and II processes can render xenobiotics more lipophilic through the addition of hydroxyl moieties (phase I) or conjugation to polar endogenous molecules (phase II). Most phase I reactions are carried out by a particular class of enzymes, the cytochrome P450 enzymes of the liver. These enzymes increase the hydrophilicity of the compound by adding a hydroxyl moiety to it. The induction of hepatic P450 enzymes after contaminant exposure is a very sensitive and phylogenetically conserved mechanism, and as such, P450 induction is commonly used to monitor contaminant exposure in wildlife species (Rattner et al., 1989). While many P450 enzymes are involved in decontamination, other P450 enzymes are involved in steroidogenesis — the production and conversion of steroid hormones. Thus, a family of enzymes that are induced by contaminant exposure to detoxify compounds could also cause an alteration in the hormonal environment of an animal. This leads to an evolutionary dilemma concerning the two fundamental constraints that dictate natural selection: the constraint of reproduction and the constraint of survival (Jacob, 1977). Survival requires that xenobiotics be detoxified and eliminated but, at the same time, reproduction can not be compromised. Therefore, to support survival and reproduction, P450 enzyme induction must be specific to detoxification and not alter steroidogenesis. The specificity of P450 induction is unknown, and this should be an area of future research. It is possible that many of the reproductive abnormalities seen in vertebrates during recent decades are due partially to the alteration of normal steroidogenesis as a result of increased exposure to a wider range of contaminants.

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7 Many plants contain endocrine-disrupting compounds such as antithyroidal goitrogens (Yamada et al., 1974), phytoestrogens (Hughes, 1988), and androgen disrupters (Gray et al., 1996) that can alter reproduction after ingestion. The innovation of such endocrine-altering compounds by plants is predicted by evolutionary theory that suggests that plants will respond to predation by herbivores (Ehrlich and Raven, 1964; Jansen, 1980). Reproduction of wildlife species is adversely affected by plants containing endocrine-disrupting compounds (Leopold et al., 1976; Berger et al., 1977; Howell and Denton, 1989), but such plant-animal interactions have evolved through time to yield an evolutionary stable strategy for the wildlife and plants. For instance, ruffed grouse utilize aspen buds as a major food source, but the grouse have had to adapt to the endocrinedisrupting effects of a component of the aspen buds, coniferyl benzoate. When ingested, coniferyl benzoate is metabolized into ferulic acid which causes anti-reproductive effects through altering estrogen and prolactin function (Jakubas et al., 1993). The grouse avoid such endocrine-altering effects by both selectively feeding on buds having low concentrations of coniferyl benzoate (Jakubas et al., 1989) and utilizing aspen buds less frequently when coniferyl benzoate levels are high (Jakubas and Gullion, 1991). In addition to such behavioral alterations, adaptation to plant endocrine disrupters can involve chronological or physiological adjustments. Wildlife can adapt to the compound by altering the timing of reproduction to avoid exposure during critical reproductive stages. Selection would favor such chronological avoidance, since animals that avoid the phytotoxicants would have increased reproductive success. Wildlife can also physiologically adapt to plant-derived endocrine disrupters. For instance, after ingestion the phytoestrogen genistein induces many of the same effects as 1 7p-estradiol (Levy et al .,

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8 1995), but genistein also stimulates production of sex-hormone binding globulin (SHBG) (Mousavi and Aldercreutz, 1993) and suppresses aromatase activity (Adlercreutz et al., 1993), both of which reduce the amount of bioavailable estrogen. The role of SHBG in controlling the bioavailability of genistein itself is unknown, but SHBG could be protective if it binds genistein as it does 170-estradiol. It is clear that coevolution has provided a means for animals to adapt to phytochemical mimicry of reproductive hormones, but such adaptations have not evolved for animals exposed to anthropogenic EDCs. From a classical gradualist perspective, adaptation to an environmental condition (here, EDC exposure) requires a predictable exposure over a relatively long period of time— generations. Whereas wildlife have coevolved with plants for hundreds of millions of years, exposure of wildlife to large numbers of anthropogenic EDCs is limited to the past two centuries— since the onset of the industrial revolution. Indeed, the largest exposure in number of compounds and concentrations of these compounds has occurred in the last 50 years. Relatively few generations have been produced in this short amount of time, precluding extensive adaptation for many long-lived vertebrates. Given generations, constant or predictable exposure of a population to a contaminant provides an opportunity for adaptation. By chance, some individuals will be better able to survive and reproduce during the exposure. These animals produce progeny with similar characteristics, and eventually the population will consist of animals able to survive and successfully reproduce during the contaminant exposure. Although exposure to some EDCs could be predictable (e.g., constant exposure of a fish population to sewage effluent), most wildlife populations are exposed to EDCs in an unpredictable manner, making chronological and physiological adaptation

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9 difficult. Moreover, the combination of chemicals that wildlife are exposed to can vary dramatically by location, season, and life stage. It has been suggested that exposure to dietary phytoestrogens is far greater than exposure to industrial estrogenic compounds and, thus, that the contribution of industrial estrogenic compounds to reproductive dysfunction is nominal (Safe, 1995). Although endocrine disruption is likely to occur in embryos exposed to high concentrations of maternally ingested phytoestrogens, laboratory data suggest that the potency of ingested phytoestrogens is nominal when compared with other modes of exposure (such as direct in utero or in ovo exposure) (Cain et a!., 1987; Lien and Cain, 1987). Guillette et al. (1996a) recently argued that the in vivo estrogenicity of an "ecoestrogen" (an environmental contaminant that acts as an estrogen agonist) is determined by many factors that differ between natural phytoestrogens and synthetic estrogens, namely (a) the binding affinity of the compound to the estrogen receptor, (b) the accumulation of the compound in the environment and the body, (c) the degradation or metabolism of the compound in the environment and the body, and (d) the availability of the ecoestrogen to the target cell. When these variables are considered, the potential "potency" of many EDCs may be substantial. Additionally, each of these variables can be contaminantand species-specific, and a significant degree of phylogenetic variation can be seen in the response to EDCs. For example, Kelce et al. (1995) found that p,p -DDE acted as an androgen antagonist in rodents whereas Soto et al. (1995) found it to be estrogenic in human MCF-7 cells.

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10 Theory of Disruption: Organization vs. Activation The effects of EDCs on reproduction can be classified as either organizational or activational. A contaminant that permanently modifies the morphology or function of a tissue as the result of exposure during a particularly sensitive period of development is said to have an organizational effect, whereas if an EDC temporarily alters the function of a normally organized tissue, it has induced an activational effect (Guillette et al., 1995a). Figure 2-1 illustrates the typical periods when organizational and activational effects are induced in a vertebrate. In general, alterations occurring from gamete production through juvenile development are permanent and organizational in nature, whereas insults during mature stages are transitory and activational. Although the dichotomy between organizational and activational effects is imperfect (Arnold and Breedlove, 1985), this concept provides an appropriate framework for the discussion of contaminant-induced endocrine effects. As an embryo develops, numerous inductive processes are initiated through extremely complex networks and cascades (Jacobson, 1966). The evolutionary result is that the overall developmental process will resist significant modification (Raff and Kaufman, 1991). However, exposure of embryos to compounds that mimic or block critical developmental signals can readily alter normal developmental processes and, thus, the organization of structures. Embryos are particularly sensitive to organizational disruption by EDCs due to (a) high rates of cellular division, differentiation, and metabolism, (b) critical windows of sensitivity, (c) the relative amount of contaminant

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1] available to cells, and (d) mobilization of maternal bioaccumulated contaminants (Guillette etal., 1995a). There is a great deal of variation in the organizational responses of various phyla to EDC exposure. For instance, consider the organizational effects of neonatal polychlorinated biphenyl (PCB) exposure in two animals that have different modes of sex determination. Sex in some species of turtles (and many other reptiles) is determined by the temperature of embryonic development (Pieau, 1996), but the development of the embryonic gonad can be altered by steroid exposure. Administration of exogenous estradiol can cause feminization of individuals incubated at male-producing temperatures (Crews et al., 1991) and, thus, estrogens and chemicals that act as estrogen agonists have the potential to alter sexual differentiation. Indeed, Bergeron et al. (1994) discovered that exposure of developing turtles (Trachemys scripta) to less than 9 ppm of some PCBs could override normal temperature-dependent male sex determination by producing turtles with ovaries and Mullerian ducts. Additionally, some PCB mixtures at less than 1 ppm synergize to produce ovarian development in T. scripta embryos incubated at a maleproducing temperature (Crews et al., 1995a). Conversely, in a species with genetic sex determination, the rat, exposure of embryos to a mixture of PCBs (Aroclor 1242 and Aroclor 1254) actually increases adult testis weight and sperm production and does not cause sex reversal (Cooke et al., 1996). These masculinizing effects are thought to be mediated through the induction of hypothyroidism, which leads to increased Sertoli cell proliferation. Such masculinizing effects were not noted in T. scripta exposed to PCBs.

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13

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14 In many instances, it is not clear whether endocrine alterations are the result of organizational or activational disruption. Consider the altered reproductive behaviors noted in herring gulls exposed to a mix of organochlorines (Fox et al, 1978; Gilman et al, 1978). The exposed gulls show a reduction in the defense of territories, a behavior mediated by the endocrine system. Wingfield (1987) noted that testosterone concentrations are related to the intensity of aggression in most birds. Thus, aggressive behavior could change if the regulatory effects of testosterone are altered. EDCs that act in an antiandrogenic manner could cause aggressive behavior to subside. DDE (1,1dichloro-2,2 bis(/?-chlorophenyl)ethylene), a major breakdown product of DDT (1,1,1trichloro-2,2 bis(p-chlorophenyl)ethane), is a widespread organochlorine contaminant known to bioaccumulate and biomagnify readily (Clement International Corporation, 1994). Since/?,/? '-DDE acts as an EDC via antiandrogenic mechanisms in rodents (Kelce et al, 1995), one could hypothesize that DDT exposure explains the alterations in herring gull behavior. It is not known if such behavioral aberrations are organizational or activational in origin. Alteration of Reproductive Tissues From the time of conception until death, hormones affect the morphology and physiology of an individual. Likewise, any contaminant that alters the dynamics of hormones can cause morphological and physiological alterations. This section examines the structural and functional alterations induced by EDCs in the reproductive system, the liver, and the thyroid, as they relate to the development and function of vertebrate

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15 reproduction. Table 2-1 presents a representative list of reproductive abnormalities caused by exposure of wildlife to EDCs. Reproductive System — Gonads The gonads serve both to provide gametes for reproduction and to produce steroid hormones that mediate reproductive physiology and behavior. Either of these functions can be altered activationally or organizationally by exposure to EDCs. Within a female's ovaries, several organizational, morphological changes are indicative of exposure to EDCs. In a normal female, the ovarian follicles contain one oocyte (the future ovum or egg) having a single nucleus. Mice experimentally treated preor postnatally with diethylstilbestrol (DES), a potent synthetic estrogen, exhibit polyovular follicles (Iguchi, 1992). The existence of such polyovular follicles has been used as a marker of endocrine disruption in American alligators (Alligator mississippiensis) (Guillette et al., 1994). Alligators from Lake Apopka, a contaminated lake in central Florida, exhibit polyovular follicles and polynuclear oocytes, suggesting organizational disruption as a result of exposure to estrogenic contaminants. Laboratory data support the hypothesis that such abnormalities are organizational in origin. Preliminary studies suggest that experimental exposure of neonatal alligators to p,p '-DDE can induce the development of polyovular follicles (Pickford, 1995). In many vertebrates, the developing gonad is composed of a cortical layer surrounding an inner medullary region. In males, the cortex degenerates to a single cell layer and the medullary region expands, whereas in females, the medullary region degenerates and the cells in the cortex proliferate. The retention of a gonadal cortex in males and the existence of primordial germ cells in this cortex are indicative of endocrine

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16 disruption (Fry and Toone, 1981; Guillette et al., 1994). Fry and Toone (Fry and Toone, 1981) found that embryonic exposure of gulls (Larus californicus) to low concentrations (from 2 ppm) of o,p -DDT and methoxychlor caused males to develop clusters of primordial germ cells in the cortex of the testes. Although it is not known if the altered gull testes are capable of producing viable sperm, many EDCs have the potential to organizationally alter the function of the testes. Lye et al. (1997) found that 53% of male flounder (Platichthysflesus) surveyed in a population exposed to sewage treatment effluent had testicular abnormalities including thick interstitial tissue, testicular cysts, and aberrations from the normal elongate structure. Alteration in sperm production can also be a sign of disruption, as decreased spermatogenesis is one of the most sensitive signs of reproductive toxicity in male mammals (Peterson et al., 1993). When mother rats are treated with 4-octylphenol and butyl benzyl phthalate (both common EDCs), male offspring have significantly reduced testicular size and sperm production (Sharpe et al., 1995). Recently, the reproductive impairment of male Florida panthers (Felis concolor coryi) has been attributed, in part, to exposure to EDCs (Facemire et al., 1995). Male panthers exhibit sterility, cryptorchidism (undescended testes), low sperm concentration, poor sperm motility, and a high proportion of abnormal sperm (Roelke, 1990). Studies examining semen quality determined that the abnormalities were not due to differences in steroid hormones but more likely were mediated at the testicular level (Barone et al., 1994). Although it has been speculated that these reproductive abnormalities are due to inbreeding (Roelke et al., 1993), Facemire et al. (1995) note high concentrations of DDE

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17 and PCBs in panther fat and hypothesized that the alterations in the male reproductive system could be due, at least in part, to EDC-induced organizational disruption. The hypothesis that p,p '-DDE can induce abnormal development of the male reproductive tract is supported by laboratory evidence. Experimental exposure of fetal male rats to the antiandrogenic metabolites of the fungicide vinclozolin produces characteristics similar to those seen in the panthers, including cryptorchidism and atrophic seminal vesicles (Gray et al., 1993b; Kelce et al., 1994). Because p,p '-DDE appears to act through mechanisms similar to vinclozolin (Kelce et al., 1995), the hypothesis forwarded by Facemire et al. (1995) is plausible and should be examined further. The antiandrogenic effects ofp,p '-DDE have also been implicated in altering the sexual characteristics of male American alligators. Androgens are responsible for the formation of sexual genitalia of male reptiles (Raynaud and Pieau, 1985), and wild alligators exposed to p,p '-DDE and other EDCs have significantly smaller penis size in relation to body size when compared to control animals (Guillette et al., 1996b). Reproductive System — Ducts EDCs can also cause abnormal formation and function of reproductive ducts In birds, only the left oviduct and ovary are functional in females of most species, a trait uncommon among vertebrates. The right ovary is rudimentary, and can be induced to develop into a mature (although non-functional) ovary upon hormonal manipulation (van Teinhoven, 1983). This observation leads to the prediction that birds exposed to estrogenic contaminants during gonadal organization could have both right and left ovaries and/or right and left oviducts. Indeed, Fry and Toone (1981) found that exposure of female embryonic gulls {Larus californicus) to o,p '-DDT (in concentrations

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18 comparable to those found in the wild 2 to 100 ppm) induced development of the right oviduct, which is usually rudimentary. Additionally, male breeding birds were extremely rare in the colony with heavily contaminated individuals, presumably due to the feminization or "neutering" effects of DDT exposure (Fry, 1995). Perhaps the most well known example of reproductive "endocrine disruption" is the effects of organochlorines (namely DDT) on eggshell thickness in birds (Elliot et al., 1988; Burger et al., 1995). The mechanism of this eggshell thinning is still debated but appears to be associated with contaminant-induced alterations of the enzymatic functioning of the oviduct. Ducks exposed to p,p -DDE have reduced prostaglandin synthesis in the eggshell gland mucosa, increased calcium content in eggshell gland mucosa, and decreased calcium in the shell gland lumen (Lundholm, 1994). Thus, the eggshell-thinning effects of DDT are consistent with the inhibition of prostaglandin synthesis in eggshell gland mucosa by/?,/? '-DDE (Lundholm 1994). Mammalian wildlife also exhibit reproductive tract alterations as a result of exposure to EDCs. Wild mink populations have declined in the Southeast, and these declines have been attributed to PCB-induced reproductive dysfunction (Osowski et al., 1995). Indeed, the reproductive tract of female mink is highly sensitive to organochlorineinduced impairment, as Patnode and Curtis (1994) found that 3,3 ',4,4', 5,6' hexachlorobiphenyl (a coplaner PCB) caused significant upregulation of uterine estrogen receptors and, most notably, significantly increased the progesterone receptor dissociation constant. Reproductive impairment in seals has also been attributed to PCB exposure. Female Baltic seals have exhibited low reproductive output in recent years, and this has

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19 been attributed largely to the occurrence of uterine occlusions (Helle, 1989). Such occlusions are thought to be caused by PCB exposure, as peak PCB levels (up to 100 ppm in blubber) in the Baltic seals preceded a rapid increase in the frequency of uterine occlusions (Helle, 1989). The detrimental impact of PCBs on seal reproduction has been documented experimentally in harbor seals (Reijnders, 1986). Reijnders (1986) found that a diet of PCB-contaminated food caused female harbor seals to resorb their embryos. This has led researchers to hypothesize that PCB-induced uterine occlusions result after an initial interruption of pregnancy around the time of implantation (Reijnders, 1984; Helle, 1989), a phenomenon dependent on the correct hormonal environment. Liver The liver is an important organ associated with reproduction in that it synthesizes proteins, such as vitellogenin and sex-hormone binding globulin, necessary for normal hormonal homeostasis in the plasma. This organ also converts steroids to more hydrophilic excretory products. A recently observed, much publicized, hepatic alteration following exposure to EDCs is the abnormal synthesis of vitellogenin (Vg) in nonmammalian male vertebrates. Vitellogenin is an estrogen-dependent protein that is the major component of yolk in eggs of oviparous vertebrates (van Teinhoven, 1983). Females normally produce Vg during the reproductive stages of egg assimilation, but Vg can also be induced by exogenous estrogens such as estradiol1 70 or estrogenic contaminants in males and non-reproductive females. For instance, adult male frogs (Xenopus laevis) and turtles (Trachemys scriptd) produce significant levels of Vg when treated with the environmental estrogen o,p '-DDT (Palmer and Palmer, 1995). The consequences of male Vg secretion are unknown, but the presence of Vg in males is

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20 correlated with a reduction in testis size in male trout (Oncorhynchus mykiss) exposed to estrogenic contaminants (Jobling et al., 1996). Vitellogenin has also been detected in juvenile male and female guppies (Poecilia reticulata) exposed to 0hexachlorocyclohexane, a persistent environmental contaminant (Wester et al., 1985). Here, the production of Vg was associated with a dose-dependent increase in hepatocellular basophilia, indicating that P-hexachlorocyclohexane induced both structural and functional damage. Similar hepatic structural damage is noted in winter flounder (Pleuromctes americanus) exposed to pulp and paper mill effluent (Khan et al., 1994). Vitellogenin production by cultured hepatocytes has been used as an indicator of the estrogenic potency of chemicals. Six different phytoestrogens (biochanin A, coumestrol, daidzein, equol, formononetin, or genistein) stimulate Vg production in rainbow trout hepatocytes, the phytoestrogens having approximately one-thousandth the potency of 17p-estradiol (Pelissero et al., 1993). Additionally, a number of EDCs stimulate in vitro Vg synthesis. Several studies indicate that alkylphenolic compounds are estrogenic based on the ability to stimulate Vg synthesis from fish hepatocytes (Jobling andSumpter, 1993; White et al., 1994; Jobling et al., 1996). Alkylphenols are environmentally persistent and the parent compounds, alkylphenol polyethoxylates, are widely used as nonionic surfactants in detergents, paints, herbicides, pesticides, and many other formulated products. Over 300,000 tons of alkylphenol polyethoxylates are produced worldwide annually, making this the second largest group of nonionic surfactants in commercial production (Chemical Manufactures Association, 1994). It is estimated that 60% of these manufactured compounds end up in the aquatic environment after sewage treatment (Giger et al., 1987). Therefore, the occurrence of Vg in the

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21 plasma of male rainbow trout (Oncorhynchus mykiss) (Purdom et al., 1994), carp (Cyprinus carpio) (Folmar et al., 1996), and flounder {Platichthys flesus) (Lye et al., 1997) following exposure to sewage effluent could be mediated, in part, by alkylphenolic compounds such as nonylphenol and octylphenol. Vitellogenin appears to be a good in vitro and in vivo indicator of exposure to estrogenic contaminants in many instances. However, the absence of Vg does not indicate the absence of endocrine disruption, as many compounds could act through mechanisms other than the estrogen receptor (see Cellular and Molecular Mechanisms of Disruption). Vitellogenin could also provide a route of exposure for metal accumulation in oocytes. Recent work on red drum (Sciaenops ocellatus) has discovered that Vg in vivo readily binds calcium, magnesium, zinc, iron and copper (Ghosh and Thomas, 1995). Furthermore, Ghosh and Thomas (1995) found that cadmium-vitellogenin injections resulted in cadmium incorporation into the ovaries. Thus, vitellogenic animals could be particularly sensitive to metal exposure. The possible role of Vg in transporting other xenobiotics into the cytoplasm of eggs is unknown. Thyroid The thyroid is often thought of as exclusively associated with the control of metabolism, but thyroid hormones are involved in numerous reproductive events, varying from stimulation of gonadal maturation in fish (Leatherland, 1994) and amphibians (van Teinhoven, 1983) to regulation of metamorphosis in amphibians or migration behavior in birds (van Teinhoven, 1983). Like steroid hormones, thyroid hormones are regulated by the hypothalamic-pituitary axis. The hypothalamus secretes thyrotropin-releasing hormone, which stimulates thyroid-stimulating hormone secretion from the anterior

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22 pituitary (adenohypophysis). Thyroid-stimulating hormone then travels via the circulatory system to the thyroid where iodine is actively transported into the gland and used to synthesize thyroxine (T 4 ) and triiodothyronine (T 3 ). Hypothyroid animals have low serum thyroid hormone concentrations which result in spontaneous abortions, stillbirths, or congenital defects (Burrow, 1986). Animals said to be hyperthyroid have excessive thyroid hormone production which can induce amenorrhea (loss of reproductive cyclicity) and reduced fertility (Burrow, 1986). Hyperthyroid animals have altered sex-hormone profiles as a result of increased levels of sex-hormone binding globulin (Tulchinsky and Chopra, 1973) and increased concentrations of circulating estrogens (Akande and Hockaday, 1972). Because sexhormone binding globulin (SHBG) has a greater affinity for testosterone than 170estradiol, an increase in SHBG causes relatively more 1 7P-estradiol to be free and active. Low reproductive success of herring gulls (Larus argentatus) in Lake Ontario was originally attributed solely to eggshell thinning, but further investigation revealed that other pollutant-induced factors contributed to the reported reproductive failure (Mineau et al., 1984). Among these factors, thyroid dysfunction was credited with altering normal reproductive behaviors such as egg incubation. Surveys of Lake Ontario herring gulls revealed several signs of hypothyroidism: decreased thyroid colloid content, increased thyroid epithelial cell height, decreased thyroid follicular diameter, and decreased circulating T 3 and T 4 concentrations (Rattner et al, 1984). Moccia et al. (1986) obtained similar results for herring gulls from other colonies in the Great Lakes basin, and both Rattner et al. (1984) and Moccia et al. (1986) attributed the hypothyroidism to a forage fish-borne goitrogenic etiology rather than iodine deficiency. In ovo hypothyroidism has

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23 been noted in cormorants (Phalacrocorax carbo) exposed to PCBs, dibenzo-p-dioxins (PCDDs), and dibenzofiirans (PCDFs), and this state could have a role in the observed low breeding success of cormorant colonies in the Netherlands (van den Berg et al., 1994). Embryos and juveniles could be particularly sensitive to contaminant-induced thyroid dysfunction, as developmental exposure to low levels of PCBs can cause organizational disruption of the thyroid (Gray et al, 1993a; Goldey et al, 1995). Some of the effects of EDCs on the thyroid are activational. Brouwer et al. (1989) found that common seals (Phoca vitulina) fed PCB-contaminated fish had significantly lower T 3 and T 4 concentrations compared to seals fed non-contaminated fish. Similarly, fish (Oreochromis mossambicus) exposed for 20 days to 1,2,3,4,5,6hexachlorocyclohexane (BHC) displayed goiter formation, decreased colloid, and atrophied follicles, but these characteristics normalized when BHC-exposed fish were subsequently exposed to clean water (Pandey and Bhattacharya, 1991). The effects of particular pesticides on thyroid function can be complex. For instance in the freshwater catfish (Clarias batrachus), endosulfan induces a decrease in T 3 and an increase in T 4 , malathion causes decreased T 3 and no change in T 4 , and carbaryl provokes an increase in T 3 and a decrease in T 4 (Sinha et al, 1991). What direct effects these alterations in plasma T 3 and T 4 have on reproductive activity have not been reported. Cellular and Molecular Mechanisms of Disruption There are numerous mechanisms through which contaminants can alter the hormonal milieu in an animal. If altered, each mechanism can produce a unique disruption profile. For example, consider two well defined case studies of animals exposed to EDCs: the white sucker fish (Catostomes commersoni) of the Great Lakes and the American

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24 alligators {Alligator mississippiensis) in Florida. White sucker fish exposed to bleached kraft mill effluent (BKME) have reduced circulating concentrations of 1 7p-estradiol (E 2 ) and testosterone (T) when compared to control fish (McMaster et al., 1991). These decreased hormone concentrations are correlated with a lower reproductive success of BKME-exposed fish. McMaster (1995) found that the decreased E 2 and T concentrations could be attributed partially to decreased steroidogenic enzyme activity (aromatase) and a reduction in the cholesterol substrate used as the precursor of steroids. Alligators from Lake Apopka, Florida, also show decreased reproductive success apparently as a function of endocrine disruption, but the hormonal profile and mechanisms of disruption differ from those of white sucker fish. Lake Apopka is contaminated with dicofol, DDT, DDD (l,l-dichloro-2,2-bis(p-chlorophenyl)ethane), and p,p'-UDE (U.S. EPA, unpublished report), and alligator eggs from Lake Apopka contain significant residues of toxaphene, dieldrin, p,p '-DDE, p,p -DDD, /raws-nonachlor, and PCBs (Heinz et al., 1991). Reproductive success is dramatically lower on Lake Apopka compared to other Florida lakes (Woodward et al., 1989; Woodward et al., 1993; Masson, 1995). Plasma samples from juvenile Lake Apopka alligators reveal an altered steroid profile: males have lower T and females have higher E 2 when compared with control animals (Guillette et al., 1994). In vitro culture of gonads found that testes from Lake Apopka males secreted more E 2 but did not differ from controls in T synthesis (Guillette et al., 1995b). Additionally, female ovaries generated significantly less E 2 compared to controls. Therefore, altered gonadal steroidogenesis does not explain the plasma steroid abnormalities in Apopka alligators, and the contaminant-induced alteration differs from that of the Great Lakes white sucker fish.

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25 To explain the variation in particular case studies, we must examine the mechanisms through which contaminants can alter the endocrine system. Contaminants could disrupt normal endocrine function by altering (a) the hypothalamic-pituitary axis of endocrine control, (b) the activity of steroidogenic enzymes, (c) the function of steroid binding molecules such as sex-hormone binding globulin, (d) the activity of hormone receptors by acting as a hormone agonist or antagonist, or (e) the hepatic clearance rate of steroids. The relationship between these functions is depicted in Figure 2-2. In considering the ways that contaminants alter an animal's hormonal environment, first consider affects on the neuroendocrine control of steroid hormone secretion. The control of reproduction is an extremely complex integration of external stimuli into internal physiological signals. Basically, external stimuli are transduced through numerous neural pathways in the brain, and one of these pathways assimilates information in the hypothalamus. Here hypothalamic neurons secrete the neurohormone gonadotropinreleasing hormone (GnRH) into a vessel (hypothalamic-pituitary portal) that carries GnRH directly to the anterior pituitary gland. The anterior pituitary gland synthesizes gonadotropins (follicle stimulating hormone, luteinizing hormone) which are glycoprotein hormones that are released into the circulatory system. The gonadotropins travel to the gonads where they stimulate gametogenesis and sex-specific steroid hormone secretions that result in reproductive maturation (in the case of juveniles) or reproductive stimulation (in the case of adults). In a normal animal, feedback circuits control the secretion of gonadotropic hormones.

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Figure 2-2. Several components of the reproductive endocrine system with specific points where endocrine disruption can occur: (a) hypothalamic-pituitary axis, (b) gonadaladrenal steroidogenesis, (c) steroid receptors, (d) sex-hormone binding globulin production and interactions, and (e) hepatic steroid metabolism.

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27 Cytoplasmic SBP

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28 Toxicants that alter gonadotropin release will change an animal's steroid-hormone profile. For instance, chlordimeform (an acaricide recently banned in the United States but used in other parts of the world) is thought to influence endocrine regulation adversely by interfering with the activity of the hypothalamus (Goldman et al., 1990). Goldman et al. (1990) found that chlordimeform treatment in male rats decreased serum gonadotropins and decreased the hypothalamic GnRH response to norepinephrine stimulation. Because brain norepinephrine increases after a single dose of chlordimeform (Bailey et al., 1982), the endocrine-disrupting effects of chlordimeform are attributed to desensitization of the hypothalamic adrenergic receptors (Goldman et al, 1990). Many other EDCs could alter reproductive endocrinology by changing the hypothalamic and pituitary control of steroid production, but there is no data available for such alterations in wildlife species. Steroid-hormone dynamics are dictated by a number of enzymes that convert steroids. The steroidogenic pathway begins with cholesterol and, through the actions of numerous steroidogenic enzymes, progresses through progestins, androgens, and estrogens (see Figure 2-3). Induction or repression of the genes encoding steroidogenic enzymes can result in altered hormone production. Therefore, some of the endocrine disruption noted in wildlife could be due to altered transcription of genes encoding steroidogenic enzymes. This hypothesis is supported by several lines of evidence: first, the endocrine-disrupting effects of many contaminants appear to be mediated independent of steroid receptor binding (McLachlan, 1993) and, second, transcriptional control of steroidogenic enzymes is extremely labile during embryonic and neonatal periods (Gustafsson, 1994). Few studies have examined the activity and transcription rate of steroidogenic enzymes after EDC exposure, but it is clear that many compounds have the

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29 potential to alter steroidogenesis in wildlife. For example in the American alligator, the actions of the enzyme aromatase (P450, rom ; the enzyme responsible for converting androgens to estrogens) were blocked in vivo after embryonic exposure to several aromatase inhibitors (Lance and Bogart, 1992). The treated alligators had inhibited ovarian development, but were not masculinized. Cholesterol P450SSC T 30-HSD P450c21 P450cll Pregnenolone ^ Progesterone Deoxycorticosterone ^ Corticosterone Isomerases P450cl7 P450cl7 T 30-HSD T 170H Pregnenolone 17-OHProge Isomerases P450cl7 P450c21 p^u • 11 Deoxycortisol ^ Cortisol P450cl7 3|3-HSD f Dehydroepianclrosterorr — Androstenediom Isomerases Reductase P450arom 17Kefa^ W' Testosterone17p-&tradiol Figure 2-3: Steroidogenic pathway illustrating the P450 enzymes (bold) associated with the production of specific steroid hormones. Based on Miller (1988). After steroid hormones are produced and secreted, they circulate throughout the body attached to carrier molecules. The most prominent of these molecules is sexhormone binding globulin (SHBG) which is produced by the liver. SHBG protects the steroid hormone from hepatic degradation and helps direct the hormone to target organs (Hammond and Boccinfuso, 1995). When attached to SHBG, steroid hormones are

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30 biologically inactive. Sex-steroid binding proteins are found in animals from all vertebrate classes, but little consideration has been given to the effects of EDCs on SHBG presence and function. One effect of antiestrogenic drugs is to increase the circulating level of SHBG. Droloxifene, an antiestrogenic drug used to treat breast cancer, increases the plasma concentration of SHBG (Geisler et al., 1995). An increase in SHBG concentration causes more native hormone to be bound in a biologically inactive form, thereby decreasing the amount of steroid hormone that reaches the inside of a cell. Little attention has been given to the binding of environmental EDCs to SHBG and other plasma proteins. If certain EDCs do not bind to SHBG, these contaminants would have an elevated availability to target cells compared to native steroids. For instance, even though the affinity of the xenoestrogen o,p -DDT for the estrogen receptor is 1 000-fold lower than E 2 , the estrogenicity of o,p -DDT would be enhanced if it did not bind to SHBG. Indeed, Arnold et al. (1996c) found that purified human SHBG selectively decreased the transcriptional activity of E 2 compared to the xenoestrogens o,p -DDT and octylphenol. The addition of alligator or human serum mimicked the results of purified SHBG, suggesting that SHBG and other proteins in serum selectively bind E 2 but do not bind certain EDCs to a similar extent (Arnold et al., 1996c). Therefore, the binding of EDCs to plasma proteins such as SHBG is a major determinant of the EDCs bioavailability and potency in target cells. Contaminants could also disrupt endocrine function by binding directly to hormone receptors. Steroid hormones are derivatives of cholesterol and easily diffuse through cellular membranes. Once inside the nucleus, steroid hormones bind to a hormone-specific receptor. The hormone-receptor complex binds directly to DNA, activating the

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31 transcription of hormone-induced genes. Many environmental contaminants can bind directly to hormone receptors, stimulating (hormone agonist) or blocking (hormone antagonist) the expression of hormone-induced genes (McLachlan, 1993). A recent study indicates that combinations of relatively weak environmental estrogens can have a synergistic effect on the activation of the estrogen receptor (Arnold et al., 1996b). These complex interactions between multiple environmental contaminants and steroid receptors are not yet understood and are currently an area of interest and importance. Although steroid hormones often are converted and "recycled" into other steroid hormones (see Figure 2-3), steroids are also excreted after hepatic conversion to hydrophilic compounds (primarily through Phase II conjugation reactions). For instance, men excrete approximately 50 ug of testosterone conjugates per day in the urine (Lipsett, 1986). Based on this principal, urine and fecal samples are used routinely for analysis of reproductive status in numerous animal species. An alteration in normal steroid-hormone excretion rates will alter the hormonal profile of an animal, and many environmental contaminants cause such alterations. It has long been recognized that the liver is the primary site of xenobiotic metabolism and steroid metabolism. Pioneering work by Conney and Klutch (1963) suggested that the administration of many drugs could alter the metabolism of in vivo steroids by altering hepatic enzyme systems. Welch et al. (1967) found that insecticides could also alter steroid metabolism in the rat liver. Specifically, organic phosphorothionate insecticides, such as parathion and malathion, inhibit liver microsomal hydroxylation (and, thus, excretion) of testosterone, whereas halogenated hydrocarbon insecticides, such as DDT and chlordane, stimulate the hydroxylation of steroids. In the case of chlordane, Levin et al. (1968) found that this decreased the

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32 stimulatory effects of androgens on seminal vesicle weight by enhancing androgen metabolism. Welch et al. (1971) found similar metabolism alterations in females. Rats and mice exposed to a number of halogenated hydrocarbon insecticides (chlordane, dieldrin, heptachlor, lindane, p,p -DDE, p,p '-DDD, or toxaphene) had stimulated hepatic metabolism of estrone and a decrease in uterine wet weight. These reactions to contaminants appear to be phylogenetically conserved, as Peakall (1967) noted that DDT and dieldrin exposure increased the excretion of testosterone's and progesterone's polar metabolites in birds. Many of the hepatic alterations induced by environmental contaminants could be sex specific. For example, research in rats indicates that embryonic steroid exposure establishes sexual dimorphism of steroid-metabolizing enzymes in the liver (Lucier et al., 1982; Lucier et al ., 1985). Male and female rodents exhibit different patterns of hepatic steroid metabolism, and these patterns are organized during development by exposure to androgens (Gustafsson, 1994). For instance, gonadectomy of males results in a femalepattern of steroid degradation, whereas administration of testosterone to a gonadectomized male results in male-pattern steroid degradation (Jansson et al., 1985). It is clear that sex steroids affect liver metabolism, such that exogenous androgens can masculinize a female liver and exogenous estrogens can feminize a male liver (Gustafsson, 1994). It is therefore plausible that exposure to EDCs, many of which mimic sex hormones, could alter normal hepatic sexual dimorphism.

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33 Conclusions In considering the effects of EDCs on wildlife reproduction, we have explored topics ranging from evolutionary considerations to molecular mechanisms of disruption. After briefly considering the topic at these scales, one theme emerges many anthropogenic compounds released into the environment have adverse effects on the development and function of the reproductive system of wildlife species. Future studies should focus on the specific mechanisms of disruption, giving consideration to "mechanisms" from gene to ecosystem. Only with a better understanding of these mechanisms can progress be made in the discipline of endocrine toxicology.

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34 u Q U B O o -*-» a. •c u « 3 D. 1 s 5 B U s •c u c M U +-> 03 03 03 On w R» RJ 2 S t> S —J U O o rj o -g -c Jill 1511. H n m m .M H H H > < > < > -§3 in 2 I 1 o 2 CO •a 3 03 1 u u < On i IS C 03 S 03 a p On ^-v On Tt ~ On . E T3 c o H Q » 0 u ^ Oh © 2 •5 60 g »> w I u a OJ u ;c 5 § a o S3 V a q

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35

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CHAPTER 3 SEX-STEROID AND THYROID HORMONE CONCENTRATIONS IN JUVENILE ALLIGATORS (ALLIGATOR MISSISSIPPIENSIS) FROM CONTAMINATED AND REFERENCE LAKES IN FLORIDA Introduction Growth and reproduction are two of the fundamental variables that dictate the life history strategy of all animals. In a simplistic model, animals must survive and grow until they reach an optimal size that maximizes reproductive output (Stearns, 1 992). Both growth and reproduction are under hormonal control and, as such, maintaining normal hormone concentrations is critical. There are numerous hormones that are involved in the complex regulation of somatic growth and development. Among these, the thyroid hormones appear to have a pivotal role in the growth of all vertebrates (Norris, 1997). The thyroid hormones thyroxine (T 4 ) and triiodothyronine (T 3 ) are necessary for normal differentiation, maturation, and growth of many systems including the central nervous system and skeletal systems (McNabb and King, 1993). Thyroid hormones also have a cooperative role in regulating the reproductive activities of vertebrates, but sex hormones are the most potent regulators of reproductive cyclicity in vertebrates (Norris, 1997). Testosterone (T) and estradiol17p (E 2 ) are particularly important in regulating the development and function of reproductive activity and behavior in both sexes. Several recent case studies have noted altered reproductive activity as a result of exposure to environmental contaminants that change the hormonal regulation of growth Note: This chapter is published in Environmental Toxicology and Chemistry (Crain et al 1997 a) 36

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37 and reproduction. Numerous studies have been conducted on white sucker (Catostomus commersoni) to determine the causes and mechanisms of reduced reproductive fitness of these fish. Depressed circulating steroid levels are noted in white suckers exposed to pulp mill effluent (McMaster et al, 1991; Munkittrick et al. ( 1991; Gagnon et al, 1994a), and these reductions correlate with reduced gonadal development, reduced expression of secondary sexual characteristics, delayed maturity, decreased egg size, and reduced fecundity with age (Gagnon et al., 1994b; Munkittrick et al., 1994a; Munkittrick et al., 1994b). A similar correlation between contaminant exposure and abnormal reproductive function has been noted for alligators {Alligator mississippiensis) from Lake Apopka, Florida, USA. The alligator population on Lake Apopka precipitously declined from 30 juveniles/km to 4 juveniles/km between 1980 and 1983, and the mean clutch viability (eggs hatched / eggs laid) dropped from 54% in 1983 to 13% in 1986 (Woodward et al., 1993). These declines followed a 1980 spill of a pesticide mixture that was primarily composed of dicofol but also had as much as 15% DDT, DDD, and DDE (Environmental Protection Agency, 1994). Recent studies have indicated that plasma steroid hormone concentrations are abnormal in the alligators of Lake Apopka (Guillette et al., 1994; Guillette et al., 1996b; Guillette et al., 1997), and these abnormal steroid concentrations correlate with altered gonadal morphology (Guillette et al., 1994) and smaller phallus size (Guillette et al., 1996b; Guillette et al., 1997) in juvenile alligators. The hormonal abnormalities in the juvenile alligators of Lake Apopka include depressed T in males (Guillette et al., 1994; Guillette et al., 1996b; Guillette et al, 1997) and elevated E 2 concentrations in females (Guillette et al., 1994). Whereas males and females have both T and E 2 circulating in their blood, it is the relative ratio of the two steroids that dictates

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38 reproduction (Bogart, 1987). For instance, females have more E 2 relative to T and this is thought to affect ovarian differentiation as well as female reproductive function. Many environmental chemicals can alter normal steroid concentrations by changing hypothalamic-pituitary control of steroid synthesis, gonadal steroid synthesis, hepatic steroid conversion and excretion rates, and receptor-mediated responses (see CHAPTER 2 for review). This study examines plasma concentrations of the sex-steroids E 2 and T and the thyroid hormones T 3 and T 4 in three populations of juvenile American alligators in Florida. Two of these populations are in historically contaminated lakes, Lake Apopka and Lake Okeechobee. The third population is in the relatively pristine environment of Lake Woodruff National Wildlife Refuge. Previous studies have noted altered circulating concentrations of steroid hormones in the alligators of Lake Apopka. The purposes of this study are to: (1) reevaluate circulating steroid hormones in Lake Apopka alligators, relative to the reference lake and another contaminated lake, (2) analyze, for the first time, circulating thyroid hormone concentrations in alligators from these three lakes, and (3) evaluate the steroid and thyroid hormone concentrations relative to body size. If alligators exposed to environmental contaminants are affected by endocrine disrupters, this may be evident in the circulating concentrations of sex steroids or thyroid hormones.

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39 Materials and Methods Study Sites and Sample Collection The three lakes in this study were chosen based on historical and ongoing episodes of exposure to environmental contaminants. Lake Woodruff (lat. 29°06'N, long. 81°25'W; sampled on April 25, 1995) is in Lake Woodruff National Wildlife Refuge and is considered a relatively pristine, natural environment. Lake Apopka (lat 28°40'N, long 81°38'; sampled on May 2, 1995) is 1.5 miles downstream from an EPA Superfiind site where an industrial spill of dicofol and DDT occurred in 1980 (Environmental Protection Agency, 1994). Lake Apopka has also received substantial agricultural and municipal runoff (Sengal and Pollman, 1991; Schelske and Brezonik, 1992) and several studies have documented reproductive endocrine disruption in the alligators of Lake Apopka (Guillette et al., 1994; Guillette et al., 1996b). Lake Okeechobee (lat. 26°56'N, long. 80°49'W; sampled on May 3, 1995) receives significant agricultural runoff that has led to eutrophication (Federico et al., 1981), and Lake Okeechobee is known to be contaminated with mercury (Jurczyk 1993; Sundlof et al, 1994) and pesticides (Turnbull et al, 1989), but endocrine-disruption has not been studied in the alligators of Lake Okeechobee. Juvenile American alligators (Alligator mississippiensis) from 60-140 cm total length were hand captured from an airboat at night. Animals of this size are 2-6 years of age (Woodward et al, 1992). To minimize temporal effects of sampling, all samples were collected during a 9-day time span. A blood sample (1-2 ml) was collected in a heparinized vacutainer from the post-cranial sinus within 30 min of capture. A previous study found that capture stress has no influence on the acute plasma steroid concentrations

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40 in alligators (Guillette et al., 1997). Individuals were sexed using the phallic criteria previously defined (Allsteadt and Lang, 1995; Guillette et al., 1996b), and body measurements (snout-vent length and total length) were recorded. Animals were released in the vicinity of their capture. Blood samples were stored on ice for 10-12 hrs until centrifugation at 1500 g for 15 min. Plasma was stored at -72°C until hormone analysis. Steroid Hormone Radioimmunoassays E 2 and T were analyzed using radioimmunoassays previously validated for alligator plasma (Folmar et al., 1996; Guillette et al., 1997). Briefly, plasma (100 u.1) was extracted 2x with diethyl ether (5 ml). The ether extracts were dried under constant air stream for 15 min. The dried samples were resuspended with borate buffer (100 u.1; 0.5 M; pH=8.0). To reduce non-specific binding, 100 uJ of borate buffer with bovine serum albumin (Fraction V; Fisher Scientific) at a final assay concentration of 0. 15% for T and 0. 19% for E2 was added to each tube. Antibody was then added (200 p.1; final concentration of 1 :55,000 for E 2 , 1 :25,000 for T; Endocrine Sciences). Finally, 100 ul of radiolabeled steroid was added (12,000 cpm per 100 ul; [2,4,6,7,1 6, 173 H]Oestradiol @ 1 mCi/ml; [l,2,6,73 H]Testosterone @ 1 mCi/ml; both from Amersham International, Arlington Heights, IL). For standard tubes, either T or E 2 was added at 0, 1.56, 3.13, 6.25, 12.5, 25, 50, 100, 200, 400, and 800 pg/tube. Tubes were vortexed for 1 min and incubated overnight at 4°C. All standards and samples were prepared in duplicate. Bound-free separation was accomplished by adding 500 ui 5% charcoal 0.5% dextran and immediately centrifuging at 1 500 g at 4°C for 30 min. 500 ul was then added

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41 to 5 ml scintillation cocktail and the tubes were counted on a Beckman scintillation counter. The steroid RIAs were validated using both plasma dilutions and internal standards. For internal standards, 100 ul of steroid-free plasma (steroids removed by incubating plasma with activated charcoal, 10% w/v, and collecting the supernatant after centrifugation at 10,000 g for 15 min) was spiked with steroid (6.25, 12.5, 25, 50, 100, 200, 400, 800 pg for T; 3.125, 6.25, 25, 50, 100, 200 pg for E 2 ). The internal standards were extracted and assayed as described above. For the T plasma dilutions, 3, 6, 12, 25, 50, 75, and 100 u.1 of plasma pool were placed in separate tubes, and the volumes were brought to 100 ul with steroid-free plasma. For the E 2 plasma dilutions, 25, 50, 75, and 100 ul of the plasma pool were pipetted into separate tubes and the volumes were brought to 100 ul with steroid-free plasma. The plasma dilutions were extracted and assayed as described above. All standards, internal standards, and plasma dilutions were performed in duplicate. Parallelism of the internal standard curve and plasma dilution curve with the standard curve was tested using the test for homogeneity of slopes (Super Anova, Abacus Concepts, Berkeley CA). Figures 3-1 and 3-2 present the validation curves for the steroid hormones.

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42 120 1 10 100 1000 Testosterone (pg) Figure 3-1 Radioimmunoassay validation for the measurement of testosterone in alligator plasma. The slopes of the internal standards and plasma dilutions were not significantly different from the standards (p=0. 184).

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43 ion 1UU Standard Cure °\ Rasma D'lutions 8 80 ^ \. Internal Standards H 60 00 20 0 1 I I — I 10 100 1000 Estradd (pg) Figure 3-2 Radioimmunoassay validation for the measurement of estradiol in alligator plasma. The slopes of the internal standards and plasma dilutions were not significantly different from the standards (p=0.73 1). Thyroid Hormone Radioimmunoassays Unextracted plasma (100 uJ) was used for determination of circulating concentrations of T 3 and T 4 . Assay buffer (100 uJ of 0.2 M borate buffer; pH=8.0) was added to each tube, followed by the addition of 200 uf of BSA/y-globulin/ANS buffer (1% bovine serum albumin, 1.25 mg/ml y-globulin, 2 mg/ml 8-alinino-l-napthalene-sulfonic acid; all from Sigma Chemical Co., St. Louis, MO). Antibody (100 p.1) was added to give

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44 a final concentration of 1 :8,000 for T 3 antisera and 1 :2,000 for T 4 antisera (Endocrine Sciences, Calabasas Hills, CA). Finally, 100 ul of iodinated T 3 (40,000 cpm/tube) or T 4 (50,000 cpm/tube) was added to the tubes (both from New England Nuclear; L-[ 125 I]thyroxine @ 1250 uCi/ug; L-3,4,3'-[ 125 I]-triiodothyronine @ 1200 uCi/ug). Standards were prepared in 100 ul of assay buffer and substituted for this constituent in the assay described above. For T 4 , 0, 25, 50, 100, 200, 400, 800, 1600, 3200, 6400, and 12800 pg/tube were prepared, whereas 0, 6.25, 12.5, 25, 50, 100, 200, 400, 800, 1600, and 3200 were prepared for the T 3 assay. Tubes were vortexed and incubated at 37°C. After 2 hrs, the tubes were incubated at room temperature for 1.5 hr. Bound-free separation was accomplished by adding 1.5 ml of 60% saturated ammonium sulfate to each tube, vortexing, and centrifuging at 1500 g for 30 min. The supernatant (containing the free hormone) was discarded and the pellet resuspended in a 9: 1 1 mixture of saturated ammonium sulfate and assay buffer with 0.5% BSA. The tubes were then vortexed, centrifuged at 1500g for 30 min, and the supernatant discarded. The pellets were counted on a Beckman gamma counter. The thyroid hormone RIAs were validated using plasma dilutions. Plasma (25, 50, 75, and 100 ul) was aliquated from a plasma pool, and the dilutions were extracted and assayed as described above. Parallelism of the internal standard curve and plasma dilution curve with the standard curve was tested using the test for homogeneity of slopes (SuperAnova, Abacus Concepts, Berkeley CA). Figures 3-3 and 3-4 present the validation curves for the steroid hormones.

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45 o o o m m 100 80 60 40 20 Standard Curve Plasma Dilutions 10 100 1000 Triiodothyronine (pg) 10000 Figure 3-3 Radioimmunoassay validation for the measurement of triiodothyronine (T 3 ) in alligator plasma. The slope of the plasma dilutions was not significantly different from that of the standards (p=0.71 1). 100 -, 1 10 100 1000 10000 Thyroxine (pg) Figure 3-4 Radioimmunoassay validation for the measurement of thyroxine (T 4 ) in alligator plasma. The slope of the plasma dilutions was not significantly different from that of the standards (p=0.454).

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Statistical Analysis 46 Hormone concentrations were determined using commercially available software (Microplate manager III; Biorad, Hercules, CA, USA). All statistical tests were performed with SuperAnova (Abacus Concepts; Berkeley, CA, USA). Mean body size was not significantly different for males (p=0. 19) or females (p=0.55) from the three lakes. Thus, an analysis of variance (ANOVA) was conducted to determine whether differences in hormone concentrations existed among alligators of the 3 lakes. Prior to this analysis, hormone concentrations were log transformed to obtain homogeneity of variance. Fisher's PLSD was used as a post-hoc test. To examine the relationship of body size to hormone concentrations, linear regression analysis was conducted for each sex on each lake. Results Mean Hormone Concentrations All the animals sampled in this study represent one life history stage juvenile. Therefore, we chose to first analyze the data by comparing hormone concentrations in alligators from the three lakes (see Table 1). There was no difference in mean E 2 concentrations among males (F=1.62; df=2, 43; p=0.21) or females (F=0.50; df=2, 40; p=0.61) from the three lakes. Mean T concentrations were not different among females from the three lakes (F=1.36; df=2, 40; p=0.27), but T concentrations were significantly different among males (F=4.96; df=2, 44; p=0.01). Woodruff males had significantly more T compared to males from either Apopka (p=0.009) or Okeechobee males (p=0.01).

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47 For triiodothyronine (T 3 ) concentrations, there was no difference among females (F=0.74; df=2, 40; p=0.48) or males (F=l. 10; df=2, 43; p=0.34) from the 3 lakes. Thyroxine (T 4 ) concentrations were significantly different among lakes for males (F=5.67; df=2, 43; p=0.007) but not for females (F=3. 16; df=2, 39; p=0.053). For males, T 4 concentrations were significantly lower in Woodruff animals compared to Okeechobee alligators (p=0.002). Table 3-1 . Mean plasma concentrations (+ 1SE) of the thyroid hormones T 3 and T 4 (ng/ml) and the steroid hormones E 2 and T (pg/ml) in male and female alligators from the three lakes. Woodruff §11! male female Apopka male female Okeechobee male female T 3 2.1±0.4 1.6+0.2 1.7+0.3 2.3+0.4 2.2+0.3 2.0+0.3 T 4 13.9±2.3 13.5+2.2 17.3+1.9 18.8+3.1 21.8+2.0 18.9+2.2 E 2 14.7±3.9 25.4±4.8 21.6+5.1 23.9+7.1 12.0+2.6 28.7+4.3 T 962.6+283.9 86.4+10.1 162.7+71.8 86.5+10.1 218.8+108.0 66.6+11.2 Body Size and Hormone Concentrations The relationship between body size and hormone concentrations was examined using regression analysis (see Table 3-2). For the steroid hormones, there was no apparent relationship between E 2 and body size in males from Woodruff and Okeechobee, but a relationship was apparent in males from Apopka (see Figure 3-5). However, this relationship disappears (r 2 =0.031, p=0.25) if the one Apopka male that is larger than 105 cm is removed from the analysis. T was positively related to body size in males from

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48 Okeechobee and Woodruff, whereas there was no such relationship in Apopka males (see Figure 3-6). Among females, a clear positive relationship existed between E 2 and body size for Woodruff and Okeechobee animals, but females from Apopka had no apparent relationship between E 2 and body size. T was positively related to body size in females from Woodruff, but T had no such relationship to body size in females from Apopka or Okeechobee. For the thyroid hormones, there was a negative relationship between both T 3 and T 4 and body size for Woodruff males and females (see Figures 3-7 and 3-8). Additionally, T 3 was negatively related to body size in males from Apopka and Okeechobee, but females from these lakes showed no clear relationship. For T 4 , Apopka females showed a clear negative relationship with body size, whereas Apopka males and Okeechobee males and females exhibited no apparent relationship.

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49 Table 3-2. Results for the linear regression analysis of hormone concentrations as a function of total body size in male and female alligators from the three lakes. Woodruff Apopka Okeechobee male female male female male female r f\ COT A T1 1 -U. 15 1 A A AO -0.493 -0.564 -0.627 -0.472 T 3 r 2 0.684 0.535 0.243 0.318 0.393 0.223 P s-r\ C\(\\ s-r\ (\(\ 1 0.089 0.012 0.088 r -U.OOZ a a i -0.691 -0.122 -0.709 -0.358 -0.521 T 4 r 2 0.438 0.477 0.015 0.502 0.128 0.271 P U.UUj A AA 1 0.001 0.660 0.022 0.189 0.067 r 0.1 0.643 0.680 0.303 0.505 0.573 E 2 r 2 0.010 0.413 0.462 0.092 0.303 0.328 P 0.701 0.003 0.004 0.541 0.569 0.032 r 0.521 0.553 0.083 0.474 0.594 0.063 T r 2 0.271 0.306 0.007 0.225 0.353 0.004 P 0.032 0.014 0.752 0.166 0.025 0.821

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50 Figure 3-5. Relationship between estradiol170 concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff. A clear relationship between body size and E2 concentration was detected in Okeechobee and Woodruff females, but not in Apopka females. The significant relationship between body size and E 2 concentration in Lake Apopka males disappears if the one individual greater than 105 cm is removed from the analysis.

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51 100 80 3 60 40 20 0 Lake Anonka A O a females o males A o o o"~ CMJ*'
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52 Figure 3-6. Relationship between T concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff. Males from Woodruff and Okeechobee exhibited a relationship between body size and T concentration, but males from Apopka showed no such relationship.

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2 1 0 4 53 Lake Apopka a females o males o g^CTznjfrt ?>y op r& 9ja y -o 60 80 100 120 140 3210 Lake Okeechobee 60 80 o --6 .Aa. 100 120 140 4 3210 Lake Woodruff XT o o 5k-— * o , o o 60 80 100 Length (cm) 120 140

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54 Figure 3-7. Relationship between triiodothyronine concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff. Both males and females from Woodruff exhibited a negative relationship between T3 concentration and body size, whereas there was no apparent relationship for females from Apopka and Okeechobee.

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55 I
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56 Figure 3-8. Relationship between thyroxine concentration and body size in alligators from Lake Apopka, Lake Okeechobee, and Lake Woodruff. A clear relationship between T4 and body size existed for Woodruff males and females, but this relationship was not seen in males from Apopka or females from Okeechobee and Apopka.

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57 Lake Apopka a females — o males —r~ 60 80 100 120 140 Length (cm)

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58 Discussion The results indicate differences in concentrations of both sex steroid and thyroid hormones among the alligators of Lake Apopka, Lake Woodruff, and Lake Okeechobee. Whereas there were no differences in E 2 concentrations among animals of the three lakes, T concentrations in Lake Apopka and Lake Okeechobee male alligators were significantly lower than T concentrations in Lake Woodruff male alligators. Concentrations of T 4 also differed in animals of the 3 lakes, with T 4 concentrations being lower in Lake Woodruff males compared to Lake Okeechobee male alligators. In general, the thyroid hormones were inversely related to alligator size, whereas the steroid hormones exhibited a positive relationship to size. Deviations from this pattern were most pronounced in Lake Apopka animals, with a lack of any relationship between T 4 or T and size in males, and a lack of any clear relationship between T 3 or E 2 and size in females. Additionally, body size had no correlation with plasma E 2 in males of Lakes Okeechobee or Woodruff, whereas Apopka males showed a positive correlation. This positive correlation is attributed to one large Apopka male, and it is unknown if this pattern is common among Apopka males of a similar size. Animals of this size class are rare on Lake Apopka due to the lack of emergent vegetation. Neither T 3 nor T 4 was correlated with alligator size in Okeechobee females, but there was a clear correlation in Okeechobee males. These results indicate differences in hormone concentrations among lakes, and these differences are dependent upon both the sex and the size of the alligator. The differences noted between males and females support previous data suggesting that the specific effects of endocrine-disrupting chemicals can be gender specific.

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59 Normally, juvenile reptiles exhibit sexual dimorphism in circulating steroid hormone concentrations. Juvenile male green sea turtles (Chelonia mydas) have higher circulating testosterone concentrations compared to females; in fact, this difference is the only means of determining the sex of these immature sea turtles (Wibbels et al., 1987). Similarly, 6month-old male alligators from Lake Woodruff have significantly more T than females, and Woodruff females have significantly more circulating E2 than males (Guillette et al ., 1994). Guillette et al. (1994) found, however, that 6-month-old alligators from Lake Apopka did not conform to this pattern. Apopka males had low concentrations of T and Apopka females had higher E2 concentrations compared to animals from Lake Woodruff. These results are from alligators reared in a control environment since hatching; thus, they represent a difference in the embryonic organization of the reproductive system (Guillette et al ., 1995a). Results of the current study support these previous data, as Apopka males have a significantly lower mean concentration of T and a pattern of higher concentrations of E 2 compared to Lake Woodruff males. These data suggest that the previously reported organizational alterations in 6-month-old alligators from Lake Apopka persist through the juvenile years and, in fact, may be more pronounced in larger (> 80 cm) juvenile alligators. Depression of circulating T concentrations has been noted before in animals exposed to environmental contaminants. Adult male rats treated with 2,3,7,8-tetrachlorodibenzo-/?-dioxin exhibit reduced androgen concentrations (Moore et al., 1985), apparently due to inhibition of cholesterol mobilization early in the steroidogenic pathway (Moore et al., 1991). These results are similar to those found in white sucker (Catostomus commersoni) exposed to bleached kraft pulp mill effluent (BKME) in Lake Superior. After exposure to BKME, male white sucker exhibit decreased concentrations

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60 of T and 1 1-ketotestosterone, and females exhibit decreased concentrations of T during both viteliogenesis and spawning (McMaster et al.„ 1991; Munkittrick et al., 1991). In populations of DalPs porpoises (Phocoenoides dalli) in the North Pacific, increasing concentrations of PCBs and DDE in the blubber are correlated with decreased circulating T levels (Subramanian et al., 1987). These altered androgen concentrations in rats, white sucker fish, and porpoises are likely activational; that is, the function of a normally organized reproductive system is altered due to exposure to an endocrine-disrupting agent (Guillette et al., 1995a). For example, normal goldfish (Carassius auratus) exposed to BKME demonstrate reduced circulating T concentrations within 4 days of exposure (McMaster et al ., 1996). Most of the evidence for Lake Apopka animals, and possibly those of Okeechobee, does not suggest activational disruption, but that an organizational change has occurred (Guillette et al., 1995a). Exposure of developing embryos to hormone-disrupting contaminants can alter the normal development of endocrine organs such as gonads and thyroids. For example, several polychlorinated biphenyls (PCBs) cause gonadal sex reversal (from default male to female) in a turtle species with temperature-dependent sex determination (Bergeron et al., 1994). Similar sex reversal from male to female has been noted in embryonic alligators exposed to the natural estrogen estradiol17p (Lance and Bogart, 1992; Crain et al., 1997b), the estrogen agonist/antagonist tamoxifen (Lance and Bogart, 1991; Crain et al., 1997b), and several inhibitors of steroidogenic enzymes (Lance and Bogart, 1992). It is unknown if such sexreversed animals exhibit normal reproductive function later in life, and it is possible that the altered steroid and thyroid hormone concentrations noted in the present study are due to such organizational endocrine disruption.

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61 Although the exact causative agents of endocrine disruption in Lake Apopka alligators have not been identified, it is likely that the alterations in steroid hormones are due to embryonic exposure to a number of different compounds. As previously mentioned, Lake Apopka is adjacent to an EPA Superfund site. Due to a 1980 spill at the Tower Chemical Co. site, a pesticide mixture composed of dicofol, DDT, DDD, and DDE entered the lake (Environmental Protection Agency, 1994). Although these compounds are known endocrine disrupters, it is impossible to draw a direct cause-effect relationship between these contaminants and the hormonal disruption in Apopka' s alligators because of the historical and current extensive agricultural practices surrounding Lake Apopka. Alligator eggs collected from Lake Apopka in 1984 and 1985 had relatively high concentrations of p,p' -DDE, p,p'-DDD , toxaphene, dieldrin, and fraws-nonachlor (see Table 3) (Heinz et al., 1991). A recent study showed that two of these compounds (p,pDDD and trans-nonachlor) exhibit binding to the alligator estrogen receptor (Vonier et al., 1996). Additionally, Arnold et al. (1996b) found that the combination of dieldrin and toxaphene actively competes for the human estrogen receptor, whereas each individual compound has no affinity for the receptor. This "synergistic" activation of the estrogen receptor emphasizes the potential for mixtures of contaminants to alter the organization of the reproductive system. Circulating thyroid hormones were highly correlated to body size in male and female alligators from Lake Woodruff. Similar strong relationships between thyroid hormone concentration and body size were noted for Lake Apopka and Okeechobee males for T 3 and for Apopka females for T 4 . However, females from Apopka and Okeechobee showed little relationship between body size and T 3 , and males from Apopka and both

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62 sexes from Okeechobee showed little relationship for T 4 . In reptiles and mammals, T3 is considered the active thyroid hormone, as it binds avidly to nuclear receptors resulting in the orchestration of metabolism and growth (Eales, 1990). However, T 4 may be the most important circulating indicator of thyroid hormone status, as T 4 is the primary hormone secreted by the thyroid. At the level of the cell, T 4 is converted to T 3 by deiodonase enzyme activity (Eales, 1990). Thus, both T 4 and T 3 concentrations are critical regulators of growth and metabolism, and both thyroid hormones can be differentially produced depending upon season and sex (Gancedo et al., 1995). These thyroid hormone concentrations can be modified by exposure to endocrine disrupters. For instance, brown trout (Salmo trutta) exposed to sublethal levels of aluminum exhibit elevated circulating concentrations of both T 3 and T 4 (Waring et al., 1996; Waring and Brown, 1997). Aluminum is thought to stimulate T 4 release from the thyroid and increase hepatic monodeiodination of T 4 to T 3 , thus resulting in elevated T 3 and T 4 in the plasma (Waring and Brown, 1997). Therefore, it is plausible that the elevated T 4 concentrations in Lake Okeechobee males and the lack of correlation between body size and thyroid hormones in Lake Okeechobee and Lake Apopka animals could be attributed to exposure to endocrine disrupters. The lack of correlation between the thyroid hormones and size in both Lake Apopka and Lake Okeechobee animals could reflect altered reproductive potential in these animals, as the thyroid hormones cooperatively regulate the reproductive activities of vertebrates. For instance, thyroidectomy causes complete inhibition of spermatogenesis in the lizard Calotes versicolor and the gecko Coleonyx variegatus, and this inhibition is restored by administration of T 4 (Plowman and Lynn, 1973; Haldar-Misra and Thapliyal,

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63 1981). However, administration of T 4 to normal, non-thyroidectomized geckos causes inhibition of spermatogenesis similar to that in thyroidectomized animals (Plowman and Lynn, 1973). These results suggest that any aberration in the concentration of circulating T 4 , whether it be increased or decreased, can have a substantial effect on spermatogenesis and, thus, male reproductive success. This corroborates previous suggestions that the gonads can only function normally within a limited range of thyroid activity (Eyeson, 1970). For this reason, contaminants that interfere with the function of the thyroid can be potent disrupters of reproduction. In Lake Apopka alligators, previous studies have documented reproductive abnormalities that are symptomatic of endocrine disruption (Guillette et al., 1994; Guillette et al., 1996b), but the pathways through which environmental chemicals elicit such abnormalities have not been fully elucidated (Guillette et al., 1995a). It is possible that the thyroid/gonad axis is involved, and future studies should examine the relationship between reproductive endocrine disruption, thyroid function, and environmental contaminants. Table 3-3. Mean concentrations (ppm) of environmental contaminants measured in alligator eggs collected from Lake Apopka during 1984 and 1985. Contaminant 1984 Concentration (ppm) 1985 Concentration (ppm) Toxaphene 0.09 2.4 Dieldrin 0.24 0.11 p,p -DDE 5.8 3.5 p,p '-DDD 0.82 0.37 /ra/w-nonachlor 0.11 0.15 Source: Data from Heinz et al. (1991) with permission.

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64 In summary, this study has shown that (1) circulating testosterone is reduced in male Apopka and Okeechobee alligators relative to males from Lake Woodruff, (2) circulating thyroxine is elevated in Lake Okeechobee males compared to Lake Woodruff males, and (3) in general, body size is correlated with steroid and thyroid hormone concentrations in animals of Lake Woodruff, but this is not always the case in animals of Lake Apopka or Lake Okeechobee. Future studies of endocrine disruption in ectotherms should consider (1) the role of thyroid hormones in contaminant-induced reproductive alterations, (2) the potential of endocrine-disrupting contaminants to alter growth via thyroid hormone disruption, and (3) size-specific responses to endocrine-disrupting chemicals.

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CHAPTER 4 TESTOSTERONE SYNTHESIS IN EMBRYONIC AND JUVENILE ALLIGATORS EXPOSED TO ENDOCRINEALTERING ENVIRONMENTAL CONTAMINANTS Introduction Environmental contaminants can alter the reproduction and growth of animals by interfering with the normal functioning of the endocrine system. The endocrine regulation of reproduction appears particularly susceptible to perturbation by endocrine-disrupting contaminants (EDCs), as many EDCs affect synthesis, availability, and binding of reproductive hormones (Chapter 2). .Contaminant-induced endocrine disruption has been shown to alter the reproduction of fish (McMaster, 1995), amphibians (Palmer and Palmer, 1 995), reptiles (Guillette and Crain, 1995), birds (Fry, 1995), and mammals (Subramanian et al., 1987; Beland et al., 1993). These endocrine-altering effects do not appear to be characteristic of any particular class of environmental contaminants, as endocrine alteration can be seen after exposure to agricultural, industrial, and municipal waste compounds (Colborn et al., 1993). One of the best-described case studies of endocrine disruption in a wildlife species involves the American alligators (Alligator mississippiensis) of Lake Apopka, Florida (USA). Lake Apopka is 1 .5 miles downstream from an EPA Superfund site (where a spill of dicofol, DDT, and other compounds occurred in 1980), receives significant agricultural runoff, and was previously used as a municipal sewage reservoir (Schelske and Brezonik 1992). A study conducted in the late 1980s and early 1990s showed that both the number Note: This chapter is in review in Comparative Biochemistry and Physiology, Part C (Crain and Guillette, 1997b). 65

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66 of juvenile alligators and the clutch viability are reduced on Lake Apopka compared to other lakes in Florida (Woodward et al., 1993), and alligator eggs taken from Lake Apopka in the mid 1980s had high residues of several organochlorines (Heinz et al., 1991). Juvenile alligators living in Lake Apopka exhibit a number of morphological modifications suggestive of endocrine disruption: (a) males on Apopka have smaller penis size compared to males on other lakes (Guillette et al, 1996b), (b) males on Apopka have poorly organized testes compared to those on other lakes (Guillette et al, 1994), and (c) females on Apopka have abnormal ovaries that exhibit polyovular follicles and polynuclear oocytes (Guillette et al, 1994). In addition to these morphological abnormalities, animals from Apopka have abnormal circulating concentrations of reproductive hormones. When compared to animals from a reference site, Apopka males have decreased circulating testosterone concentrations (Chapter 3) (Guillette et al, 1994; Guillette et al, 1997) and Apopka females have increased circulating estradiol concentrations (Guillette et al, 1994). There are a number of ways that contaminants alter circulating hormone concentrations, such as (a) directly affecting gonadal steroid production, (b) indirectly altering steroid production by changing the hypothalamic-pituitary control of gonadal steroid synthesis, (c) changing the normal hepatic excretion rate of steroids, or (d) affecting the amount of circulating "free hormone" by altering sex-steroid binding protein synthesis. The purpose of this study is to test the first of these possibilities: that contaminants can directly alter gonadal steroid production. A descriptive and an experimental study are utilized to assess the effects of contaminants on gonadal testosterone synthesis. This approach is taken to further elucidate the mechanism(s) of endocrine disruption in alligators exposed to EDCs.

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67 Materials and Methods Descriptive Study Eggs were collected from Lake Woodruff National Wildlife Refuge, Florida, and Lake Apopka, Florida, under permit from the Florida Game and Freshwater Fish Commission. Eggs were incubated at 3 1°C, a temperature that produces mostly females (Lang and Andrews, 1994). After hatching, neonates were transported to the Santa Fe Teaching Zoo (Gainesville, Florida) where they were housed in an outdoor semi-aquatic enclosure. Animals were fed ad libitum daily with a commercial alligator chow (Burris Mill and Feed, Inc., Franklinton, LA). At 9 months of age, animals were sexed by examination of phallus development (Allsteadt and Lang, 1995), and the animals were transported to our lab for examination. Five males and 5 females were examined for Lake Apopka hatchlings, whereas 3 males and 7 females were examined for Lake Woodruff hatchlings. We were unable to examine more male alligators from Lake Woodruff due to a lack males in this cohort group. Experimental Study Eggs were collected from 5 nests at the reference site, Lake Woodruff, and transported to our laboratory. One egg from each nest was opened to determine the developmental stage of the embryo (staging based on criteria defined by Ferguson (1985)). Eggs from each clutch were separated equally among 5 groups (15 eggs per group), and each group was exposed to a particular treatment at developmental stage 21, just prior to the onset of sexual differentiation (Lang and Andrews, 1994). Table 1 summarizes the experimental design. Two of the treatment groups were controls (one group incubated at

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68 a male-producing temperature 33°C, and one group incubated at a female-producing temperature 30°C), one group was an endocrine-disrupting standard (0.1 ppm estradiol1 7p), and two groups were exposed to common environmental contaminants (5 ppm p,p 'DDE or 5 ppm p,p '-DDD, both metabolites of DDT). Except for the control females, all eggs were incubated at 33°C, a temperature that normally produces 100% males (Chapter 4, Chapter 5) (Lang and Andrews, 1994). Five embryos from each treatment group were sampled at each of the following stages: stage 23 (in the middle of gonadal differentiation), stage 25 (at the end of gonadal differentiation), and hatching (approximately 2 weeks after gonadal differentiation). Table 4-1. Design for the experimental dosing of alligator eggs. Eggs were assigned to an experimental group, exposed to the treatment, and sampled at either developmental stage 23, developmental stage 25, or hatching. Five animals were sampled for each treatment group at each developmental stage. Stage 23 Stage 25 Hatching Control Male 5 5 5 Control Female 5 5 5 p,p '-DDD (5 ppm) 5 5 5 p,p '-DDE (5 ppm) 5 5 5 Estradiol17p (0.1 ppm) 5 5 5 Tissue Culture A lethal injection of sodium pentobarbital (0.4 mg/g) was administered into the dorsal postcranial sinus of hatchlings and 9-month olds and into the vitellein vein of the

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69 embryos. The gonad/adrenal/mesonephric complex (hereafter referred to as gonad) was removed and weighed. Tissue culturing procedures were modified from McMaster et al. (1995a). For the descriptive study, the left gonad (termed "stimulated gonad") was placed in 700 ul media (Medium 199 with Hanks salts, L-glut, and 25 mM Hepes; Gibco BRI, Gaithersburg, MD) supplemented with 5 uM forskolin as a cAMP stimulator, 0. 1 mM 3isobutyl 1-methylxanthine (IBMX) as a cAMP protectant, and androstenedione as a steroid precursor. The right gonad (termed "unstimulated gonad") was placed in 700 u.1 media supplemented with 0. 1 mM IBMX. The gonads were incubated at 32°C for 5 hrs; the media was removed and flash frozen in liquid nitrogen, and fresh media (either supplemented media for the left gonad or non-supplemented media for the right gonad) was added. After 5 more hours, the media was again collected and changed. This last aliquot of media was collected after 15 hrs of incubation. After flash freezing, the media was stored at -72°C prior to analysis for steroid hormones. For the experimental study, the right gonad was removed for histology and the left gonad was placed in 500 u.1 media supplemented with forskolin, IBMX, and androstenedione as described above. After 5 hours of incubation, the media was collected, frozen in liquid nitrogen, and stored at -72°C. Histological Analysis Gonads from the experimental animals were examined by histology to document which compounds induced the production of females at a male-producing temperature and to observe tissue-level changes induced by the compounds. The gonad was preserved in

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70 Bouin's fixative, serial sectioned at 7 um following paraffin embedding, and stained with a modification of Harris' trichrome staining procedure (Humason, 1972). Gonads were inspected and scored as testis or ovary by two independent researchers. Histological criteria originally reported by Forbes (Forbes, 1 940) and recently reestablished by Guillette et al. (Guillette et al., 1994) were used to determine sex. In brief, criteria for testes included reduced cortex and medullary sex cord proliferation, whereas criteria for ovaries included hypertrophied cortex, medullary reduction, the presence of lacunae in the medulla, and germ cells in the cortex. Radioimmunoassays T concentrations were determined using a radioimmunoassay (RIA) previously validated for alligator plasma (Chapter 3(Guillette et al., 1997). Briefly, 70 u.1 of culture media was extracted 2x with 5 ml diethyl ether. The ether extract was air dried and the following were added to the dried assay tubes: Assay buffer (100 ul; 0.05M borate buffer, pH=8.0), assay buffer with bovine serum albumen (100 u,l; BSA at a final assay concentration of 0.15%), antibody (200 ul; final concentration of 1:25,000; Endocrine Sciences, Calabasas Hills, CA), and radiolabeled tracer (100 ul; 12,000 cpm per 100 ul; 1 mCi/ml; Amersham International, Bukinghamshire, England). Samples were compared to reference standards prepared at 0, 1.56, 3.13, 6.25, 12.5, 25, 50, 100, 200, 400, and 800 pg/tube. All samples and standards were prepared in duplicate. Tubes were incubated overnight at 4°C. Bound-free separation was accomplished by adding 500 u.1 5% charcoal 0.5% dextran and immediately centrifuging at 1500 g at 4°C for 30 min. Supernatant (500 ul) then was added to 5 ml Scintiverse BD (Fisher Chemical Company) and the tubes

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71 were counted on a Beckman scintillation counter. All samples were assayed in duplicate in a single assay, and the intraassay variability averaged 6.78%. The T RIA was validated for the culture media by comparing media dilutions to the T standard curve. A pool was made from an aliquot of each sample, and the following volumes of this pool were aliquanted: 0, 10, 25, 40, 55, and 70 ul All tubes were brought to 70 u.1 with fresh, uncultured media. These samples were extracted and assayed as described above. Figure 1 presents the results of the RIA validation. Statistical analysis (test for homogeneity of slopes; Super Anova; Abacus Concepts, Inc., Berkeley, CA) revealed that there was no significant difference (p=0.361) between the slopes of the culture media dilutions and the standard curve. We attempted to measure estradiol17p (E 2 ) in the culture media using similar extraction and assay techniques to those in the T RIA. While this E 2 RIA is successful in measuring plasma concentrations of E 2 down to 12 pg/ml (Chapter 3) (Guillette et al., 1997), there was not sufficient E 2 in the culture media to produce a validation for the RIA. Therefore, results for gonadal production of E 2 are not presented. Statistical Analysis Hormone concentrations were determined using a commercially available software package (Microplate manager III; Biorad, Hercules, CA). A comparison of T production between Apopka and Woodruff animals was conducted using a repeated measures ANOVA (SuperAnova; Abacus Concepts, Inc., Berkeley, CA). For the results of the Experimental Study, a 2way ANOVA was used to evaluate the effects of treatment and developmental stage on T production. To obtain homogeneity of variances, T

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72 concentrations were log transformed prior to this analysis. Fisher's Protected LSD was used as a post hoc test, and p=0.05 was the accepted level of significance. 100 1 10 100 1000 Testosterone (pg) Figure 4-1 . Radioimmunoassay validation for the measurement of testosterone in the culture media. The statistical test for homogeneity of slopes revealed that the media dilution curve was not significantly different from the standard curve (p=0.361). Results Descriptive Study Figure 4-2 presents the results of the descriptive study. Exposure to androstenedione, forskolin, and IBMX ("stimulated gonads") stimulated significantly greater T secretion from ovaries (p<0.0001) and testes (p<0.0001) compared to unstimulated gonads. The interaction of lake and time of incubation was not significant for either testes or ovaries. Time of incubation had a significant influence on ovarian T

PAGE 80

73 production in unstimulated (p=0.003) and stimulated (p=0.002) ovaries. However, time of incubation had no effect on testicular T production in either unstimulated (p=0.209) or stimulated (p=0.672). There was no difference in T production between Apopka and Woodruff animals for either stimulated ovaries (p=0.733), unstimulated ovaries (p=0.936), stimulated testes (p=0.395), or unstimulated testes (p=0.799). Experimental Study We were unable to differentiate between testes and ovaries in stage 23 and stage 25 embryonic alligators. However, clear differences were seen in the hatchling alligators. Histological examination of the hatchling gonads revealed that female alligators were produced at a male-producing temperature in the estradiol1 70 (5 of 5) and p,p '-DDD (2 of 5) treatment groups (Figure 4-3). The remaining treatment groups produced the sex defined by the temperature regime. Gonads from the p,p '-DDD-treated male hatchlings appear different from control males (Figure 4-4). Testes from animals exposed to 5 ppm p,p -DDD appear developmental^ accelerated, as there is a clearly defined lumen in the seminiferous tubules. Additionally, these testes have poorly organized seminiferous tubules when compared to testes from control males. Exposure to 5 ppm p,p '-DDE produced no observable abnormalities in the hatchling testes. Figure 4-5 presents mean T concentrations of the alligators treated in ovo. The interaction of treatment group and embryonic stage was not significant (p=0.067), indicating that there is no apparent stage-specific treatment effect. However, both developmental stage (p=0.001) and treatment group (p=0.050) did influence T concentrations. Among stages, stage 23 animals produced more T than stage 25 animals

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74 (p=0.017) and hatchlings (p^O.002). Among treatment groups, p,p '-DDD treated animals produced significantly more T than control females (p=0.004) and E 2 -treated animals (p=0.038). Discussion This study detected no difference in testosterone (T) production between juvenile alligators from Lakes Apopka and Woodruff. However, Lake Woodruff alligators treated in ovo with p,p '-DDD (one of the compounds found in Lake Apopka eggs) had significantly higher in vitro gonadal T production compared to control female alligators and E 2 -treated alligators. Therefore, it appears that exposure to p,p '-DDD is able to alter gonadal steroid production, but this alteration does not explain the hormonal or morphological alterations noted in Lake Apopka alligators. Several studies have documented reduced circulating T concentrations in juvenile male alligators from Lake Apopka (Chapter 3(Guillette et al., 1997). This reduction appears to be defined in ovo, as 6-month-old Apopka males reared in the lab show reduced plasma T concentrations compared to lab-reared Lake Woodruff males (Guillette et al., 1994). The mechanisms of T reduction in the Apopka males are unknown, but a previous study found no difference in the production of T in ovaries or testes from 6month-old Apopka and Woodruff alligators (Guillette et al., 1995b). These results are consistent with those of this more detailed study that found no difference in gonadal T production between 9-month-old Apopka and Woodruff alligators. If the Apopka animals produced decreased T as a result of direct gonadal suppression of steroidogenesis, then the cAMP-stimulated gonads from Apopka animals would show reduced T synthesis

PAGE 82

75 (jq/6/6u) auoj3)so)sai 3 O £_ E (jq/6/6u) auojajsojsai (jq/B/Bu) auoj3)so)S9i (jq/6/Bu) auojaisoisai 1 CX, -o O CO Q, C
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Figure 4-3. Photomicrographs of hatchling alligator ovaries from a control female (A, incubated at 30°C) and a alligator that was incubated at a male-producing temperature (33°C) and treated with p,p'-DDD (B). Numerous germ cells are present in the cortex (C) of both animals, as well as a clearly defined medullary region (M). Both gonads appear to be normal ovaries, even though the p,p'-DDD-treated animal was incubated at a temperature that produces testes. Magnification 25x.

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77

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Figure 4-4. Photomicrographs of hatchling alligator testes from a control male (A) and a p,p'-DDD treated male (B). Tissue from the control males had well-organized seminiferous tubules with little or no lumen. In contrast, testes from hatchlings treated in ovo with p,p'-DDD have poorly organized testes and pronounced lumen. Magnification lOOx.

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79

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80 30 25 O) O) 20 0) § 15 0) go 10 o -4— » h5 Control Female 30°C Control Male 1 Stage 23 Stage 25 Hatching ft | p.p'-DDD p.p'-DDE Estradiol 33°C Figure 4-5. Testosterone production (±1 SE) of gonads from embryonic and hatchling alligators in the experimental treatment groups. Control females were incubated at 30°C (female-producing temperature), whereas all other treatment groups were incubated at 33°C (male-producing temperature). Gonads from alligators treated with p,p'-DDD produced significantly more T compared to control females and E 2 -treated animals. Although p,p'-DDD induced ovarian development in 2 of the 5 treated alligators, analysis revealed that the elevated T concentrations in the p,p'-DDD treatment group can be attributed solely to T production from the testes (p: testicular T production = 77.3 ng/g/hr; [i ovarian T production = 6.59 ng/g/hr).

PAGE 88

81 compared to cAMP-stimulated Woodruff gonads. This is not the case and, therefore, the reduced circulating T concentrations previously reported for juvenile Apopka alligators can not be attributed to a direct reduction in gonadal T synthesis. Alternative sites for the androgen reductions in the Apopka males include (1) increased conversion of T to estradiol170 by the enzyme aromatase, (2) increased excretion of T, (3) decreased availability of free T as a result of increased sex-steroid binding protein production, (4) decreased gonadotropin synthesis from the anterior pituitary. Future studies should explore the effects of endocrine altering contaminants on these potential sites of disruption. The experimental egg dosing study found that both developmental stage and treatment group had an influence on gonadal T synthesis. T synthesis is greatest during the period of gonadal differentiation (Stage 23 embryos). This period coincides with the window in development when temperature influences the determination of sex in alligators (Lang and Andrews, 1994). We were unable to detect significant concentrations of estradiol170 (E 2 ) in the culture media with our E 2 radioimmunoassay, but a previous study found that female alligators progressively produce more E 2 from stage 23 to hatching (Smith et al., 1995). Conversely, we found that both male and female alligators produce progressively less T from stage 23 to hatching. Therefore, it appears that T is produced in both males and female embryos at the onset of gonadal differentiation (Stage 23) and, in females, aromatase enzyme activity is elevated as development proceeds from Stage 23 to hatching. This causes increased E 2 in females but not males (Smith et al., 1995).

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82 Hatchling alligators that were exposed in ovo to 5 ppm p,p '-DDD exhibited increased T production compared to control females and E 2 -treated hatchlings. In the hatchlings that were exposed in ovo to p,p '-DDD, consideration of both the histology and the T concentrations revealed that the elevated T production can be attributed to only those hatchlings that were males. Further, the testes of these p,p '-DDD treated males appear to be developmentally advanced compared to control males, as the sex cords have clearly defined lumen. Therefore, p,p '-DDD can override the effects of temperature by producing female alligators at a male-producing temperature, but also can induce testicular changes in alligators that are not "sex reversed." These effects are likely mediated through the estrogen receptor, as/>,/;'-DDD binds to the alligator estrogen receptor (Vonier et al, 1996). Future studies should determine if these testicular abnormalities are prevalent and persistent among males that are exposed to p,p'-DDD as embryos. Unlike those exposed in ovo to p,p '-DDD, hatchlings that were exposed in ovo to 5 ppm/?,/> '-DDE exhibited no structural or functional gonadal alterations. All p,p '-DDE exposed animals had apparently normal testes and produced normal concentrations of testosterone. This lack of effect may be due to the dynamics ofp,p '-DDE with steroid receptors. Unlike p,p '-DDD, p,p '-DDE exhibits weak affinity to the alligator estrogen receptor (Vonier et al ., 1996). In rodents, p,p '-DDE is an antagonist of the androgen receptor both in vitro (Kelce et al., 1995) and in vivo (Kelce et al., 1997), although p,p 'DDE may interact with the androgen receptor at a dose approximately 10 6 -fold higher than endogenous testosterone (Gaido et al., 1997). Therefore, it is possible that higher dosages ofp.p '-DDE could cause testicular abnormalities through anti-androgenic mechanisms.

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83 Although the present study found no difference in testicular T production between Lake Apopka and Lake Woodruff alligators, it is possible that such a difference would occur or become evident in older alligators. Environmental contaminants can differentially affect plasma androgen concentrations depending on the animal's developmental stage, sex, and reproductive status. The effects of bleached kraft mill effluent (BKME) on androgen concentrations in white sucker fish (Catostomus commersoni) best illustrate this. Gagnon et al. (1994a) found that BKME exposure causes no change in T concentrations in males, but induces increased T concentrations in females. However, a more detailed study of male white suckers found that while suckers sampled in August were not influenced BKME exposure, BKME caused increased T concentrations in males sampled in September (Munkittrick et al., 1992b). Conversely, BKME-exposed male fish had reduced 1 1-ketotestosterone in August but not in September (Munkittrick et al., 1992b). Although the mechanisms for increased circulating androgens remain unknown, it has been suggested that the decreased androgen concentrations are due to BKME-induced reduction in steroid biosynthesis (McMaster et al., 1991; McMaster et al., 1995b). This study has shown that in ovo exposure to p,p -DDD can cause structural and functional changes in the gonads of hatchling alligators. In some individuals, exposure to 5 ppm/?,/? -DDD caused the production of females in prospective males. In other individuals exposed to p,p -DDD, sex reversal was not induced, but testes appeared to be developmentally accelerated and poorly organized. These testes also produced significantly more T compared to ovaries, whereas testes from control animals did not. Although alligators from Lake Apopka are exposed to similar concentrations ofp,p '-DDD (up to 1.8 ppm (Heinz et al, 1991)) during embryonic development, nine-month-old

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84 alligators from Lake Apopka do not show such abnormalities. No difference was seen in T synthesis between alligators from Lake Apopka and Lake Woodruff. Future studies should more fully characterize both the endocrine-altering effects of p,p '-DDD and the mechanisms of endocrine disruption in Lake Apopka alligators.

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CHAPTER 5 ALTERATIONS IN STEROIDOGENESIS IN ALLIGATORS {ALLIGATOR MISSISSIPPIENSIS) EXPOSED NATURALLY AND EXPERIMENTALLY TO ENVIRONMENTAL CONTAMINANTS Introduction Environmental contaminants alter the reproduction of a number of wildlife species by changing the normal endocrine environment that mediates sexual differentiation and function (Chapter 2). Many of these endocrine alterations are thought to occur by direct interactions between the contaminants and hormone receptors (McLachlan, 1993), but the specific mechanisms by which most environmental contaminants cause endocrine disruption are unknown. Although all vertebrates are potentially susceptible to reproductive disruption by endocrine-disrupting contaminants (EDCs), many ectothermic vertebrates are particularly sensitive due to the processes mediating the organization of the reproductive system (Guillette et al., 1995a). Unlike birds and mammals, many fish, amphibians, and reptiles exhibit environmental sex determination, by which the gender of the undifferentiated embryo is determined by an environmental variable. In many reptiles, the temperature of egg incubation determines the sex of the offspring (Bull, 1980). Exposure of developing reptile embryos to exogenous chemicals can mimic the effects of temperature on sex determination. For example, when red-eared turtle (Trachemys scripta) embryos are incubated at a male-producing temperature and exposed to estradiol17P during the window of developmental sex determination, phenotypically female turtles Note: This chapter is published in Environmental Health Perspectives (Crain et al 1997b) 85

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86 are produced (Wibbels et al., 1991; Wibbels et al, 1993). This estrogen-induced sex reversal appears to be dose dependent (Crews et al., 1991), and suggests that other steroidal agonists, steroidal antagonists, and steroidogenic disrupters could alter normal sexual differentiation. Indeed, Wibbels and Crews (1992) found that steroid hormones are not exclusive in their ability to alter normal sex determination, as many estrogen agonists and steroidogenic modifiers mimic and/or reverse the effects of temperature on the differentiation of primary sex organs in red-eared turtles. The specific mechanisms by which temperature determines gender are unknown, but it is hypothesized that temperature stimulates or suppresses pivotal steroidogenic enzymes (Wibbels et al., 1994). These enzymes then propagate a cascade of events leading to the organization of a testis or ovary. This hypothesis is supported by work conducted on the steroidogenic enzyme aromatase. Aromatase converts androgens to estrogens by binding the Ci 9 androgen substrate and catalyzing several reactions leading to a phenolic ring characteristic of estrogens (Simpson et al., 1994). Several lines of evidence support the pivotal role of aromatase in temperature-dependent sex determination. First, several studies indicate that aromatase activity is increased in prospective females during periods coinciding with thermosensitivity (Desvages and Pieau, 1992; Chardard et al., 1995; Smith et al., 1995). Second, high doses (50-100 ug per egg) of testosterone cause feminization of 71 scripta at a male-producing temperature (Wibbels and Crews, 1992; Crews et al., 1995b). Because testosterone is the precursor to estradiol170, this phenomenon is thought to be mediated by the enzyme aromatase. Third, administration of an aromatase inhibitor induces male sex determination in both a female unisexual (parthenogenetic) lizard and a turtle with temperature-dependent sex

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87 determination (Wibbels and Crews, 1994). Collectively, these studies suggest that aromatase is an enzyme critical to thermosensitive sex determination and is capable of being modified by extrinsic factors. In consideration of these studies, I propose that the endocrine-altering effects of some environmental contaminants may be mediated via changes in the expression or activity of the aromatase enzyme. Two studies, one descriptive and one experimental, were conducted to test this hypothesis. First, juvenile alligators from a control lake and a lake historically contaminated with a number of persistent organochlorines were analyzed for plasma steroid hormones and in vitro gonadal-adrenal aromatase activity. Second, embryos from a control lake were exposed to several known hormonal modifiers and two common herbicides, and hatchlings were analyzed for (a) egg chorioallantoic fluid hormones, (b) plasma steroid hormones, and (c) in vitro aromatase activity. Using these studies, I sought to determine whether aromatase function could explain endocrine alterations in alligators exposed to EDCs. Materials and Methods Animals and Treatments For the descriptive study, eggs were collected from 6 nests on Lake Apopka, Florida (contaminated lake) and 6 nests on Lake Woodruff National Wildlife Refuge, Florida (control lake) during the first week of July 1995. Lake Apopka is designated as one of Florida's most polluted lakes (Schelske and Brezonik, 1992) due to extensive agricultural activities around the lake, a sewage treatment facility associated with the city of Winter Garden, FL, and a major pesticide spill from the Tower Chemical Company.

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88 The pesticide spill, which occurred in 1980, consisted primarily of dicofol but had significant amounts of DDT, DDE, and DDD in the mixture (U.S. EPA, 1994). Analysis of alligator eggs taken from Lake Apopka in 1984 and 1985 revealed significant residues of toxaphene, dieldrin, p,p -DDE, p,p -DDD, /raAw-nonachlor, and PCBs (Heinz et al., 1991). A previous study found evidence of "estrogenic contamination" among the female alligators of Lake Apopka (Guillette et al ., 1994) and, as I wanted to minimize the number of eggs taken from Lake Apopka, eggs were incubated only at a female-producing temperature (30°C) (Lang and Andrews, 1994). After hatching, alligators were housed at Sante Fe Teaching Zoo (Santa Fe Community College, Gainesville, FL) in outdoor, semiaquatic enclosures. At nine months of age, the female alligators were transported to the laboratory for collection of tissues. For the experimental study, eggs were collected from 7 nests on Lake Woodruff, Florida during the first week of July 1995. Eggs were transported to the lab, placed in an incubator at 30°C, and one egg from each clutch was opened to stage the embryos. Staging was based on criteria defined by Ferguson (1985). Five days after collection (and prior to the temperature sensitive period when sex determination occurs), eggs were separated into two groups such that half of the eggs from one clutch were incubated at 30°C (female producing) and half at 33°C (male producing). Eggs were maintained at approximately 90% humidity using sphagnum moss as incubation material. Within each incubation group, eggs from each clutch were distributed among 6 treatment groups of varying dosages (see Table 5-1). One treatment group served as a control and three groups served as endocrine-disrupting standards: estradiol17p, tamoxifen which acts as an estrogen in embryonic alligators but as an antiestrogen in hatchlings (Lance and Bogart,

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89 1991), and vinclozolin which is a potent antiandrogen in rodents (Gray et al., 1993b). The two remaining treatment groups were the modern-use herbicides atrazine and 2,4dichlorphenoxyacetic acid (2,4 D). Treatments were applied topically to the eggshell in 50 ul of 95% ethanol, a technique frequently used to transport compounds inside reptilian eggshells (Crews et al., 1991; Wibbels and Crews, 1992). Using this method, Crews et al. (1991) found that greater than 89% of the applied compound is incorporated into the embryo. The treatments were applied at stage 21 of embryonic development, the beginning of the critical period of gonadal differentiation (Lang and Andrews, 1994). Table 5-1. Experimental treatments and dosages (in parts-per-million ppm) that were applied to different groups of eggs. A sample size of 5 eggs was included in each dosetreatment group. Treatments Effect Doses Control" None Nothing; Diluent only Estradiol Natural Estrogen 0.014, 0.14, 1.4, 14 ppm Vinclozolin Androgen antagonist in rodents 0.14, 1.4, 14 ppm Tamoxifen Estrogen agonist/antagonist 0.14, 1.4, 14 ppm 2,4-D ?, Common herbicide 0.14, 1.4, 14 ppm Atrazine ?, Common herbicide 0.14, 1.4, 14 ppm a Each chemical was solubilized in 95% ethanol prior to topical application on the egg and, thus, two control doses were used — one with 95% ethanol and one without. There was no difference between these controls for any of the variables measured. Upon pipping, chorioallantoic fluid was collected and frozen at -72°C until steroid hormone analysis. Total protein content in the CAF samples was determined using a commercially available Bradford assay kit (Biorad, Hercules, CA), and CAF steroid

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90 hormone concentrations are presented as per ug protein. This was necessary due to differential hydration states of the CAF samples. Aromatase Assay Hatchlings were individually housed for 10 days prior to tissue collection. Following the collection of blood from the dorsal post-cranial sinus, a lethal injection of sodium pentobarbital (0.4 mg/g) was administered in the sinus. Animals are anesthetized within thirty seconds using this method. The right gonadal-adrenal-mesonephros (GAM) complex was immediately removed for the aromatase bioassay. Aromatase activity was measured indirectly based on the release of tritium from ip3 H-androstenedione during aromatization of the substrate into estrogen (Smith and Joss, 1994). Briefly, the tissue was placed in 400 ul culture media (RPMI-1640; Sigma Chemical Co., St. Louis, MO) supplemented with 0.8 mM tritiated androstenedione (DuPont NEN Research Products; # NET-926). After a 6 hr incubation at 32°C, 300 ul of the media was transferred to a new tube. Chloroform (1.5 ml) was added, the tube was vortexed, and centrifuged for 15 min at 2000 g. A 200 uJ aliquant of the aqueous phase was added to a new tube. 5% charcoal 0.5% dextran (200 u.1) was added, the tube was vortexed and then immediately centrifuged for 15 min at 2000 g. Scintiverse BD (5 ml) was added to 300 ul supernatant and the tube was counted on a Beckman Scintillation counter. Aromatase activity is proportional to the amount of tritium in the scintillation vial and is calculated as a percentage of the total substrate added. After subtracting the nonspecific tritium release, the DPM of the sample tubes are converted to a percentage of the total DPM added. This percentage is multiplied by the mass of the substrate added. After adjusting for extraction

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91 loss, the value obtained represents the amount of substrate converted to tritiated water, which is proportional to aromatase activity. Assay sensitivity was defined as twice the mean cpm of blank tubes, which corresponds to 0. 15 pmol/g/hr for the average weight GAM (0.032g). GAMs from three additional control female alligators were used to determine the specificity of the aromatase assay. The left GAM was incubated as above, while the right GAM was exposed to media supplemented with the aromatase inhibitor 4-hydroxy androstenedione (100 uM). Alligators exposed to the aromatase inhibitor had significantly lower GAM aromatase activity (u.=0.45 pmol/g/hr) compared to the individuals incubated normally (u,=3.15 pmol/g/hr). Histology Histology was conducted to determine histological sex in order to document which compounds induced sex reversal. A complete histopathological examination of the GAMs was beyond the scope of this study. The left GAM was preserved in Bouin's fixative, serial sectioned at 7 urn following paraffin embedding, and stained with a modification of Harris' trichrome staining procedure (Humason, 1972). Gonads were inspected and scored as testis or ovary by two independent researchers. Histological criteria originally reported by Forbes (1940) and recently reestablished by Guillette et al. (1994) were used to determine sex. In brief, criteria for testes included reduced cortex and medullary sex cord proliferation, whereas criteria for ovaries included hypertrophied cortex, medullary reduction, the presence of lacunae in the medulla, and germ cells in the cortex.

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92 Radioimmunoassays Estradiol17(3 (E 2 ) and testosterone (T) concentrations were measured in plasma of the 9-month-old descriptive study animals and in plasma and chorioallantoic fluids (CAF) of all hatchlings that provided ample fluids. Radioimmunoassays for E 2 and T were performed as previously described (Folmar et al., 1996) with the following modifications in sample extraction. CAF (750 p.1) was mixed overnight (15 hrs) with 2 ml of 95% ethanol. The suspension was centrifuged at 1200 g for 20 min. Supernatant (500 |il) was pipetted in duplicate for each sample and dried under constant air stream. Extraction efficiency averaged 92% for T and 94% for E 2 with this method. For plasma extraction, plasma (125 ul) was mixed with 4 ml ethyl ether for 1 min. The aqueous layer was frozen in a dry ice-methanol bath (-25°C), and the ether phase decanted into an assay tube. The aqueous pellet was reextracted with ether, and the ether added to the assay tube. The ether was dried with constant air stream. Extraction efficiency was consistent and averaged 95% for T and 94% for E 2 . Crossreactivities of the T antisera (T3-125, Endocrine Sciences, Calabasas Hills, CA) to other ligands are as follows: dihydrotestosterone, 44%; A1 -testosterone, 41%; A-l-dihydrotestosterone, 18%; 5 aandrostan-3p,17p diol, 3%; 4-androsten-3p,17P-diol, 2.5%; A-4-androstenedione, 2%; 5P-androstan-3p,17P-diol, 1.5%; estradiol, 0.5%; all other ligands <0.2%. Crossreactivities of the E 2 antisera (E26-47, Endocrine Sciences, Calabasas Hills, CA) to other ligands are as follows: estrone, 1.3%; estriol, 0.6%; 16-keto-estriol, 0.2%; all other ligands <0.2%. For the plasma RIAs, inter-assay variance was 15.0% for T and 12.6% for E 2 , and intra-assay variance was 3.6% for T and 3.7% for E 2 . For the CAF RIAs inter-

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93 assay variance was 1 1 .8% for T and 16. 1% for E 2) and intra-assay variance was 4.68% for T and 3.5% for E 2 . Statistics Hormone concentrations were estimated from raw data with the commercially available Beckman EIA/RIA ImmunoFit tm software program (Fullerton, CA). Statistics were performed with the software packages Stat View (Abacus Concepts, Inc., Berkeley, CA, 1992) and SuperAnova (Abacus Concepts, Inc., Berkeley, CA, 1989). For the descriptive study, a t-test was used for between-lake comparisons. In the experimental study, a two-factor ANOVA was used to test the effects of treatment and dose on aromatase activities. Where the interaction of treatment and dose was not significant, the factor of dose was removed from the analysis and a one-factor ANOVA was used to test the effects of treatment on hormone concentrations and aromatase activity. Fisher's protected LSD was used as a post-hoc test to discriminate which groups differed significantly. Results Descriptive Study Female Juvenile Alligators Results of the aromatase enzyme assay are expressed both as fmol/hr and pmol/g/hr. The former is used in other studies of alligator gonadal aromatase activity (Smith and Joss, 1994; Smith et al., 1995) and is presented here for comparative purposes only. A comparison of the female juvenile alligators found that gonadal-adrenalmesonephros (GAM) aromatase activity was significantly elevated in Lake Woodruff alligators compared to Lake Apopka alligators (see Table 5-2 and Figure 5-1). Mean

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94 concentration of E 2 were not different between lakes (p=0.5178), but concentrations of T were significantly lower in Apopka animals compared to Woodruff animals (p=0.05; see Figure 5-1). Table 5-2. T-test results from GAM aromatase activities of alligators from Lake Apopka (contaminated) and Lake Woodruff (control). n fmol/hr(± 1SE)* pmol/g/hr (± 1SE) Lake Apopka 12 204.5 ±21.792 7.549 ± 1.264 Lake Woodruff 14 364.4 + 47.82 11.4241 1.159 df 24 24 tvalue -2.877 -2.262 p-value 0.0083 0.0331 a Aromatase activity is expressed as both fmol/hr and pmol/g/hr; in the discussion, we refer to the more descriptive expression of pmol/g/hr, but fmol/hr is presented only to facilitate comparison with other studies (i.e., (Smith and Joss, 1994; Smith et al., 1995)). Experimental Study-Neonatal Alligators Four response variables were measured for the alligators treated in ovo. sex reversal, chorioallantoic fluid (CAF) hormone concentrations, hatchling plasma hormone concentrations, and GAM aromatase activity. Tables 5-3 and 5-4 summarize these results. One-hundred percent sex reversal male to female was noted for all dosages of E 2 and tamoxifen. No other treatments caused sex reversal. Two-way ANOVA revealed that dose had no influence on plasma hormone concentration or gonadal aromatase activity in the eggs incubated at male and female

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95 temperatures. Dose did have an influence on CAF E 2 concentrations, but this was due only to the E 2 treatment group. Therefore, dose was removed from the statistical analysis, and a one-way ANOVA was used to determine if differences existed among treatment groups. Treatment group had no significant influence on CAF hormone concentrations, with the exception of tamoxifen treatments on eggs incubated at a male-producing temperature. These "sex-reversed" females had significantly more CAF T compared to control males (p=0.024). Neither CAF T concentrations (p=0.92) nor CAF E 2 concentrations (p=0.77) were different between control males and control females. Plasma E 2 concentrations were not significantly different among treatment groups, but plasma T was different among treatment groups incubated at a male-producing temperature. Tamoxifen (p=0.028) and E 2 (p=0.027) treated animals (which were sex reversed) had elevated plasma T concentrations compared to control males. The ratio of E/T was not significantly different for eggs incubated at male or female temperatures. Neither plasma E 2 (p=0.72) nor plasma T (0.09) were significantly different between control males and control females. No differences in aromatase activity were detected among treated eggs at the female-producing temperature, but treatment groups in the male-temperature regime were significantly different. Tamoxifen-treated eggs had significantly more aromatase activity compared to control males (p=0.012). Among the other eggs incubated at a male temperature, both E 2 and atrazine-treated hatchlings appeared to have elevated aromatase activity. Although not significantly different from control males (E 2 , p=0. 15; atrazine, p=0.65), further analysis revealed that these E 2 and atrazine-treated animals were also not significantly different from control females (E 2 , p=0.65; atrazine, p=0.13).

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96 Apopka WxxJnff Figure 5-1. Steroid hormone concentrations (± 1 SE) and GAM aromatase activity (± 1 SE) for 9-month-old female alligators from Lake Apopka and Lake Woodruff. Testosterone concentrations are significantly lower in animals from Lake Apopka (p=0.05), whereas aromatase activity is significantly elevated in the alligators from Lake Woodruff (p=0.03).

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99 Discussion There are numerous mechanisms through which environmental contaminants potentially can cause endocrine alterations (Chapter 1, Chapter 2), and this study has documented that one such mechanism is the alteration of the steroidogenic enzyme aromatase. Results from the descriptive study indicate that gonadal-adrenal-mesonephros (GAM) aromatase activity is significantly different between control and contaminated juvenile alligators, but this difference does not correspond with the alterations in circulating hormones in these animals. A previous study has shown that juvenile female alligators from Lake Apopka have significantly higher concentrations of plasma estradiol170 (E 2 ) when compared to juvenile females from Lake Woodruff (Guillette et al., 1994). One hypothesis for this increase in plasma E 2 is that aromatase activity is increased in the contaminant-exposed Lake Apopka alligators. However, the current study shows that female alligators from Lake Apopka have a significantly lower mean aromatase activity when compared to females from Woodruff. These results are consistent with data from in vitro cultures of gonads from Apopka and Woodruff juvenile alligators which show that ovaries from Lake Apopka animals produce significantly less E 2 in vitro than do Lake Woodruff animals (Guillette et al., 1995b). Although this study did not detect a difference in plasma E 2 between Apopka and Woodruff female alligators, significantly less testosterone (T) was present in the plasma of the female Apopka alligators. Several studies have noted a decreased plasma T concentration in male and female juvenile alligators from Lake Apopka (Guillette et al., 1994; Guillette et al., 1996b). The extent to which such decreases in plasma T alter the

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100 physiology of alligators is unknown, but a recent study has shown that mean phallic size in alligators from Lake Apopka is significantly smaller than that of males from Lake Woodruff (Guillette et al., 1996b). Because alligator phallic development and size are androgen dependent (Pickford, 1995), it is probable that the decrease in plasma T concentrations contributes to an inhibition of penis growth in males. The effects of decreased T on females are unknown. Results from the experimental study indicate that treatment with exogenous chemicals can alter the endocrine system of developing alligator embryos. Among the endocrine-disrupting standards, E 2 and tamoxifen treatments caused the development of ovaries in embryos incubated at a male-producing temperature. However, neither atrazine nor 2,4 D had such obvious endocrine-altering effects. In an attempt to assess the impact of the treatments on the developing embryo, we measured hormone concentrations in the chorioallantoic fluid (CAF). The urinary wastes of developing oviparous embryos are stored in a membrane-bound sac called the allantois. As development proceeds, the allantoic membrane fuses with the chorion, and the fluid inside is termed chorioallantoic fluid. A previous study measured sex-steroids in this fluid and indicated that the measurement of these steroids could provide an assessment of the embryonic hormonal environment (Gross et al., 1995). Unlike the study by Gross et al. (1995), we were unable to detect any difference in steroid hormones between control males and females. The only difference in CAF hormones was for T concentrations between the tamoxifen treatment and controls at a male-producing temperature. Therefore, measurement of steroid hormones in CAF fluid does not appear to be a useful technique for assessing the embryonic hormonal environment of alligators.

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101 Measurement of aromatase activity indicates that the enzymatic activity of GAM tissue from hatchlings incubated at a male-producing temperature can be altered by exposure to exogenous compounds. None of the treatment groups had an effect on aromatase activity of hatchlings incubated at a female-producing temperature, but several differences were noted among alligators incubated at a male-producing temperature. All dosages of E 2 and tamoxifen caused ovarian differentiation at the male-producing temperature, and aromatase activity in these groups was increased accordingly (although only the tamoxifen-treated animals had aromatase activity that was significantly different from control males). The aromatase activity of the ovaries from these sex reversed animals was not significantly different from ovaries of control female alligators. Interestingly, atrazine treatments did not induce sex reversal at a morphological level but did stimulate testicular aromatase activity that was not different from that of ovaries from control animals. Neither vinclozolin nor 2,4 D treatments had any effect on testicular aromatase activity. This suggests that although sex reversal was not induced by atrazine, atrazine altered steroidogenesis in the hatchling alligators such that more estrogen was produced. This increase in circulating estrogen was not detected by radioimmunoassay, although this is expected because there is no difference in control male and female hormone concentrations. A more sensitive and precise assay for E 2 and T could potentially uncover differences in these individual hormones. Using in vivo and in vitro techniques, Connor et al. (1996) concluded that chloro5-triazines do not interact with the estrogen receptor and, thus, any estrogenic or antiestrogenic effects must occur at mechanistic levels other than the estrogen receptor ligand interaction. The data presented in this study suggest that atrazine may affect

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102 organisms at the level of steroidogenic enzyme activity. The ability of atrazine to alter the activity of steroidogenic enzymes has been noted previously in several studies. For instance, the female offspring of rats treated with atrazine during pregnancy and lactation show increased pituitary 5a-reductase and 3a-hydroxysteroid dehydrogenase activity (Kniewald et al., 1987). Conversely, atrazine significantly decreases pituitary 5areductase, 3a-hydroxysteroiddehydrogenase, and 30-hydroxysteroid dehydrogenase activity in adult male rats (Kniewald et al., 1979; Babic-Gojmerac et al., 1989). The aromatase enzyme is of the cytochrome P450 enzyme family and, therefore, it is likely that many compounds (such as atrazine) increase aromatase activity in a similar manner to that of P450 enzymes involved in detoxification. Atrazine could cause greater endocrine disruption in alligators hatched in the wild than that revealed by the present controlled laboratory experiment. As explained in the introduction, the temperature of egg incubation determines the sex of many reptiles, including alligators. Previous studies have indicated that steroid hormones synergize with incubation temperature to exert a greater effect at intermediate temperatures that produce both males and females (Wibbels et al., 1991; Crews et al., 1995b). For instance, males incubated at a male-producing temperature close to the temperature threshold for female development are more sensitive to the effects of estradiol170 (Wibbels et al., 1991). The present experimental design used two temperatures that would produce either 100% females (30°C) or 100% males (33°C). However, if eggs were incubated at a temperature closer to that that produces 50% males and 50% females (pivotal temperature), the effect of aromatase and the resultant steroid environment could be magnified. Future studies should explore this hypothesis.

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103 Previous studies have shown that E 2 and tamoxifen treatments promote ovarian differentiation in reptiles incubated at male-producing temperatures (Gutzke and Bull, 1986; Lance and Bogart, 1992). Although it is not known if such animals are capable of successful reproduction during adulthood, the present study finds no significant difference in aromatase activity between the female control animals and the sex-reversed tamoxifen and E 2 treated animals. This would suggest that at least at hatching, exogenously sexreversed animals are capable of normal steroid production. The present study has indicated that induction and suppression of aromatase enzyme activity are potential modes of contaminant-induced endocrine disruption in a species with temperature-dependent sex determination. Such disruption may not be limited to species with environmental sex determination, as aromatase function can also be affected in species with genetic sex determination. For example, the newt Pleurodeles waltl exhibits a ZZ/ZW system of genetic sex determination in which females are the heterogametic sex and, although sex determination is under gametic control, the sexual differentiation of the gonads can be modified by alterations in temperature (Chardard et al., 1995). When ZW females are incubated at 32°C, aromatase activity in the gonadalmesonephric complex is decreased to male-like levels. This suggests that although sex determination is genetic, the steroidal environment can be easily manipulated by exposure to various extraneous factors including compounds that modify steroidogenic enzymes. Indeed, addition of aromatase inhibitors modify the sexual differentiation of animals with genetic sex determination. When treated with an aromatase inhibitor, genetically female chickens are masculinized but eggs fertilized by these males are not viable (Elbrecht and Smith, 1992). This infertility is species dependent, as genetically female Chinook salmon

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104 (Oncorhynchus tshawytscha) develop into functional males if exposed to an aromatase inhibitor for as little as 2 hours (Piferrer et al., 1994). This study attempted to determine if the endocrine alterations previously observed for wild alligators could be explained by alterations in aromatase activity and if endocrine disruptions, including alterations in aromatase activity, could be induced by embryonic exposure to herbicides used extensively in the habitat of these wild alligators. Aromatase activity was significantly lower in female Apopka animals compared to control animals, a result which supports data from in vitro culture of Apopka gonads (Guillette et al., 1995b). This increased aromatase activity does not, however, explain increases in circulating E2 that have been described previously for wild Apopka females (Guillette et al., 1994). Embryonic exposure to 2,4 D had no effect on the endocrine parameters measured, but atrazine exposure caused aromatase activity that was neither characteristic of males or females. Thus, atrazine exposure may induce endocrine alterations in embryonic alligators. These data emphasize the importance of considering the alteration of steroidogenic enzymes when analyzing contaminant-induced endocrine disruption.

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CHAPTER 6 CELLULAR BIOAVAILABILITY OF NATURAL HORMONES AND ENVIRONMENTAL CONTAMINANTS AS A FUNCTION OF SERUM AND CYTOSOLIC BINDING FACTORS Introduction Environmental contaminants that disrupt the endocrine system can decrease the reproductive success of wildlife and humans. Endocrine disruption from such contaminants can result from an insult to adult reproductive activity (activational disruption) or from an alteration in the embryonic development of the endocrine system (organizational disruption) (Guillette et al., 1995a). Contaminants have been shown to affect endocrine activity by interacting through the hypothalamo-pituitary axis of endocrine control, steroidogenic enzymes, the hepatic clearance rate of steroids, and hormone receptors (Chapter 2). The best described mechanism-of-action of endocrinedisrupting contaminants (EDCs) is their interaction with the estrogen receptor. Many contaminants bind directly to the estrogen receptor, either mimicking (agonistic) or blocking (antagonistic) the effects of estradiol. The ability of an EDC to interact with the estrogen receptor appears to be independent of the chemical class, as many agricultural, industrial, and municipal compounds are capable of binding the estrogen receptor (Guillette et al., 1996a). The ability of contaminants to disrupt endocrine processes depends not only on their ability to interact with hormone receptors, but also on the bioavailability of the EDC Note: This chapter is published in Toxicology and Industrial Health (Crain et al., 1997c). 105

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106 to the receptor. This is controlled by several factors including concentration, sequestration, clearance, and metabolism. These factors are not characterized for most EDCs. However for native steroid hormones, binding proteins in the blood and the cell regulate metabolism, clearance, and bioavailability. Serum or extracellular binding proteins interact with steroids and protect them from renal and hepatic clearance, whereas intracellular proteins reduce steroid metabolism and serve as molecular chaperones. Two of the most well-studied serum binding proteins are sex hormone binding globulin (SHBG) and corticosteroid binding globulin (CBG). When bound to these binding proteins, steroid hormones are unable to interact with receptors and are thus unable to elicit biological effects (Mendel, 1989). The synthesis of both CBG and SHBG are stimulated by estrogens, but SHBG preferentially binds testosterone and estradiol whereas CBG preferentially binds Cortisol and progesterone (Baxter et al., 1995). Circulating binding proteins related to CBG and/or SHBG have been characterized in fish (Martin, 1975), amphibians (Ozon et al., 1971; Martin and Ozon, 1975), reptiles (Salhanick and Callard, 1980; Ho et al., 1987; Paolucci and Di Fiore, 1992), birds (Wingfield et al., 1984), and mammals (Siiteri et al., 1982), indicating a conserved role for these proteins. Intracellular steroid binding proteins are also thought to regulate the amount of free hormone available for receptor binding. Fox (1975) characterized a protein that binds estradiol in the brains of neonatal rats. This neonatal binding protein (NBP) is believed to protect the brain cells from high levels of maternal estrogen. Fox (1975) suggests that without the binding protein, the rats likely would be "sterilized" due to brain exposure to estrogens. Proteins similar to NBP have been characterized in the cytosol of turtle oviduct

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107 (Salhanick et al., 1 979) and avian liver (Dower and Ryan, 1976). Recent studies have begun to characterize the function of a-fetoprotein, a protein found in the cells of developing reproductive and nervous systems of animals (Mizejewski, 1995). afetoprotein regulates the amount of unbound estradiol available for cellular activity and is influenced by local production of growth factors (Nunez 1 994). Wildlife studies have provided specific examples where environmental contaminants have influenced reproductive activity and embryonic development by endocrine disruption (Chapter 1). In particular, reptiles have served as important model species (Guillette and Crain, 1995). Field studies examining alligators in central Florida (USA) have provided observations of abnormal gonadal development in male and female neonates, abnormal plasma sex steroid concentrations in neonates and juveniles, and reduced phallus size in juvenile males (Guillette et al., 1994; Guillette et al., 1995b; Guillette et al., 1996b). The alligators exhibiting these abnormalities hatched from eggs obtained at Lake Apopka, Florida. Previous studies have shown that eggs from this lake have elevated concentrations of a variety of persistent bioaccumulated pesticides and their metabolites such as p,p '-DDE, p,p '-DDD, trans-nonachlor, dieldrin, toxaphene, oxychlordane, and various PCBs (Heinz et al., 1991). A number of these compounds have been shown to have an affinity for the alligator estrogen receptor (Vonier et al., 1996). Do they show any affinity for various plasma binding proteins? The purpose of this study is to examine the binding affinity of circulating and intracellular binding proteins for various environmental contaminants. Given an equivalent concentration of steroid hormone and environmental contaminant, the intracellular bioavailability of contaminants would be approximately equal to that of the native steroids

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108 if steroid binding proteins bind to environmental contaminants with affinities similar to native steroids. If, however, the contaminants do not bind to the binding proteins, then the contaminants are available at a greater concentration for intracellular activities compared to native hormones. This study utilizes competitive binding assays to examine the ability of steroid binding proteins to bind native hormones and environmental contaminants. It is hoped that such an approach will facilitate our understanding of the complex mechanisms of contaminant-induced endocrine disruption. Materials and Methods Buffers and Chemicals All reagents used in buffers were purchased from Sigma Chemical Company (St. Louis, MO). The buffers used in this experiment were as follows: TEEG buffer (20 mM tris, 1 mM EDTA, 1 mM EGTA, 1 mM sodium metavanodate, 10% glycerol, pH 7.4); homogenization buffer (TEEG with 200 uM leupeptin, 525 uM PMSF, 360 uM pepstatin, 10 ug/mL chymostatin, 2 nI/mL aprotinin); and binding buffer (TEEG with 4 uM leupeptin, 10.5 uM PMSF, 7.2 uM pepstatin, 5 mg/mL y-globulin). For binding assays, 170-estradiol, testosterone, and diethylstilbestrol were purchased from Sigma Chemical Company. All environmental chemicals tested were purchased from Chem Service (West Chester, PA). These compounds were: o,p '-DDT (98% purity); p,p '-DDE (99% purity); p,p '-DDD (99.2% purity); dieldrin (98% purity); toxaphene (mixture of isomers); and a,P endosulfan (60% a, 38% P).

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109 Tissue Preparation Oviducts were obtained from 2 gravid female alligators as part of a larger interdisciplinary study of the reproductive biology of the alligators in Florida. Pieces of the anterior section of the oviduct were flash frozen in liquid nitrogen and stored at -72°C. To obtain tissue from gravid Trachemys scripta turtles, 3 adult female turtles were injected with 1 7(}-estradiol (800 ng/g) every other day for 21 days at which point the turtles were in the late gravid reproductive stage. Portions of oviducts from these turtles were flash frozen in liquid nitrogen upon collection and stored at -72°C prior to homogenization. Alligator and turtle oviducts were homogenized using a Polytron homogenizer (Brinkman, Westbury, NY) for 20 sec in 5 volumes of homogenization buffer. Homogenates were centrifuged at 1000 g for 15 min to pellet cellular debris and connective tissue. The supernatant was centrifuged at 100,000 g for 1 hr in a L8-M Beckman Ultracentrifuge. The resultant pellet includes nuclear material, whereas the cytosolic fraction remains in the supernatant. For this assay, only the cytosolic fraction was utilized. This supernatant was flash frozen in liquid nitrogen and stored at -72°C prior to analysis. DNA Cellulose Trachemys scripta cytosol was separated with DNA cellulose to determine if the binding was specific to binding proteins, the estrogen receptor, or both. This separation technique is explained fully elsewhere (Alberts and Herrick, 1970) but, briefly, the DNA on the cellulose beads adheres to any molecules which have DNA binding domains. Because the estrogen receptor (ER) has a DNA binding region, ER binds to the DNA-

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110 coated cellulose beads. After washing the beads of all other material, the ER can be eluted with high salt concentrations. To accomplish DNA-cellulose separation, cytosol was diluted with 5 volumes of binding buffer and this was mixed 1 : 1 with DNA cellulose beads (Pharmacia, Piscataway, NJ). The slurry was mixed for 30 min at 4°C followed by centrifiigation at 5000 g for 10 min. The supernatant (termed the non-ER fraction) was flash frozen and saved for analysis, and the pellet was washed 3x with binding buffer supplemented with 50 mM NaCl. Elution of proteins bound to the DNA cellulose (termed the ER fraction) was accomplished by washing the beads with binding buffer supplemented with 600 mM NaCl. After centrifiigation, the supernatant was flash frozen for analysis. Binding Assay A competitive binding assay (modified from Vonier et al. (1996)) was conducted on the alligator and turtle cytosol. A summary of the tissue preparation, binding assay, and validation is given in Figure 6-1. Cytosol (30 uL) was combined with 165 u.1 binding buffer, and incubated with 6.5 ul tritiated 17-P estradiol (100,000 cpm; dissolved in 10% ethanol; specific activity =157 Ci/mM; Amersham International, Arlington, IL) and competitor (6 uX; dissolved in DMSO) for one hour at room temperature. Competitors used in this study were estradiol170 (E 2 ), testosterone (T), diethyl stilbestrol (DES), o,p DDT, p,p '-DDD, p,p '-DDE, toxaphene, endosulfan, and dieldrin. After adding the assay constituents, tubes were then cooled on ice for 5 min, followed by incubation with 100 ul charcoal-dextran (0.25% charcoal, 0.0025% dextran in 0.0 1M Tris; pH=7.5). After 5 min, the tubes were vortexed again and allowed to sit 5 additional min on ice. Tubes were

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Ill vortexed at 13,000 g for 4 min. Supernatant (200 uL) was diluted with scintillation cocktail and counted on a Beckman LS-20 Scintillation Counter. For all assays, each point was tested in triplicate. The range of estradiol, testosterone, and DES tested was from 1.95 nM to 1 uJvl, whereas the environmental chemicals were tested from 0.58 u.M to 300 uJvi. To calculate % binding, the non-specific binding (no cytosol, but all other assay constituents) was subtracted from the average of the triplicates, and this value was expressed as a percentage of the total binding (cytosol and label, but no competitor). Yeast-Estrogen Screen (YES) Assay The potency of sex hormone binding globulin (SHBG) to bind ligands, and thus reduce transcriptional activity of the estrogen receptor, was measured using a yeast estrogen screen (YES) assay as previously described (Arnold et al., 1996c). This assay is an ideal system for studying the interaction of EDCs with a specific binding protein because yeast have no endogenous steroid binding proteins. Briefly, the yeast expression plasmid pSCW231-hER was incubated with 10 nM 17p-estradiol, ethinyl estradiol, or DES in the presence of varying concentrations of human SHBG (0-0. 1 mg/mL). In this system, activation of the ER stimulates transcription of P-galactosidase, which was measured on a spectrophotometer at 420 nm. Increased p-galactosidase activity indicates greater ER stimulation and less SHBG binding. P-galactosidase activity was expressed as a percentage of 100% transcription in the absence of SHBG.

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112 Oviducts were collected from gravid females. Tissue was homogenized. 1 Cytosolic fraction was separated via centrifugation. Estrogen receptor fraction Cytosol was incubated with removed with DNA cellulose ^ H-estradiol and the competitor niirififfltinn (steroid or contaminant). 1 Unbound H-estradiol was removed with charcoal-dextran. i Amount of 3 H-estradiol bound to the binding proteins was counted on a scintillation counter. Figure 6-1 Overview of the methods used to assess the binding of steroids and contaminants to cytosolic binding proteins. DNA-cellulose purification was used to assure that the binding was not due to estrogen receptor species. See text for an elaboration of the methods.

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113 Results Binding Assays Results for steroid and contaminant binding were similar for alligator and turtle oviductal cytosol. In both alligator and turtle, testosterone and estradiol displaced [ 3 H]-E 2 from the binding proteins. The synthetic estrogen diethylstilbestrol (DES) did not compete for the binding protein site (see Figures 6-2 A and 6-3 A). Neither o,p -DDT nor its metabolites, p,p'-DDD and p,p'-DDE, were able to displace more than 50% of the [ 3 H]-E 2 from the binding proteins. However, o,p '-DDT exhibited greater displacement of [ 3 H]-E 2 than did the other DDT metabolites (300 uM caused 46% and 32% displacement in turtle and alligator cytosol, respectively; see Figures 6IB and 6-2B). Among the other environmental contaminants tested, only toxaphene induced greater than 50% displacement of [ 3 H]-E 2 from the binding proteins (see Figures 6-1C and 6-2C). Toxaphene caused 48% and 66.3% displacement in turtle and alligator cytosol, respectively, at a concentration of 18.7 uM. DNA Cellulose To determine if the binding activity was specific to the E 2 receptor or other proteins, Trachemys scripta cytosol was purified with DNA cellulose. Both the receptor and the non-receptor fractions were tested for their ability to bind [ 3 H]-E 2 . Minimal binding occurred in the ER fraction, whereas the non-ER fraction exhibited significant binding of [ 3 H]-E 2 (see Figure 6-4). E 2 (500 nM) was able to displace [ 3 H]-E 2 in the nonER and ER fractions of the cytosol (see Figure 6-4).

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Figure 6-2. Binding of (a) native estradiol and testosterone and the synthetic estrogen DES, (b) o,p'-DDT, p,p'-DDE, and p,p'-DDD, and (c) toxaphene, endosulfan, and dieldrin to the alligator cytosolic binding proteins. % Binding was calculated by averaging the cpm of the triplicate tubes, subtracting the non-specific binding (no cytosol, but all other assay constituents), and expressing this as a percentage of the total binding (cytosol and label, but no competitor).

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115 A 80.00 e 60.00 .1 40.00 -£0 v» 20.00 -0.00 -20.00 0.001 0.01 0.1 1 10 Ligand Concentration (|xM) B loo.oo j 80.00 Ml .5 60.00 1 40.00 5 20.00 0.00 -20.00 0.001 0.01 0.1 1 10 100 1000 Ligand Concentration ( (t 1Y1 ) 100.00 j 80.00 o 60.00 3 .5 40.00 CP 20.00 -0.00 --20.00 -0.001 0.01 0.1 1 10 100 1000 Ligand Concentration (jiM )

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Figure 6-3. Binding of (a) native estradiol and testosterone and the synthetic estrogen DES, (b) o,p'-DDT, p,p'-DDE, and p,p'-DDD, and (c) toxaphene, endosulfan, and dieldrin to turtle cytosolic binding proteins.

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117 A c -3 e • — 02 100.00 0.001 0.01 0.1 1 Ligand Concentration (\iM) 10 B 100.00 M 3 B CO o.p'-DDT p,p'-DDD p,p'-DDE Estradiol 0.001 0.01 0.1 1 10 100 Ligand Concentration (\iM) 1000

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118 Cytosol Non-ER ER Figure 6-4. Binding of the non-ER and ER fractions of the turtle cytosol. To determine % binding, the binding of each fraction is expressed as a percentage of the binding to pure cytosol. YES Assay The yeast estrogen screen (YES) was used to determine the ability of sex hormone binding globulin (SHBG) to affect gene expression induced by various ER ligands. In the YES assay, SHBG inhibited E 2 -, EE-, and DES-induced 0-galactosidase activity (ER activation) in a dose dependent manner (see Figure 6-5). E 2 -induced ER activation displayed a greater sensitivity to SHBG compared to EE and DES. For example, a physiological concentration of SHBG (0.01 mg/mL) inhibited 30% of estradiol activity, 2% of EE activity, and 0% of DES activity as indicated by ER activation.

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119 120 100 i( 80 > f \ < 40 20 -0 *— Estradiol DES —EE DDT 0.02 0.04 0.06 0.08 SHBG (mg/mL) o.i 0.12 Figure 6-5. YES assay results for 10 nM estradiol, ethinyl estradiol (EE), and DES in the presence of increasing SHBG concentrations. For comparison, YES results for o,p'-DDT and octyl phenol (OP; both at 10 uM) are plotted on the same figure. Data for o,p'-DDT and OP are from Arnold et al. (1996b). Discussion Results of this study indicate that cytosolic and circulating binding proteins have a significantly lower affinity for some environmental chemicals compared to native steroids. This may increase the availability of the environmental chemicals to intracellular steroid receptors and, therefore, increase the cellular potency of such contaminants relative to native hormones. Whereas previous studies have focused on the affinity of the estrogen receptor for environmental chemicals as a major factor determining the estrogenicity of a contaminant, this study indicates that intracellular availability is also important in evaluating the potency of an endocrine-disrupting contaminant.

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120 Both turtle and alligator oviductal cytosol exhibited binding activity consistent with sex steroid binding proteins. These binding proteins were not steroid receptors, as indicated by several lines of evidence. First, testosterone (T) and estradiol (E 2 ) had approximately equal ability to displace [ 3 H]-E2 from the binding proteins. This would not be the case if the binding protein were a receptor species, as E 2 would be a much more potent competitor for the E 2 receptor than would T. Second, diethylstilbestrol (DES) was unable to displace [ 3 H]-E 2 in the cytosol. DES is a potent competitor for the estrogen receptor in numerous species, including turtles (Dufaure et al., 1983) and alligators (Vonier et al., 1997). Third, DNA-cellulose purification exhibited minimal binding activity in the purified estrogen receptor (ER) fraction, whereas significant binding was noted in the non-ER fraction. The minimum binding activity in the ER fraction indicates that there is some ER in the cytosolic preparations, but the activities of such few receptors is masked by other cytosolic binding proteins. Because of this, the cytosolic preparations provide an appropriate system for studying the interaction of EDCs with binding proteins. Binding proteins have been characterized previously in the oviductal cytosol of other animals. Studies on rabbits (Korenman and Rao, 1968) and rats (Katzenellenbogen et al., 1973) have revealed a uterine cytosolic binding protein with similar characteristics to the binding proteins in this study. Salhanick et al. (1979) conducted a detailed characterization of the estrogen-binding proteins in the oviduct of the turtle Chrysemys picta. As in the present study, they found a sex steroid binding protein in the oviductal cytosol that functions similarly to plasma binding proteins such as sex hormone binding globulin (SHBG). Salhanick et al., however, were also able to isolate a significant amount of ER in the oviductal cytosol. Due to the reproductive stage at which the oviducts were

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121 taken, it is likely that our alligator and turtle oviductal tissues contained low ER concentrations. All oviduct samples were taken from gravid animals, just before the animals were to lay eggs. ER levels are likely to be low during this period due to ER down regulation via negative feedback. Few of the environmental contaminants tested in the cytosol binding assay displayed significant ability to displace [ 3 H]-E 2 from the binding proteins. Of the DDT metabolite compounds tested in this study, o,p '-DDT was the only compound to significantly influence the binding of E 2 to cytosolic proteins. Proteins in the turtle cytosol bound o,p '-DDT more effectively than did those in alligator cytosol, with the highest dose (300 uM) displacing 46% of the [ 3 H]-E 2 in turtle cytosol. None of the other DDT congeners displaced more than 33% of [ 3 H]-E 2 in turtle or alligator cytosol. Of the other environmental contaminants tested, only toxaphene exhibited significant binding to the proteins (18.7 uM caused displacement of 48% in turtle and 66.3% in alligator). It is evident that although the same major binding characteristics are present in alligator and turtle cytosol, there is some species specificity in the binding of environmental contaminants to cytosolic proteins. This should be considered in future assessments of the potential potency of endocrine-disrupting contaminants. As cytosolic binding proteins, circulating steroid binding proteins can affect the bioavailability of endogenous hormones and environmental contaminants. Results from this study show that SHBG inhibited E 2 from activating the ER in a dose-dependent manner. Normal circulating concentrations of SHBG in nonpregnant females and males range from 0.003-0.015 mg/mL (Cheng et al., 1983). Using a concentration of 0.01 mg/mL, we found that 30% of E 2 -induced transcription was inhibited while virtually none

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122 of the DESand ethinyl estradiol-induced transcription was inhibited. These results are similar to those of Arnold et al. (1996a; 1996c) who found that synthetic estrogens, p,p DDD, and o,p '-DDT were less affected by SHBG than E 2 in their ability to activate the ER. Arnold et al. (1996c) noted an even greater discrepancy between the influence on E 2 and o,p '-DDT when whole serum from alligators or humans was added instead of SHBG. This suggests that binding factors in blood have a lower capacity to bind environmental contaminants than purified SHBG, and raises further concern about the bioavailability of endocrine-modifying contaminants relative to endogenous hormones. The entire phenomenon of contaminant and hormone binding to plasma proteins is extremely complex, as the most realistic model of interactions involves numerous steroids binding to at least three plasma binding proteins (albumin, CBG, and SHBG) (Tait and Tait, 1991). Levels of SHBG vary throughout the year in seasonally breeding animals, and this can change the exposure of contaminants that have minimal binding to SHBG. In alligators, sex steroid binding protein concentrations are dependent on the gender, sexual maturity, and reproductive stage of the animal (Ho et al., 1987). Ho et al. found that plasma levels of a sex steroid binding protein declined in adult females during the breeding season. Therefore, if some contaminants do bind to these proteins, exposure to these contaminants would increase during reproduction even if total body burdens and environmental exposure remained the same. The results of this study, when synthesized with other related studies, indicate that the potency of some environmental contaminants may be greater than previously suspected based on the cellular availability of the compounds. Figure 6-6 presents a model that depicts this increased cellular availability when compared to native hormones.

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123 Blood-borne contaminants bind to circulating binding proteins with much lower affinity than do native hormones and, thus, the contaminants are more available for transport inside the cell (Figures 6-5A and 6-5B). The actual processes controlling contaminant transport across the cell are poorly understood, and membrane transport itself could be dramatically altered by the interaction of binding proteins and contaminants. Once inside the cell, intracellular binding proteins have low binding affinities for environmental contaminants compared to native steroids, and these environmental contaminants are able to readily interact with hormone receptors (Figure 6-5C). This theoretical model should be considered and tested in future studies of the potency of environmental contaminants that alter the endocrine system.

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Figure 6-6. Theoretical model for the interaction of native hormones and contaminants with serum (sBP) and cytosolic (cBP) binding proteins. (A) Circulating hormones bind readily to blood-borne steroid binding proteins, whereas such proteins have low affinity for most contaminants. This could lead to increased cellular availability for environmental contaminants. The roles of membrane transport facilitators (MTFs) as cellular chaparones are poorly understood. It is possible that serum binding proteins bind to MTFs and influence intracellular transport of hormones or contaminants. For example, Nakhla and Rosner (1996) have shown that SHBG binds to a receptor on the membrane of prostate cells, regulating the intracellular activities of androgens and estrogens. (B) Once inside the cell, intracellular binding proteins bind to hormones, whereas the proteins do not bind contaminants. Therefore, the contaminant is more available for transport into the nucleus. (C) The contaminant binds to the estrogen receptor (ER) and the complex resides in a bound form in the nucleus, where transcription for ER-regulated transcripts will be either increased (in the case of contaminants with agonistic activity) or blocked (in the case of contaminants with antagonistic activity).

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125 Q Extracellular

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CHAPTER 7 SUMMARY AND CONCLUSIONS The Rationale Revisited In Chapter 1, a simplistic model (Figure 1-1) for understanding normal steroid hormone dynamics was introduced. This model illustrates the potential sites at which endocrine-disrupting contaminants (EDCs) can alter the normal dynamics of hormones. The research in this dissertation, as well as many other studies that are reviewed in this dissertation, provide evidence that environmental contaminants can alter steroid dynamics at numerous points in the endogenous cycle of hormones. Figure 7-1 revisits the model presented in Figure 1-1, citing representative studies that describe the effects of EDCs on particular sites in the cycle of steroid dynamics. By far, the most thoroughly studied aspect of this cycle is the effect of EDCs on steroid action. Numerous studies have noted that many EDCs bind with steroid receptors, eliciting either an agonistic or antagonistic effect. Because of the prevalence of these studies and the availability of numerous techniques to measure steroid receptor binding, the phenemenon of receptor interaction has become synonymous with endocrine disruption. As is apparent in Figure 7-1, however, there are many sites independent of the receptor that should be considered when assessing endocrine disruption. 126

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127 The Purpose Revisited The research contained in this dissertation has contributed to the body of knowledge about endocrine-disrupting contaminants by: (1) outlining concepts, such as organization vs. activation and mode of gender determination, that must be considered when endocrine disruption is suspected (Chapter 2), (2) consolidating and summarizing the evidence for contaminant-induced endocrine disruption in wildlife species (Chapter 2), (3) describing hormone profiles of steroid and thyroid hormones in alligators from contaminated and reference lakes (Chapter 3), and (4) examining several specific mechanisms through which environmental contaminants can elicit their responses (Chapters 4, 5, and 6). This research has documented that T production and aromatase activity in vitro can be altered by embryonic exposure to a persistent organochlorine contaminant (p,p -DDD) and a modern use herbicide (atrazine), respectively. However, these changes do not explain the altered hormone profiles noted in wild alligators exposed to these compounds. This suggests that mechanisms other than steroid synthesis are involved in the endocrine disruption of wild alligators. One mechanism that could be involved is an alteration in the bioavailability of steroid hormones. Chapter 6 showed that many pesticides and industrial waste compounds do not interact with steroid-binding proteins and, thus, the contaminants may be more readily available for cellular action than endogenous steroids. Future studies should consider the effect of EDCs on steroid binding protein concentrations. A mechanism that likely is involved in contaminant-induced endocrine disruption is the alteration in the rate of steroid excretion and steroid biotransformation. Enzymes in

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128 the cytochrome P450 family catalyze both steroid excretion and biotransformation. Cytochrome P 450 is a generic term for a superfamily of more than 230 enzymes, all of which contain a heme group and approximately 500 amino acids (Miller, 1988). P 450 enzymes have been studied intensively due to their action in phase I detoxification of natural toxins, drugs, and environmental pollutants. A lesser-studied action of the P 450 enzymes is their role in synthesis/degradation of steroids. The induction of steroidogenic enzymes is responsible for the transformation of one steroid metabolite into another. Several P 450 enzymes regulate adrenal and gonadal steroidogenesis in mammals (see Figure 2-3). Thus the evolution of P 450 enzymes provides two critical adaptations to animals: the ability to detoxify contaminants and the ability to exhibit sexual dimorphism. That is, by modifying the synthesis and degradation of steroid hormones, vertebrates have achieved an internal environment that is dimorphic in sex steroid concentration. The detoxifying and steroidogenic actions of P 450 s are normally treated as separate and distinct functions, but it is proposed that these activities are closely related and, perhaps, regulated by similar mechanisms. Therefore, it is likely that an alteration in steroid excretion and steroid biotransformation rates is involved in endocrine disruption by EDCs. Future studies should test this hypothesis, and examine other potential mechanisms of contaminant-induced endocrine disruption. The Problem Revisited The observation that P 450 induction can lead to enhanced detoxification and altered steroidogenesis leads to an evolutionary dilemma. If an animal is to survive exposure to xenobiotics, that animal requires a well-developed P 450 detoxification system. Evolution is

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129 not driven by survival, however, but by reproduction. In terms of evolution, it matters not how long an individual survives, but how many viable offspring that individual contributes to the population. For example, consider an individual that is born, by chance, with a superior P 450 detoxification system. This individual may live a long life, but never successfully reproduce because of altered steroidogenesis. Another individual with a less effective detoxification system but normal reproduction has higher fitness, because this individual produces viable offspring. Therefore, evolution would favor individuals having efficient detoxification systems (to promote survival so future reproduction can occur) that do not alter reproduction. Millions of years of evolution have produced such a hepatic P450 detoxification system in vertebrates. However, during the last 100 years, animals have been exposed to an increased number of novel, persistent anthropogenic compounds. This brief time period, a matter of several generations in most vertebrate species, would not allow evolution to generate a balance between survival and reproduction. Therefore, many of the reproductive abnormalities currently seen in wildlife and humans could be due to the balance shifting toward survival, with reproduction being compromised. An increase in hepatic P450 induction is but one potential mechanism that could lead to increased survival but decreased reproduction in vertebrates. What are the characteristics of a population experiencing this increased survival but decreased reproduction? First, population demographics would be altered. Fewer offspring would be produced, but offspring that are produced would exhibit an increased survival. There may also be an increased percentage of older animals in the population, as these animals would have an enhanced ability to combat insults from xenobiotics. Males and females have sexually dimorphic hepatic function (Gustafsson, 1994) and, therefore,

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130 the population may also exhibit gender bias. Second, animals in the population would exhibit alterations in their reproductive systems. Abnormal steroid hormone concentrations are expected as a result of altered steroidogenesis. With these altered hormones, reproductive behaviors and sexual characteristics would be changed. Many of these characteristics have been noted before for populations exposed to environmental contaminants (McMaster et al., 1991; Munkittrick et al., 1991; Guillette et al., 1994). As noted in Chapter 1 , the study of the adverse effects of xenobiotics dates back to the earliest humans. We are still far from understanding all the ways that xenobiotics alter our physiology, but our increased understanding of physiological mechanisms has elucidated many of the mechanisms through which xenobiotics can elicit unfavorable effects. For instance, thirty years ago the interaction of contaminants with estrogen receptors was unknown, mainly because the basic understanding of steroid-receptor dynamics were not understood. Therefore, our future understanding of xenobiotic actions is dependent upon both basic and applied research. In a day when humans and wildlife are exposed to increasing amounts of novel compounds, these basic and applied research efforts are both timely and essential.

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131 "2 'o 2 £ 8 5 5 o o o c •> u -3 C >. O aj a & > * 8 ? w 0 *-' t« 1 M X) C (« 3 4> ea 3 P? »-i H 3 ft -2 o J § S -a *"* Z « '5 V ro » 8 •§ 01 * 3 S S o I ill h t w o iS O 3 K sa
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BIOGRAPHICAL SKETCH David Andrew Crain (Drew) was born May 13, 1970 in Spartanburg, South Carolina. He attended public school in Spartanburg and graduated with honors from Dorman High School in 1988. He studied at Clemson University from August of 1988 until May of 1992, receiving Sigma Tau Epsilon, Phi Kappa Phi, and Golden Key National Honor Society memberships. While at Clemson, he received the Mary J. Cox Scholarship, Hessie T. Morrah Scholarship, and Tigerama Scholarship. He received a Bachelor of Science in Biological Sciences degree from Clemson in May 1992. In August 1992, Drew entered the University of Florida as a graduate student in the Department of Zoology. He received his master's degree from the Zoology Department in August, 1994, with a thesis entitled "Insulin-like Growth Factor-I in the plasma of two chelonians: Caretta caretta and Trachemys scripta elegans " While at the University of Florida, Drew received a Grinter Fellowship, a University of Florida College of Liberal Arts and Sciences Fellowship, an International Women's Fishing Association Fellowship, a Sigma Xi Grant-inAid of Research, a Division of Sponsored Research Assistantship, a Learner-Gray American Museum of Natural History Grant for Marine Research, and an Environmental Protection Agency Graduate Student Fellowship. Also while at the University of Florida, he was employed as a graduate teaching assistant for Introductory Biology, Animal Physiology, and Developmental Biology. 153

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154 After graduating with his Ph.D., Drew will join the faculty of the University of Mississippi as an Assistant Professor in the Department of Biology.

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I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Karen A. Bjorndaf Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. David H. Evans Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Larry R. McEdward Associate Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Daniel C. Sharp Professor of Animal Science This dissertation was submitted to the Graduate Faculty of the Department of Zoology in the College of Liberal Arts and Sciences and to the Graduate School and was accepted as partial fulfillment of the requirements for the degree of Doctor of Philosophy. August, 1997 Dean, Graduate School

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I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Karen A. Bjorndaf Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. David H. Evans Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Larry R. McEdward Associate Professor of Zoology I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a thesis for the degree of Doctor of Philosophy. Daniel C. Sharp Professor of Animal Science This dissertation was submitted to the Graduate Faculty of the Department of Zoology in the College of Liberal Arts and Sciences and to the Graduate School and was accepted as partial fulfillment of the requirements for the degree of Doctor of Philosophy. August, 1997 Dean, Graduate School