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The environmental contaminants, PCBs impact on androgen receptor action and prostate growth and development Portigal, Cheryl Lynn 2000

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T H E E N V I R O N M E N T A L C O N T A M I N A N T S , P C B s I M P A C T O N A N D R O G E N R E C E P T O R A C T I O N A N D P R O S T A T E G R O W T H A N D D E V E L O P M E N T by C H E R Y L L Y N N P O R T I G A L B.Sc. M c G i l l University, 1996 A thesis submitted in partial fulfillment of the requirements for the degree of M a s t e r o f Science in The Faculty of Graduate Studies Department of Pathology and Laboratory Medicine We accept this thesis as conforming to the required standard The University of British Columbia 2000 © Cheryl Lynn Portigal, 2000 In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission. Department The University of British Columbia Vancouver, Canada DE-6 (2/88) ABSTRACT Poly chlorinated biphenyls (PCBs) are synthetic chemicals that were used in industrial lubricants and household goods until they were banned in open systems in North America in 1977, after millions of tons were manufactured. P C B s are currently widespread major environmental contaminants that persist in the environment and due to their lipophilic nature, bioaccumulate-exponentially up the food chain. A l l humans, particularly in industrialized countries, carry a burden of P C B s primarily in their adipose tissue. Studies in laboratory animals, wildlife, and humans exposed to P C B spills suggest that P C B s interfere with normal steroid hormone action. Therefore, P C B s and other persistent organochlorines, pose a risk to human health as endocrine disrupting compounds that may target hormone sensitive organs such as the prostate which require steroid hormones for appropriate growth and differentiation. In utero exposure .to compounds that affect steroid hormone action can be the most detrimental to development. In this study the influence of P C B s on the androgen axis in vitro and in vivo were examined. The effects of P C B s on androgen and glucocorticoid regulated reporter gene expression in a prostate cell line, and the ability of P C B s to influence binding of endogenous hormones to the androgen receptor (AR) were tested. Aroclor 1254 acted as a weak androgen receptor agonist, and Aroclor 1254, Aroclor 1242, Aroclor 1260, P C B 42, and P C B 31 were antagonistic to androgen activity at high concentrations relative to natural ligand concentrations. P C B 42 was an additive glucocorticoid receptor (GR) agonist and increased reporter activity to a level 150% higher at 10000 n M than at 0 n M P C B 42. Aroclor 1254 and P C B 42 significantly reduced the levels of D H T binding to the androgen receptor by 25% and 50% respectively, at 1000 n M . Both these compounds reduced D H T levels by 90 to 95% at 10000 n M . The treatment of mice transgenic for a prostate specific probasin promoter driven a C A T reporter gene, with Aroclor 1254 resulted in changes in prostate growth and development. In a single dose level study, mice were treated in utero and until four weeks (prepubertal) and eight weeks (post pubertal) of age with 10 mg/kg/day of Aroclor 1254. Additionally, in a dose response study, mice were treated in utero and until eight weeks of age with 10, 20, and 40 mg/kg/day. Prostate, testis, epididymis, heart, kidney, and liver weights were measured for comparison between groups. Results indicated that Aroclor 1254 has the capacity to reduce prostate weight and increase liver weight (both normalized for body weight) in a dose dependent manner. Finally, prostatic C A T activity was measured. This P C B mixture significantly reduced C A T activity in eight week old treatment mice in the single dose level study. Histology of prostate and liver were also examined in the dose response group. Histological alterations included dose related changes in liver vacuolization, as well as lymphocytic infiltrations and an increase in the presence of predominantly dilated acini in the prostate in a dose dependent manner. These findings demonstrate that P C B s can influence steroid hormone action through the androgen axis in vitro, and interfere with androgen-regulated gene expression in vivo. TABLE OF CONTENTS A B S T R A C T i i L I S T O F T A B L E S v i i i L IST O F F I G U R E S ix A C K N O W L E D G E M E N T S x i LIST OF A B B R E V I A T I O N S x i i I N T R O D U C T I O N 1 1.1 Endocrine Disruptors 1 1.1.1 Discovery 1 1.1.2 Target Sites .. 2 1.1.3 Mechanisms of Act ion of Endocrine Disruptors 2 1.1.4 Diethylstilbesterol 6 1.1.5 Endocrine Disruptors in the Environment 8 1.2 Polychlorinated Biphenyls (PCBs) 10 1.2.1 Industrial Use of P C B s 10 1.2.2 P C B Exposure in the General Population and Groups with High Levels of Exposure 13 1.2.3 Health Effects of P C B s 15 1.3 Endocrine Disrupting Characteristics of P C B s 18 1.3.1 In vitro Evidence 18 1.3.2 In vivo Evidence 19 1.3.1.1 Female Reproductive Effects 21 1.3.1.2 Male Reproductive Effects 22 iv 1.4 The Prostate 22 1.4.1 Structure and Function 22 1.4.2 The Androgen Receptor and Hormonal Regulation 23 1.4.3 Xenobiotic Effects on the Prostate 24 1.5 Rationale and Objectives of this Study 27 E X P E R I M E N T A L P R O C E D U R E S 29 2.1 In vitro Studies 29 2.1.1 Chemicals 29 2.1.2 Plasmid Constructs 31 2.1.3 Ce l l Culture Transfections 32 2.1.4 Ce l l Culture Ligand Displacement Assay 33 2.2 In vivo Studies 35 2.2.1 Animals 35 2.2.2 Treatment and Groupings 36 2.2.3 C A T Assay 37 2.2.4 Histology 37 2.2.4.1 Measurement of Percent Area of Stroma and Ducts in Prostate Tissue Sections 38 2.2.5 Measurement of Serum Testosterone Levels 38 2.2.6 Statistics 39 R E S U L T S 40 3.1 Induction of Luciferase Reporter Gene 40 3.1.1 Induction by D H T and D E X 40 V 3.1.2 Induction of Luciferase Activi ty Through A R by Aroclor 1254 42 3.1.3 Alteration of Luciferase Activi ty Induction Through A R by Aroclor 1242 45 3.1.4 Alteration of Luciferase Activi ty Induction Through A R by Aroclor 1248 48 3.1.5 Alteration o f Luciferase Activi ty Induction Through A R by Aroclor 1260 48 3.1.6 Alteration of Luciferase Activi ty Induction Through A R and G R by P C B 42 50 3.1.7 Alteration of Luciferase Activi ty Induction Through A R by P C B 31 50 3.2 Displacement of 3 H - D H T as A R Ligand by P C B s 53 3.3 Analysis o f C A T Activi ty 56 3.3.1 Single dose level Group 57 3.3.2 Dose Response Group 59 3.4 Alteration of Organ Weights 59 3.4.1 Liver 61 3.4.2 Prostate 61 3.4.3 Serum Testosterone Levels 63 3.5 Pathological Analysis of Histological Sections 65 3.5.1 Liver 65 3.5.2 Prostate 68 3.5.2.1 Comparison of Stromal Areas in Prostate Tissue 71 3.5.3 Testes 72 D I S C U S S I O N 74 4.1 Luciferase Reporter Gene Assay 74 4.2 Ligand Displacement Assay 76 4.3 C A T Assay .' 77 vi 4.4 Organ Weights and Histopathology 78 4.4.1 Liver ; 78 4.4.2 Prostate , 79 4.5 Future Directions 81 B I B L I O G R A P H Y 83 vii LIST OF TABLES Table 1 Percent content by weight of some individual congeners in four Aroclor mixtures 30 Table 2 Corrected organ weights of L P B - C A T mice treated to four and eight weeks of age with vehicle only (control) and 10 mg/kg/day of Aroclor 1254 62 Table 3 Corrected organ weights of L P B - C A T mice treated to eight weeks of age with vehicle only (control), 10, 20, 40 mg/kg/day of Aroclor 1254 62 vm LIST OF FIGURES Figure 1 Chemical structure of some steroid hormones 5 Figure 2 Chemical structure of some endocrine disruptors 7 Figure 3 General chemical structure of chlorinated biphenyls 12 Figure 4 Functional domains of the androgen receptor 26 Figure 5 Luciferase activity in L N C a P cells transiently transfected with A R or G R and A R R 3 t k - l u c and exposed to D H T and D E X titrations 41 Figure 6 Luciferase activity in L N C a P cells transiently transfected with A R and A R R 3 t k - l u c and exposed to Aroclor 1254 43 Figure 7 Luciferase activity in L N C a P cells transiently transfected with A R and ARR3tk-luc and exposed to Aroclor 1242. 46 Figure 8 Luciferase activity in L N C a P cells transiently transfected with A R and A R R 3 t k - l u c and exposed to Aroclor 1248 , 47 Figure 9 Luciferase activity in L N C a P cells transiently transfected with A R and A R R 3 t k - l u c and exposed to Aroclor 1260 49 Figure 10 Luciferase activity in L N C a P cells transiently transfected with A R or G R and A R R 3 t k - l u c and exposed to P C B 42 51 Figure 11 Luciferase activity in L N C a P cells transiently transfected with A R and A R R 3 t k - l u c and exposed to P C B 31 . . . . : 52 Figure 12 Percent of 3 H - D H T bound to A R in H e L a FLAG-tagged A R cells after, exposure to tritiated and nontritiated D H T , and tritiated D H T and P C B s 55 Figure 13 Activi ty of C A T reporter transgene linked to the prostate specific probasin promoter from prostate tissue extracts of four and eight week old control and treatment L P B - C A T mice in the single dose level study 58 Figure 14 Activi ty of C A T reporter transgene linked to the prostate specific probasin promoter from prostate tissue extracts of eight week old L P B - C A T mice in the dose response study treated with vehicle only (control), 10, 20, and 40 mg/kg/day 60 Figure 15 Scatter plot of serum testosterone levels from eight week old L P B - C A T mice in the dose response study treated with vehicle only (control), 10, 20, 40 mg/kg/day 64 Figure 16 Centrilobular vacuolozation in the liver of an eight week old L P B - C A T mouse treated with 40 mg/kg/day of Aroclor 1254 67 Figure 17 Periportal vacuolization in the liver of an eight week old L P B - C A T mouse treated with 20 mg/kg/day of Aroclor 1254 67 Figure 18 Accumulation of lymphocytes and small macrophages in periportal spaces in the liver of an eight week old L P B - C A T mouse treated with 20 mg/kg/day of Aroclor 1254 67 Figure 19 Dilated acini of the prostate of an eight week old control group L P B - C A T mouse 69 Figure 20 Non-dilated acini of the prostate of an eight week old L P B - C A T control group mouse 69 Figure 21 Infiltration of discrete aggregates of small lymphocytes in the interstitial tissue of the prostate of an eight week old L P B - C A T mouse treated with 40 mg/kg/day of Aroclor 1254 69 Figure 22 Percent ratio of stromal area to total area in prostate tissue from eight week old L P B - C A T mice treated with vehicle only (control), 10,20,40 mg/kg/day of Aroclor 1254 73 ACKNOWLEDGEMENTS I would like to thank my supervisor Dr. Colleen Nelson for her guidance, support, and enthusiasm throughout the time we have worked together, particularly my master's research project. I would also like to thank my research supervisory committee—Drs. Bal ly , Gleave, and Rennie, for their advice and constructive suggestions. I appreciate the members of the Prostate Centre at V G H and Cancer Endocrinology, and Dr. Stephane Lair of the U B C Animal Care Centre, who shared their time, knowledge, and expertise with me. Finally, I would like to express my gratitude to my family and friends for their continual support and encouragement. XI LIST OF ABBREVIATIONS A B P = Androgen Binding Protein A h R = A r y l Hydrocarbon Receptor A N O V A = Analysis of Variance A P E O s = Alkylphenol Polyethoxylates A R = Androgen Receptor C A T = Chloramphenicol Acetyl Transferase C Y P = Cytochrome P450 D B D = D N A Binding Domain D E S = Diethylstilbesterol D E X = Dexamethasone D H T = Dihydrotestosterone D M E M = Dulbecco's Modified Eagle's Medium E D T A = Ethylenediaminetetraacetic A c i d E G T A = Ethyleneglycol-bis[B-aminoethylether]-N,N,N,N',N'-tetraacetic A c i d E R = Estrogen Receptor G R = Glucocorticoid Receptor H E P E S = N-2-Hydroxyethylpiperazine-N'-2-ethanesulfonic A c i d L B D = Ligand Binding Domain L D 5 0 = Dose which is Lethal to 50% of the Test Animals L O A E L = Lowest Observed Adverse Effect Level M F O s = Mixed Function Oxidases M R = Mineralocorticoid Receptor N B F = Neutral Buffered Formalin N O A E L = N o Observed Adverse Effect Level P C B - Polychlorinated Biphenyl P C D F = Polychlorinated Dibenzofuran P R = Progesterone Receptor SPBs = Serum Binding Proteins S-FBS = Stripped Fetal Bovine Serum S E M = Standard Error of the Mean T3 = Triiodothyronine T4 = Thyroxine T C D D = 2,3,4,8-Tetrachlorodibenzo-P-Dioxin T E F = T C D D Equivalency Factor T R = Thyroid Hormone Receptor CHAPTER 1. INTRODUCTION 1.1 Endocrine Disruptors 1.1.1 Discovery The concept of endocrine disruptors has come to light over the past 60 years, as scientists have begun to piece together evidence of alarming changes in wildlife populations. The disappearance of mammal and bird populations in Europe and North America, and the high incidences of behavioral abnormalities, and reproductive and developmental defects were initially inexplicable. These anomalies were observed in invertebrates, fish, reptiles, birds, and mammals (including humans). The possibility that this phenomenon was due to environmental contaminants was considered (J). Abnormalities in specific groups in locations with high levels of pesticides and other organochlorine compounds suggested the link between environmental contaminants and the observed alterations in wildlife species (2-6). A decrease in sperm counts and sperm quality during the 20 t h century was reported in men from Asia , South America, North America, and Europe (7 ) . The rate of such changes could not be explained simply by genetic factors, and again, environmental contaminants became suspect (8). Endocrine disrupting compounds mimic, inhibit, or alter the ability of natural hormones to act in the regulation of endocrine systems (9). More specifically, they are exogenous agents that interfere with the synthesis, storage, release, transport, metabolism, binding, action or elimination of natural hormones (1). Endogenous hormones regulate homeostasis and development, and thus alteration of these processes 1 causes disruption in critical aspects of cellular, organ, and organismal function. There is a wide range of effects on steroid hormone systems due to exposure to endocrine disruptors which depends on the nature of the compound such as toxicity, and persistence, as well as the amount, time of life that exposure occurs, and duration of exposure (10). 1.1.2 Target Sites Endocrine modulating compounds can affect a number of systems in the body. The more common sites of disruption include the immune system, thyroid function, and neurodevelopment and behaviour (10). These chemicals have also been linked to hormone-related cancers such as those of the breast, testis, and prostate, and can be harmful to the reproductive system (9). The threat of reproductive effects such as alterations in hormone balance and production (4), developmental abnormalities of the reproductive tract (11), and a global decrease in human sperm counts (8) implicates endocrine disruption as a contributing factor to the compromised reproductive fitness of many species, including humans. 1.1.3 Mechanisms of Action of Endocrine Disruptors Different mechanisms of action of environmental contaminants have been determined as pathways that lead to impaired steroid hormonal control. Normal steroid hormone function begins with a small molecule synthesized from blood cholesterol and absorbed into the appropriate gland (12). These molecules generally act at the level of transcription initiation and affect m R N A synthesis, which results in a modification in protein synthesis. Once a steroid hormone enters the cell, it forms a high affinity complex with the cognate steroid hormone receptor by binding to the ligand binding 2 domain. The complex enters the cell nucleus and binds as a dimer to D N A via an interaction between a discrete domain in the protein with a specific response element sequence in the regulatory region of the target genes. Recruitment of coregulatory proteins to the DNA-bound dimer enables the hormone to modulate gene transcription and ultimately protein levels (12). The synthesized proteins mediate the effects of steroid hormones on target tissues including metabolic activation of cellular growth, and activation of cellular differentiation. In the case of the steroid hormone testosterone (Figure 1), this hormone is synthesized mainly in the testes, through a reaction involving conversion from androstenedione by the action of 1713-hydroxysteroid oxidoreductase. Most of the biosynthesized testosterone from the testis goes into peripheral circulation and is bound to serum binding proteins (SBPs), although 2% of the total in the plasma is free testosterone and remains unbound. A fraction of the testosterone diffuses from the Leydig cells into the seminiferous tubules, Sertoli cells and germ cells for the maintenance of spermatogenesis and to the genital tract, including the epididymis. One of the mechanisms by which testosterone reaches the epididymis, is through transport by androgen binding protein (ABP) . Testosterone also contributes to the biosynthesis of A B P , and is involved in the activation of spermatozoa (12). Testosterone is converted to dihydrotestosterone (DHT) (Figure 1) by 5(X-reductase. D H T exerts it action by binding to the ligand binding domain of the A R , enters the nucleus and binds as a dimer through the D N A binding domain to nuclear D N A . Coregulatory proteins are recruited and androgen regulated responses in the cell are initiated through gene transcription and protein production. 3 T and other steroid hormones are made by a series of biosynthetic enzymes and are metabolized by enzymes known as mixed function oxidases (MFOs) which are part of the cytochrome P450 ( C Y P ) superfamily of enzymes involved in the synthesis and catabolism of steroid hormones and in the detoxification of xenobiotics (13). These processes enable the synthesis of steroid hormones from cholesterol, and transform these hormones from one type to another. For example, aromatase transforms androgens to estrogens (12). The catabolism of steroid hormones involves reductions and hydroxylations to conjugate the steroid with an acid to allow for elimination, which is mostly through the urine. Many o f these reactions take place in the liver, particularly in the case of androgens.' A n endocrine disruptor can alter the synthesis of steroid hormones (14) or alter the transport and clearance of hormones by binding to steroid binding proteins in the serum. This prevents endogenous steroid hormones from binding to these carrier proteins, resulting in an increased level of steroid hormones within the cell (75). Environmental hormones are also known to alter steroid hormone receptor activation indirectly by downregulating the receptor by decreasing sensitivity of the receptor to the ligand (9). Finally, the most common known mechanism of endocrine disruption is through direct hormone receptor binding by the contaminant and its metabolites. Steroid hormone receptors including the A R and estrogen receptor (ER) are considered to have a significant amount of latitude in binding specificity (16, 17). OH Dihydrotestosterone (DHT) Testosterone Dexamethasone (DEX) Figure 1. Chemical structure of some steroid hormones. B y mimicking the natural ligand or inhibiting the natural ligand from binding to the receptor, a compound can act as a steroid hormone agonist or antagonist, respectively. This has been shown through in vitro hormone binding and transcriptional assays that demonstrated the ability of many environmental contaminants to interact at the molecular level with one or more steroid hormone receptors (17-20). 1.1.4 Diethylstilbesterol One of the first compounds found to interfere with normal hormone action was diethylstilbesterol (DES) (Figure 2). This drug is a synthetic estrogen discovered in 1938 and administered to pregnant women to prevent miscarriages and premature births. It was also used to suppress lactation after childbirth and to reduce symptoms of menopause. In agriculture, D E S was used to increase the rate of weight gain in livestock. In the 1970s the teratogenic nature of D E S was discovered. This highly estrogenic non-steroidal chemical causes reproductive abnormalities such as vaginal adenocarcinoma in females (21) and hypospadias and other genital tract abnormalities in males (22, 23) exposed in utero. D E S has also been associated with a higher risk of breast cancer in women who took this drug during pregnancy to prevent miscarriages (23). Some of these abnormalities include effects on the prostate. Animal studies have shown that pregnant female mice fed 200 ug/kg/day have male offspring with reduced prostate weights (24, 25). A reduction in prostate weight was also observed in male offspring of female rats whose drinking water contained 100 ug/1 of D E S (26). B y contrast, lower doses of D E S ranging from 0.02 to 2 ng/kg/day to female mice during gestation caused an increase in prostate weight in the male offspring in adulthood (24, 25). 6 OH Diethylstilbesterol (DES) Figure 2. Chemical structures of some known endocrine disruptors in the environment. 7 In utero exposure to D E S has been linked to compromised immune function in animal models (27, 28) and in humans (7). The high potency of D E S can be partly explained because it has a low binding affinity for serum hormone binding proteins and is therefore more bioavailable than endogenous estrogens. This characteristic suggests that in vivo, D E S can cross the placental barrier and accumulate inside target cells at a higher rate than estradiol (10). The unfortunate finding that this drug which millions of people were exposed to is carcinogenic and teratogenic was an important finding regarding the harmful effects of synthetic agents that act as hormones, particularly in utero. 1.1.5 Endocrine Disruptors in the Environment In addition to D E S , the existence of other compounds that possess endocrine disrupting activity has been discovered. Most of these chemicals are organohalide environmental contaminants such as pesticides, fungicides, herbicides, alkylphenols, and polychlorinated biphenyls (PCBs) (29). The Canadian Environmental Protection Agency and the United States Endocrine Disruptors Screening and Testing Advisory Committee ( E D S T A C ) are examining chemicals in the environment for endocrine disrupting potential. There are currently 87 000 chemicals in commerce which require evaluation for all potential endocrine disrupting capacities (30). Over 120 chemicals that are released into the environment have been found to have endocrine disrupting activities (10). The pesticide l,l,l-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) (Figure 2) and its metabolites are quite lipophilic, poorly metabolized, and bioaccumulate in the food chain (11). D D T was banned in the U.S . in 1972, although it is still in use in developing 8 countries and continues to accumulate in the global environment through air currents (11). This compound is therefore found ubiquitously in all individuals at varying levels. Reduced breeding in fish-eating birds has been attributed to the environmental toxicity of D D T (7). This compound binds to the E R , activating estrogen regulated gene transcription and increasing estrogen metabolism (10, 31). The D D T metabolite, p , p ' - D D E (Figure 2) binds to the A R and is antagonistic, therefore causing inhibition of androgen regulated sexual differentiation in rodents and reptiles at environmentally relevant levels (19). The dicarboxamide fungicide vinclozolin (3-(3,5-dichlorophenyl)-5-methyl-5-vinyl-oxazolidine-2,4-dione) (Figure 2) is used to control fungal growth on fruits, vegetables, ornamental plants, and turfgrass (32). This compound has also been shown to have endocrine disrupting capabilities. Vinclozol in alters sexual differentiation in male rats treated in utero and modifies their development in an antiandrogenic manner (33). Metabolites of vinclozolin ( M l and M 2 ) compete with D H T for binding to the A R and inhibit DHT-induced transcriptional activity in vitro (18). The fungicide hexachlorobenzene (Figure 2) was originally introduced in 1945 for agricultural use as a seed protectant treatment, particularly to control bunt growth on wheat. This chemical interferes with mechanisms that regulate ovarian steroidogenesis (34), alters menstrual cycle characteristics (35), and decreases progesterone serum levels (36) in mammals. Alkylphenol polyethoxylates (APEOs) are nonionic surfactants, which have been produced since the 1940s. These compounds are widely used in detergents, herbicides, pesticides, and paints. A P E O s and their metabolites, such as nonylphenol (Figure 2) are 9 released through sewage treatment and are leached from plastics, particularly when heated. A P E O s have been known to displace estrogen from the E R since 1978 (20). The observation that polystyrene tubes can release nonylphenols, which stimulate the estrogen-dependent growth of M C F - 7 breast adenocarninomca cells (15), was a powerful endocrine effect that brought attention back to this group of endocrine disruptors. The ubiquitous environmental contaminants P C B s (Figure 3), the focus of this study, are also considered to have endocrine disrupting potential, in addition to a number of other health effects (37). The magnitude of the effects of endocrine disruptors on steroid hormone systems is vast and certainly not yet fully elucidated. B y determining the endocrine effects of environmental contaminants such as P C B s , we may begin to understand and address the rapidly changing state of human and ecological reproductive health. 1.2 Polychlorinated Biphenyls (PCBs) 1.2.1 Industrial Use of PCBs Polychlorinated biphenyls are persistent, toxic environmental pollutants which have bioaccumulated in the environment. These compounds were produced commercially as mixtures of up to 209 isomers or congeners. P C B s have been manufactured in the U S A since 1929 and had common commercial use, mainly as coolants and lubricators in electrical equipment by 1930 (38). P C B s also became common components of general use products such as solvent extenders, flame retardants, organic diluents, inks, dyes, paints, and adhesives, and were found in carbonless copy paper, newsprint, and caulking compounds (39). 10 12 Although P C B manufacture was generally banned in the United States in the late 1970s, millions of tons were released into the environment between 1929 and 1977 and are still detected in all levels of the food chain in a biomagnified manner (37), as P C B s become further concentrated in animals with each successive step up the food chain (40). Slow accumulation of P C B s in adipose tissues over time leads to tissue concentrations far above acutely toxic intake levels that would be detected in the blood, immediately after exposure. However, through the food chain, a predator can ingest a fairly high dose of P C B s from its prey. This process continues up to the top predators as P C B biomagnification takes place (41). For example, in the Lake Michigan food web P C B s increase 12.9 times from plankton to fish (42). 1.2.2 PCB Exposure in the General Population and Groups with High Levels of Exposure P C B s are highly stable compounds, which are fat soluble, bioaccumulative, and ubiquitously present in all humans. The estimated daily exposure in the United States of America from 1982 to 1984 was approximately 40 ng/day (43). In this study, 85% of the P C B s that adult men acquired from their diet was from meat products (43). The offspring of women exposed to P C B s also experience P C B exposure in utero and during lactation. Even though P C B s are detected in human tissues and breast milk at levels which are considered low regarding most toxicological endpoints, these concentrations have been reported to be harmful to reproductive, developmental, and endocrine processes (44, 45). The average P C B concentration in whole human breast milk ranges from 10 to 180 ng/ml, which leads to a range of 1.5 to 27 |^g/kg/day of P C B s that breast fed infants may consume (46). Infants and young children consume more food per body weight than 13 adults, and may have a greater relative exposure to P C B s than adults which may contribute to the higher susceptibility of infants and young children to P C B s (47). P C B s have been reported to cross the placenta in a study which correlated the concentration of P C B s in maternal serum (4.7 ng/ml) to that in cord serum (2 ng/ml) (48). Due to their lipophilic nature, P C B s are stored in adipose tissues. Therefore, it should be noted that while many studies report P C B levels in serum, this value is likely well below the levels stored in tissues such as fat. Individuals exposed to P C B s by occupational exposure or ingestion of highly contaminated foods can have much higher body burdens of P C B s . Capacitor plant workers in the United States have reported serum P C B levels ranging from 1 up to 1700 ng/ml (49). People who eat a diet rich in Great Lakes fish have serum P C B levels ranging from 7 to 366 ng/ml, depending on the amount of fish they consumed (49). Individuals exposed to P C B s by eating P C B contaminated rice o i l in Japan in the late 1960s developed a condition termed "Yusho". A similar case of P C B poisoning occurred in Taiwan in 1979 which was termed "Yu-Cheng", resulting in the same symptoms seen in Japan. Yusho and Yu-Cheng patients had levels of P C B s reported in the 50 ng/ml range as well as polychlorinated dibenzofurans (PCDFs) at levels around 0.1 ng/ml in their blood (50). Children of exposed Yusho and Yu-Cheng mothers developed distinct clinical manifestations including dark brown skin and mucous membrane pigmentation, edematous eye, dentition at birth, abnormal skull calcification, and low birth weight (51). The effects seen in these children demonstrates the teratogenicity of organochlorines such as P C B s and P C D F s . 14 Finally, the Inuit population is exposed to large amounts of P C B s through a diet rich in seals and other marine mammals which are near the top of the Arctic aquatic food chain (52). These animals themselves have bioaccumulated levels of P C B s as high as 8ug/g of l ipid, which are stored at highest levels in the blubber at 1.9 ug/g (wet weight) and are also detected in other tissues consumed as meat at 0.8 u.g/g (53). Inuit women in northern Quebec have approximately 7 p.g/g of lipid of P C B s in their breast milk, which is 7 fold higher than that found in Caucasian women in southern Quebec (52). 1.2.3 Heal th Effects of P C B s P C B s have a wide range of health effects on different systems depending on the type of P C B , the amount of P C B in the exposure, the length of exposure, and the organism exposed (37). The more highly chlorinated P C B s tend to bioaccumulate more and have a lower rate of metabolism compared to lower chlorinated P C B s . This is well demonstrated in a study of the tissues of rats dosed with Aroclor 1254. Relative to Aroclor 1254, the brain, liver, blood, and adipose tissue contained a lower concentration of tetra- and penta- congeners, and a greater concentration of heavily chlorinated congeners in the hexa- to nona- range (54). The more P C B s that enter the body, the more harmful effects that may be seen. Although these compounds have high L D 5 0 s of approximately 1300 mg/kg (depending on the components of the P C B mixture), some effects due to acute exposure are more severe than those due to chronic exposure. The bioaccumulative nature of P C B s is also a factor regarding the length of exposure. A n individual w i l l accumulate P C B s in their tissues during the time they are exposed and some of these P C B s w i l l remain for months or years without being metabolized. P C B toxicity studies have been performed in various organisms including mice, rats, mink, 15 guinea pigs, and monkeys. In some studies mink and monkeys exhibit a higher degree of sensitivity to P C B s than other species such as mice, rats, and rabbits, particularly regarding the influence of P C B s on body weight (55, 56). Variation of P C B effects in different species may be modulated by species-specific discrepancies in l ipid metabolism, enzyme induction as wel l as quantitative differences in P C B s binding to receptors in target organs (57). The toxicity of P C B s in humans has only been measured in individuals exposed through occupational or accidental exposure. Therefore, we must also rely on evidence of these effects in animal studies to suggest how human systems may be affected. P C B s have caused renal damage in animal studies. Rats treated with 100 mg/kg/day of P C B s for 3 days per week for 3 weeks developed renal tubular damage, including vacuolated tubular epithelial cells with fatty deposits, and epithelial cells and proteinaceous casts in the tubular lumens and urine (58). N o significant effects on kidney weight has been reported (59). Dermal effects due to P C B exposure has been reported in humans as well as laboratory animals. The levels of P C B s in these exposures range from less than 0.1 mg/m to 11 mg/m over at least 4 months in the humans and as low as 0.1 mg/kg/day in monkeys fed P C B s for 2 months (59). The symptoms include facial edema, chloracne, alopecia, thickening and pigmentation disturbances of the skin and nails, and fingernail loss (60-64). The immune system has also been shown to be susceptible to the effects of P C B s . Intermediate duration exposure studies have induced a decrease in serum levels of immunoglobulins and an increase in susceptibility to bacterial infections in mice treated 16 with 22mg/kg/day of P C B s for 3 to 6 weeks (65-67). P C B s have also been shown to cause similar effects in monkeys (68, 69). Neurological effects of P C B s have been reported in humans and animals. Individuals such as the Inuit with a diet high in fish from P C B contaminated waters and those consuming fish from the Great Lakes are susceptible to these effects. P C B exposure has been correlated to changes in neuropsychological functioning in groups of people with a diet rich in fish (70). Studies of P C B effects at different stages of development in humans and rodents show alterations in dopamine levels (59). P C B s can also affect memory, attention, and learning in humans and monkeys exposed in utero and through lactation (71,72). Although P C B s can effect all systems in the body, one of the most notable and consistent effects is on the liver. P C B doses above 0.3 mg/kg/day for 3 to 15 weeks have been shown to cause increased liver weights in rats (44, 73-78), although biochemical alterations were not observed in all of these studies. The lowest observed adverse effect level ( L O A E L ) for P C B induced hepatic effects in mice is 200 mg/kg, and the no observed adverse effect level ( N O A E L ) is a single dose of 50 mg/kg. Histological effects of P C B s in rats are seen at doses of 50 mg/kg/day for 30 days and 11 months and of 5 mg/kg/day for 6 months. The effects from these studies include hepatocyte hypertrophy, fat deposition, fibrosis, necrosis, and changes in serum levels of liver associated enzymes indicative of possible hepatocellular damage (59). P C B s are known to have effects on induction of enzymes in the cytochrome P450 (CYP) system. C Y P and M F O s involved in P C B metabolism are primarly located in the liver (79) and are involved in the biosynthesis and catabolism of steroid hormones as well 17 as the detoxification of foreign compounds that enter the body. The presence of chlorine substitutions on the carbon atoms in a P C B molecule blocks metabolism at the position of substitution. M F O s in the liver metabolize P C B s to arene oxide intermediates, which undergo oxidationto polychlorinated hydroxybiphenyls, followed by spontaneous or epoxide hydrolase mediated reduction (80). P C B s that have been hydroxylated or sulfonated by phase II enzymes are known to bind to steroid hormone receptors (81-85) and to be estrogenic (85, 86). C Y P 1A and 2B enzymes are both thought to be involved in P C B metabolism (87). One of the intriguing aspects of P C B action on metabolism is that they induce the production of the enzymes that metabolize them. Aroclor 1254 is known to stimulate the expression of C Y P enzymes including some of the C Y P 1A, 2A, 2B, 3 A enzymes. Epoxide hydrolase, glucoronosyl transferase, glutathione S transferase, and some reductases are also induced by Aroclor 1254 (88). In addition to effects on specific organ systems in the body, P C B s can affect body weight. In animal studies, weight loss after a single high dose was attributed to dehydration (58). Intermediate and chronic duration dietary administration of P C B s often causes a decrease in body weight or body weight gain, which is considered to constitute a wasting syndrome (59). 1.3 Endocrine Disrupting Characteristics of PCBs 1.3.1 In vitro Evidence A common method used to assess the estrogenic effects of a compound is the E -S C R E E N assay. This method tests the estrogenicity of environmental chemicals by measuring the proliferative effect on M C F - 7 cells (86). This study found five P C B congeners as wel l as some of the hydroxylated metabolites of some congeners to be estrogenic. Other studies examined the binding affinity of hydroxylated P C B congeners 18 and found the most potent congener was 2,4,6,2',6 '-pentachloro-4-biphenylol, which was approximately 5 times less potent than 1713-estradiol (89). The congeners found to be estrogenic were not coplanar, ie. they did not have the two benzene rings in the same plane. The degree of planarity is determined by the presence of ortho-substituted chlorines, which cause steric hindrance to rotation. Coplanar P C B s have one or less ortho-substituted chlorine atoms (90). Coplanar P C B s are thought to act through the aryl hydrocarbon receptor (AhR) and exhibit dioxin-like characteristics, and are the most toxic P C B s when assessed for their T C D D equivalency factor (TEF) (57). Transcriptional assays have found some hydroxylated P C B congeners bound to the E R and altered gene expression mainly in an antiestrogenic manner and decreased reporter gene expression (82, 91). There have been no previous reports demonstrating the ability of P C B s to bind to the thyroid hormone receptor (TR) or A R in vitro. 1.3.2 In vivo Evidence Studies of endocrine disruption by P C B s have examined a number of hormonal systems and have been conducted in animals and in humans through epidemiological analysis. A study was conducted on the effects of P C B s on turtles that utilize temperature-dependent sex determination (common in many egg-laying reptiles). The investigators found that in turtle eggs incubated at male-producing temperatures, the estrogenic effects of P C B s reversed the gonadal sex to female (92). Many epidemiological studies on breast cancer and environmental exposure to P C B s have been conducted. The previously mentioned in vitro studies on the estrogenicity of P C B s suggest that these compounds are contributing factors to high rates of breast cancer in industrialized countries. However, the relationship between P C B s and 19 breast cancer remains unclear. Some studies show a significant relationship involving a correlation between breast adipose tissue levels of P C B congeners and breast cancer risk (93), and the ability of P C B s to produce free radical-mediated oxidative D N A damage during oxidation of lower chlorinated biphenyls (94). The detection of increased oxidative damage of D N A in human breast tumour tissue suggests that this area warrants further investigation. Another study found that P C B s are present in breast cyst fluids and therefore in contact with the breast ductal epithelium (95). Despite this evidence, other epidemiological studies have found no correlation between serum P C B s and breast cancer risk (96-98): Another component of the endocrine system that is markedly affected by P C B s in vivo is the thyroid hormone system. Studies of rats treated with 0.2 to 1.8 mg/kg (99) and 5 to 50 mg/kg/day for 5 to 7 months (100) found that serum and plasma triiodothyronine (T3) and thyroxine (T4) levels were suppressed in a dose-related manner in neonatal, pregnant, and adult rats. This reduction results from direct thyroid damage, rather than an increase in hepatic catabolism (100). A s thyroid hormones are involved in fetal brain development, these observations hold important relevance regarding the role o f thyroid hormones in brain maturation and the developmental neurotoxicity induced by P C B s (99). Furthermore, P C B s have been shown to reversibly increase Sertoli cell proliferation and testis weight through suppression of T4 in rats treated with 1.6 and 3.2 mg/day during gestation (101). In addition to inducing hypothyroidism, 50 and 500 mg/kg of P C B s administered to rats translactationally as well as directly via the oral route produced ultrastructural lesions such as an increased development of rough endoplasmic reticulum and mitochondrial vacuolization in thyroid follicular cells. This contributed to 20 decreases in serum thyroid hormone levels (102). The decrease in serum thyroxine levels in this case is attributed to an interference with hormone secretion and an enhanced peripheral metabolism of thyroxine (103). Evidence of decreased serum levels of adrenal cortex hormones has also been reported in animal studies. Alterations in adrenal function including decreased serum corticosterone levels in rats treated orally with 1 to 50 mg/kg/day P C B s for at least 5 months has been reported (104). B y contrast, an increase in serum corticosterone levels after treatment with at least 8.1 mg/kg/day of P C B s for 2 weeks has been observed in mice (105). However, a study of monkeys exposed to 0.08 mg/kg/day o f P C B s for 22 months showed no change in serum hydrocortisone (106). These discrepancies may be due to species-specific differences in principal glucocorticoids and the effects of P C B s on the metabolism of glucocorticoids in each species (59). 1.3.1.1 Female Reproductive Effects P C B s were uterotropic when administered to immature female rats in doses of 30 and 120 mg/kg (107). P C B s at 0.1 to 10 fig/ml also decreased the in vitro fertilizing ability of exposed mouse oocytes (108). Monkeys treated with 20 and 80 ng/kg/day P C B s for about 2 years experienced changes in luteal phase progesterone levels and a marginally longer duration of menses (109). Neonatal female rats exposed to 110 umol/kg P C B s exhibited increased hepatic basal testosterone hydroxylase activity, androstenedione formation, and testosterone metabolism (110). Female offspring of rats exposed to P C B s as low as 8 ug/kg during lactation underwent a delay in puberty. When treated with doses between 32 and 64 ug/kg these rats had decreased uterine wet weight, and offspring in the 64 ug/kg group exhibited impaired fertility and irregular estrus cycle patterns (111). 21 Endocrine modulating effects related to estrogen have also been observed in humans. Women who consumed P C B contaminated fish had a significant reduction in menstrual cycle length (112). Another study found high levels around 790 ng/ml of some P C B congeners in the blood of women with repeated miscarriages which correlated with immunological and hormonal changes, such as a decrease in testosterone levels (r s=-0.23126) (113). 1.3.1.2 Male Reproductive Effects Exposure of sperm to a capacitation medium containing P C B s did not effect in vitro fertilizability or sperm motility in mice (108). However, male rats treated with 100 umol/kg P C B s neonatally exhibited decreased hepatic basal testosterone hydroxylase activity and androstenedione formation (110). The male offspring of rats dosed with 8, 32, and 64 mg/kg/day P C B s for 6 days during lactation developed significantly smaller prostates with fewer acini and altered morphology of the epithelial cells compared to controls in adulthood (114). Studies have also shown rats treated during lactation and/or in utero developed larger testes (101, 114). A reduction in seminal vesicle and epididymal weights, and caudal epididymal sperm counts has been reported (44) in rats treated during development. These studies indicate some of the potential reproductive target sites of P C B s in males; however, these findings are not consistent with other reports (44, 115). The effects of P C B s on male reproductive organs such as the prostate, epididymis, and testis are examined in this study. 1.4 The Prostate 1.4.1 Structure and Function The prostate is an exocrine gland that surrounds the urethra immediately below the bladder. This organ functions to produce secretions containing proteins and proteases 22 from the ducts in the prostate. Aliphatic polyamines such as spermidine and spermine are components of the secretions which are discharged into the urethra and may play a role in reducing the rate of seminal clot formation during ejaculation {116). In humans, the / -prostatic ducts originate from the urethra and radiate peripherally to completely surround the urethra. In rats and mice, the ducts are organized and encapsulated into individual lobes around the urethra (117). 1.4.2 The Androgen Receptor and Hormonal Regulation The androgen receptor (AR) is a member of the superfamily of nuclear receptors that act as transcription factors. Once the receptor binds to the ligand, this complex binds to D N A regulatory elements and activates gene transcription for specific genes. The receptors in this superfamily have structural and functional characteristics in common. The A R belongs to a subgroup of steroid hormone receptors that also includes the E R , G R , progesterone receptor (PR), and mineralocorticoid receptor ( M R ) (118). In the steroid receptor family, there are several conserved functional domains, such as a transactivating N-terminal domain (TAFi ) , a DNA-binding domain (DBD) , a hinge region involved in nuclear translocation and dimerization, and a C terminal ligand binding domain ( L B D ) which also has a transactivating function ( T A F 2 ) (119) (Figure 4). The prostate is an androgen target organ for its growth, development, and differentiation, and is responsive to androgens during fetal, pubertal, and adult stages (117). The secretory epithelial cells in the prostate require androgens, which are produced in the testis to maintain normal structure and function. The removal of androgens causes an inhibition of cellular proliferation and an induction of cell death mechanisms, which results in involution of the prostate (120). The development of 23 carcinoma of the prostate is an androgen dependent process (121). Prostate cancer is the most commonly diagnosed lethal cancer in Canadian men and has become the second greatest cause of cancer deaths in Canadian men (122). 1.4.3 Xenobiotic Effects on the Prostate There are a variety of components which contribute to the patterns of prostate cancer occurrence observed in epidemiological studies. These include age, ethnic background, country of residence, and diet (123). Genetic factors also contribute to the development of this disease. A n individual in North America with a first-degree relative who is affected has a 2 fold increased risk of developing prostate cancer (124). Genes linked to prostate cancer have also been found, including HPC1 on l q (125), and some genes on the X chromosome (126, 127). It has been proposed that high incidence of this disease in industrialized countries may be influenced by environmental factors (128). A s these environmental contaminants can interfere with normal androgen function, the changes that could occur in all stages of life may aid in the development and progression of prostate cancer. A statistically significant association was found between prostate cancer deaths and the number of acres that Canadian farmers sprayed with herbicides (129). This suggests a possible role of chemical exposure, perhaps through endocrine disrupting mechanisms in the etiology of adenocarcinoma of the prostate (130). The hormonal sensitivity of the prostate makes it vulnerable to endocrine disrupting effects, and the prostate is considered one of the most effected sites in the male reproductive endocrine system (130). This sensitivity is demonstrated in part by the influence of estradiol on prostate development in utero. Neonatal exposure to estradiol caused a decrease in A R expression in prostatic cells and significantly retarded growth 24 and epithelial cytodifferentiation, which demonstrates that basal epithelial cells are a possible target of androgen action during prostate morphogenesis (131). In addition, a low concentration of free serum estradiol during murine fetal life causes an increase in the number of developing prostatic glands during fetal life. In adulthood these mice had an increase in the number of prostatic A R s per cell and the prostate was enlarged due to hyperplasia. M i c e fetuses exposed to high concentrations of estradiol experienced a decrease in prostate weight in adulthood (25). Interestingly, murine fetal exposure to low doses of the xenoestrogen D E S caused an increase in prostate weight in adulthood (132). In addition, male rats exposed postnatally to P C B s developed histological alterations in the ventral prostate (133). These studies demonstrate the sensitivity of the prostate gland to hormones and xenobiotics during development and the possibility of long term effects such exposures can produce. 25 oo ON fe VO oo CN VO i n oo co m I "9 •» 3 & +-> *T3 fl> C » 1 op a o 0) Q CQ Q H .S 0 o 1 & 6 .a & M 2 < •a g Q CQ t i vo .s <^  o T3 0) O ID B 5= fe •H <T 00 s • s m •I 00 c .S -o o 13 C TJ S =5 g .2 o _> S fe 3 ^ 1 H .IP G O 3 3 26 1.5 Rationale and Objectives of this Study The influence of organochlorine environmental contaminants such as P C B s on biological systems can be severe and deleterious to all organisms in the ecosystem. Although P C B s are no longer manufactured and large releases of P C B s are prohibited in industrialized countries, due to their chemical properties and bioaccumulative nature, P C B s are still found in organisms at all trophic levels. These compounds are known to have endocrine disrupting capabilities. Most studies of the effects of P C B s on hormonal systems have focused on estrogen and thyroid hormone as targets. Indeed, endocrine disruption as a field of research has had great emphasis historically on estrogen mimics and environmental estrogens. Only a few studies have shown that environmental contaminants can block or increase the activity of androgens at the molecular level and affect reproductive development. The role of P C B s as androgenic endocrine disruptors has not been extensively studied. In order to further assess the impact of P C B s on the androgen axis, analysis of A R interactions and alterations in prostate growth and development are warranted. Hypothesis: The hypothesis to be tested is that P C B s have the ability to alter androgen regulated processes in vitro and in vivo. Objective 1. To test the influence of four different P C B mixtures, Aroclors 1242, 1248, 1254, and 1260 on androgen-regulated reporter gene expression in the L N C a P prostate cell line. Also to examine the influence of these chemicals on glucocorticoid-regulated reporter gene expression in order to demonstrate the presence of any androgen-specific effects. Objective 2. To determine the effects of two congeners, 2,3,4, trichlorobiphenyl (PCB 31) and 2,2',3,4' tetrachlorobiphenyl (PCB 42) in the same system 27 described above in order to gain insight into the action of individual congeners that are present at different concentrations in each mixture. Objective 3. To test the ability of these mixtures and congeners to interfere with binding of D H T (a natural A R ligand) in a cell line stably transfected with a FLAG-tagged A R . Objective 4. To evaluate the effects of P C B s on the prostate in the mouse model transgenic for a prostate-specific, androgen sensitive reporter gene. This animal model was used to assess the ability of the P C B mixture, Aroclor 1254, to interfere with specifically androgen regulated gene expression in vivo and the alterations in prostate size and development that occur as a result of P C B exposure. 28 CHAPTER 2. EXPERIMENTAL PROCEDURES 2.1 In vitro Studies 2.1.1 Chemicals P C B s were produced for industrial purposes and sold commercially as mixtures under a number of names, including Aroclor. These mixtures are composed of up to 209 congeners, although some congeners are only detectable at trace levels in some Aroclor mixtures. Four Aroclors, 1242, 1248, 1254, and 1260 were tested in this study. These four digit identifiers shows the parent molecule is a 12 carbon backbone by the first two digits in the identifier (12) and the last two digits indicate the overall chlorine content by weight percent of the entire mixture (90). The individual P C B congeners tested in this study were 2,3,4, trichlorobiphenyl ( I U P A C P C B 31) and 2,2',3,4' tetrachlorobiphenyl ( I U P A C P C B 42). These congeners are present in the Aroclor mixtures at different levels (Table 2.1). P C B 31 is 4.5% of Aroclor 1242, 9.3% of the content of Aroclor 1248, 0.7% of Aroclor 1254, and is not detectable in Aroclor 1260. P C B 42 is not detectable in Aroclor 1242, comprises 7% of Aroclor 1248, 2.2% of Aroclor 1254, and 0.7% of Aroclor 1260. Thus, these congeners represent a distinct component of each mixture, but exert their influence on Aroclor mixture action at different levels. Aroclor 1242 (mol wt.=266.5), Aroclor 1248 (mol wt.=299.5), Aroclor 1254 (mol wt.=328), Aroclor 1260 (mol wt.=375.7), P C B 31, and P C B 42 were purchased from Ultra Scientific (North Kingstown, RI). Compounds were dissolved in absolute ethanol (Aroclors) or hexane (PCB 31 and 42) for a 0.1 M stock. 29 Table 1. Percent content by weight of some individual congeners in four Aroclor mixtures. * indicates congeners tested in this study. Aroclor Congener 1242 1248 1254 1260 IUPAC # 31* 4.53 9.31 0.72 trace 42* trace 7.05 2.18 0.66 45 0.9 5.73 0.15 trace 66 0.81 4.95 2.24 0.22 99 0.55 2.52 6.1 0.82 101 0.27 1.5 6.98 5.04 118 trace trace 8.09 2 168 trace 0.56 4.23 0.59 193 trace trace 2.3 trace 198 trace trace 1 0.15 For cell culture studies serial dilutions were prepared in absolute ethanol with a final concentration of ethanol of 0.1% in culture. The final concentration of ethanol in the D H T and dexamethasone ( D E X ) (Steraloids Inc., Pawlings, N Y ) were dissolved in absolute ethanol for 1 mg/ml stocks. Serial dilutions of D H T and D E X were prepared fresh for each experiment. These dilutions were in 20% ethanol, for a final ethanol concentration of approximately 0.2% in culture. 2.1.2 P lasmid Constructs The cell culture transfection assays utilized four different plasmids. The first plasmid contained the full length c D N A of the rat A R driven by the C M V promoter; the plasmid has been named p A R 6 C M V (134). The rat G R plasmid was obtained from Dr. R. J. Matusik. This construct consists of a 2.8 kb insert containing 24 nucleotides upstream of the rat G R primary translation initiation site to 360 nucleotides downstream of the termination codon and is driven by the double R S V - S V 4 0 promoter; this plasmid is named p rGR (135). The reporter used was the ARR3tk-luc which contains three tandem repeats of -244 to -96 of the promoter region of the rat probasin gene (136). This promoter region contains androgen response elements to which both A R and G R bind to and transactivate from in vitro (in vivo, the probasin promoter is regulated by the A R endogenous gene only). The promoter is ligated in the pT81 vector (American Type Culture Collection, Rockvil le, M D ) , which contains a minimal thymidine kinase promoter, and the firefly luciferase gene, as described previously (134). The firefly luciferase gene is translated to a 61 k D a monomeric protein that does not undergo post-translational modification for 31 enzyme activity and is a genetic reporter that functions immediately upon translation (137, 138). A n additional plasmid was used as a transfection efficiency control. The p R L - T K transfection control vector contains a thymidine kinase promoter upstream of Rluc. Rluc is the slightly modified c D N A encoding Renilla luciferase from the sea pansy Renilla reniformis (Promega, Madison, Wisconsin). A l l plasmid D N A was propagated in JM109 E. coli and was prepared using Q I A G E N Maxiprep K i t ( Q I A G E N , Mississauga, O N , Canada). 2.1.3 C e l l Cul ture Transfections L N C a P cells originated as a lymph node metastasis from the prostate in a 50 year old Caucasian male. These cells are well differentiated androgen-sensitive prostate cells which express an endogenous mutated yet functional A R with an alanine to threonine mutation at position 868 (139). L N C a P cells were maintained in R P M I 1640 defined medium (GibcoBRL, Burlington, O N , Canada) supplemented with 5% fetal bovine serum under standard conditions (37°C, 5% CO2). For transfections, cells were cultured in R P M I 1640 defined medium supplemented with 5% hormone stripped (with dextran coated charcoal) fetal bovine serum (S-FBS) at a density of 3x 10 5 cells/well in 6 well cell culture plates (Costar, Coming N Y ) and incubated overnight under standard conditions. Plasmids were cotransfected into the cells using L I P O F E C T I N Reagent (GibcoBRL, Burlington, O N , Canada). Each plate received 1.5 pg of the A R or G R plasmid, 0.01 pg of pARR 3 t k - luc , and 0.005 pg of p R L - T K . The cells were incubated overnight and hormone and P C B were added simultaneously. Cells were harvested 48 h later in P B S with I m M E D T A and microcentrifuged at 3000 rpm at 4°C for 4 min. 32 Supernatant was removed and 0.1 ml Passive Lysis Buffer (Promega, Madison, WI) was added. Cells remained on ice for 15 min and were then gently vortex ed and frozen at -80°C until analysis. Luciferase assays were performed using the Dual-Luciferase Reporter Assay System (Promega, Madison, WI) and the E G & G Berthold Microplate Luminometer L B 96V. A n average of at least two experimental determinations with three replicates each were done for each P C B and hormone combination to give at least six samples per group. Firefly luciferase values were normalized for transfection efficiency using the activity of Renilla Luciferase as baseline levels. Values were presented as mean of total replicates (minimum of 6) corrected for background and control value ± standard error of the mean (SEM). 2.1.4 Cell Culture Ligand Displacement Assay H e L a cells expressing a stably transfected FLAG-tagged A R were obtained from Dr. M . Carey (140). The F L A G - A R fusion protein is an aspartate rich octamer at the N terminus of the A R . These cells were cultured in Dulbecco's Modif ied Eagle's Medium ( D M E M ) (Gibco B R L , Burlington, O N , Canada) supplemented with 10% S-FBS at 30 x 10 5 cells/plate in 10 c m 2 plates (Costar, Corning, N Y ) . Cells were incubated for 5 h in standard conditions. Media was changed to D M E M supplemented with 10% S-FBS, 0.3 n M [1,2,4,5,6,7-3H(N)]-DHT ( N E N Life Science Products, Inc., Boston, M A , specific activity=123 Ci /mol , initial concentration^ mCi/ml) in 20% ethanol and P C B in 100% ethanol. Cells were incubated 20 to 24 h and washed with P B S and harvested by scraping. Nuclear extracts (141) were prepared and frozen at -80°C prior to use. Briefly, harvested cells were centrifuged for 4 min at 4000 rpm. P B S was removed, and add 4 ml 33 cold buffer A (10 m M H E P E S p H 7.9, l O m M KC1, O . l m M E D T A , O . l m M E G T A , fresh I m M dithiothreitol, fresh 0 .5mM phenylmethylsulfonyl fluoride) vortexed carefully and set on ice 15 min. 250 p i of 10% I G E P A L was added, vortexed vigorously for 10 seconds and centrifuged immediately for 1 min at 4000 rpm. Supernatant was removed, and 500pl cold buffer C (20mM H E P E S p H 7.9, 0 .4M N a C l , I m M E D T A , I m M E G T A , fresh I m M dithiothreitol, fresh I m M phenylmethylsulfonyl fluoride) was added. Nuclear pellets were resuspended, vortexed vigorously for 15 min at 4°C, centrifuged for 5 min at 4000 rpm at 4°C and frozen. Extracts were thawed and half of the extract was incubated for 6 h with 10 p i of agarose beads conjugated with a n t i - F L A G monoclonal antibodies (Sigma Chemical Co., St-Louis, M O ) at 4°C for 6 h. Beads were washed 2 times with Buffer D (140) containing 2 0 m M H E P E S p H 7.9, 20% glycerol, 0 .3M K C 1 , 0 .2mM E D T A , 0.05% I G E P A L , 0 .5mM dithiothreitol, 0 .5mM phenylmethylsulfonyl fluoride, and then resuspended in buffer D . Resuspensions were measured for levels of tritium by scintillation counting in a Beckman L S 6500 scintillation counter. Western Blotting was performed according to the method of Harlow and Lane, (142) using the remaining portion of the extracts. Thawed nuclear extracts were heated at 100°C for 5 minutes in 2 X sample buffer (0.125M Tr i s -HCl p H 6.8, 4% SDS, 20% glycerol, 5% mercaptoethanol), and put on ice and then loaded into gels. Samples were electrophoresed on 10% S D S - P A G E minigels and transferred to P V D F membranes overnight. After transfer, membranes were washed 2 X for 5 minutes in T B S T (lOOmM Tr i s -HCl p H 7.5, 0.9% N a C l , 0.1% Tween-20) and blocked with 10% dried skim milk in T B S T for 1 hour at room temperature. The primary antibody, an anti-mouse monoclonal antibody to human A R D B D (Pharmingen, Mississauga, O N , Canada) was added to the 3 4 blots in T B S T (1:500) for 2 hours. Membranes were washed 3 times for 10 minutes with T B S T . Following the washes, membranes were incubated with anit-mouse-HRP (Santa Cruz Biotechnology, Santa Cruz, C A ) at a dilution of 1:5000. Membrane bound antibodies were detected using chemiluminescence with E C L ™ Western blotting detection agents (Amersham Pharmacia Biotech Inc., Baie d'Urfe, Quebec) for an incubation period of 30 to 60 s, and exposed immediately to scientific imaging film (Kodak Eastman Co. , Rochester, N Y ) . Films from Western blots were analyzed using BioRad Quantity One Quantitation Software (BioRad, Hercules, C A ) to determine and compare band densities for nuclear extracts from each ligand displacement assay. Values in this assay are presented as the amount of 3 H - D H T ( D P M ) bound to the A R and corrected for the amount of A R present in the sample (dividing by the value calculated as the density of the band after Western blotting). 2.2 In vivo Studies 2.2.1 Animals This study used CD1 transgenic mice possessing the androgen sensitive L B P -C A T reporter gene (143). This transgene is comprised of a large fragment of the rat probasin promoter (-11500 to +28 bp) (LPB) linked to a chloramphenicol acetyl transferase ( C A T ) gene to achieve highly prostate-specific expression of C A T in the prostatic secretory epithelial cells. Expression is highest in the ventral prostate, and therefore, this lobe of the prostate was used for biochemical assays. Sexually mature L P B - C A T mice were combined in mating pairs and females were treated daily by gavage. Four to five pairs were used for each treatment group to ensure a minimum of five male pups per group. Females were checked daily for copulatory plugs. Males were removed on approximately day 18 of gestation to prevent re-impregnation after birth of 35 the litter. Females, litters, and male weanlings were weighed daily. A l l mice were fed Lab Diet 5015 (PMI Foods, Richmond, IN) and water ad libitum and were kept on a cycle of 12 hours light followed by 12 hours darkness at 20-25°C. These animal studies were conducted according to the Canadian Council o f Animal Care Guidelines and were reviewed by the University of British Columbia Animal Care Committee. 2.2.2 Treatment and Groupings Treatments used were 10 mg/kg/day Aroclor 1254 (Ultra Scientific, North Kingstown, RI) in 0.1 m l canola oi l , and 0.1 m l canola o i l (Safeway Brand, Safeway Canada) for the control in the single dose level group. The dose response group received 10, 20, or 40 mg/kg/day Aroclor 1254 in 0.2 m l canola o i l and 0.2 m l canola o i l (Ultra Scientific) for the control group. Females were dosed until weaning when the pups were 21 days old. Weaned male pups from the litter were treated via gavage daily until the ages of 28 days old (single dose level group only) and 56 days old. A t the end of the treatment the animals were sacrificed by cardiac exsanguination under Metofane anesthesia (Janssen, Toronto, O N , Canada). The left epididymis, left testis, and left kidney were removed, weighed and frozen. Organ weights were calculated as organ weight/body weight in order to correct for mouse to mouse variation in body weight. This data is presented as group mean ± S E M . The right epididymis and right testis were removed and fixed in 10% neutral buffered formalin (NBF) for histology. The liver, heart, and ventral prostate were removed, weighed, subdivided and half frozen. The remaining half was fixed in 10% N B F . The doses were selected based on the solubility limits of Aroclor 1254 in canola oi l of 1 mg/ml, and on the basis of toxic levels reported for the liver, as P C B s are known 36 to cause severe hepatic effects. A dose of 50 mg/kg/day of P C B s administered to rats for 30 days resulted in degenerative histological liver changes (144). A dose of 2.5 mg/kg/day of P C B s administered to rats for 5 months did not cause any significant changes in liver weights or histology (104). Therefore, doses used in our study were within this range of level and duration of treatment. 2.2.3 CAT Assay Frozen tissues were thawed on ice and homogenized in 300 ul lysis buffer (0.1M Tris, p H 7.8, 0.1% Triton X-200) with 10 strokes of a Pyrex Tenbroeck Glass homogenizer. Extracts were removed from homogenizer and centrifuged at 8000 rpm for 3 minutes at 4°C. Supernatant was assayed for protein concentration using the Pierce B C A protein assay kit (Rockford, Illinois) and results read at 560 nm in a Titertek Multiskan platereader (Flow Laboratories, Mississauga, O N , Canada). Aliquots containing 0.1 to 1 ug of protein from each sample in duplicates in 200 ul lysis buffer were heated at 65°C for 10 minutes and cooled at room temperature for 10 minutes. A reaction mixture combining 3Ff Acetyl-Co A , 5 m M Chloramphenicol, and I 'M Tris p H 7.8 was added to the sample. 3 m L of ScintiLene (Fisher Scientific, Whitby, O N , Canada) was added and incubated at room temperature for 30 minutes. Samples were read in a scintillation counter for 1 minute per sample for 5 cycles. Slope of the line from the five cycles was calculated and converted to dpm/min/mg protein for the C A T activity. C A T activity was presented as mean ± S E M . 2.2.4 Histology Fixed tissues were paraffin embedded and slides prepared at V G H Pathology Laboratories. A l l slides were H and E stained. Sections were then assessed for 37 histological anomalies by a veterinary pathologist (Stephane Lair). A photonic microscope was used. The parameters examined include evaluation of different histological structures and grading of the distribution of such structures in each section. Other histological findings such as anomalies in the tissues were also reported and graded. 2.2.4.1 Measurement of Percent Area of Stroma and Ducts in Prostate Tissue Sections Three photographs of random areas of stained slides of prostate tissue from each experimental animal were captured using C o o l S N A P software connected to an inverted microscope (Leica, Wetzlar, Germany). Each captured image was exported to Adobe Photoshop 5.5. Drawing tools in this program were used to cover all the ductal structure in the image with the colour red. The remaining area in the image was the stroma. This area was coloured white. The number of red and white pixels relative to the total number of pixels was calculated in Photoshop to give a representative value of the percent area for stroma and ducts. A ratio of stromal to ductal area was also calculated based on the number of pixels assigned to each type of area. The percent area for each image was averaged with others for the same treatment group for comparison. This calculation was used to determine i f the treatment related increase in dilated acini was due to a change in the stromal or ductal area, which may account for the decrease in prostate weight with increasing treatment dose. 2.2.5 Measurement of Serum Testosterone Levels Serum collected by cardiac exsanguination was frozen at - 2 0 ° C . Samples from the single dose level group that were in excess of 200 pi were sent to the Tumour Marker 38 Laboratory at the British Columbia Cancer Agency for total testosterone radioimmunoassay (DPC, Los Angeles, C A ) . The sensitivity of this assay is at 0.14 nmol/1, and the upper limit of the range of detection is 56 nmol/1. Crossreactivity with D H T is less than 5% and the antiserum is considered highly specific for T. Results were provided in duplicate in nmol/1, and were presented in a scatter plot by treatment group. 2.2.6 Statistics Statistical analysis for results from the 10 mg/kg/day single dose level group C A T assay and organ weights were analyzed using the Student's Mest. The one way analysis of variance ( A N O V A ) and Dunnett's test for multiple comparisons of means were used for the dose response study and the cell culture ligand displacement assay. p< 0.05 was considered significant. The serum testosterone levels data set was analyzed by one way A N O V A as well as by the Kruskal-Wallis test using IMP® (SAS Institute Inc.). 39 C H A P T E R 3. RESULTS 3.1 Induction of Luciferase Reporter Gene A n endocrine disruptor may act as an A R agonist by directly activating androgen receptor function in the absence of hormone, or D H T . In the presence of D H T , there is competition for the receptor ligand binding sites between D H T and the endocrine disruptor, which may result in an antagonistic effect. These same mechanisms for changes in receptor activation may also be seen with the G R and D E X . This cell culture assay of P C B alteration of steroid hormone action involved transient transfection of L N C a P cells, a well differentiated prostate cancer cell line. Cells were transiently cotransfected with plasmids containing the A R or G R , and the ARR.3t.k-luc hormone response elements to which the D N A binding domain of both A R and G R bind. p R L - T K was also cotransfected into the cells. This plasmid is transcribed relative to cell proliferation and was therefore an internal control for this assay. The addition of a ligand such as D H T or D E X induced transcription of the luciferase gene from ARR^tk-luc. Addit ion of P C B s can alter such transcription i f these compounds have the ability to influence A R or G R activity. 3.1.1 Induction by D H T and D E X To determine i f in vitro effects on the luciferase reporter gene assay are specific to a certain steroid hormone receptor, both A R and G R were tested. L N C a P cells transfected with the A R and exposed to a titration of D H T concentrations showed a steep increase in luciferase activity from 0.01 to 0.1 n M followed by a leveling off of activity 40 [DHT], nM b. 120 . [DEX], nM Figure 5. L N C a P cells transfected with a) A R and exposed to D H T titration and b) G R and exposed to D E X titration. Concentration of D H T (a) and D E X (b) in n M is plotted on the x axis, and the percent fold activation with the maximal activation level set at 100% plotted on the y axis. Luciferase activity was normalized relative to Renil la luciferase activity, and the maximal activity was set at 100% fold activation (not shown on log graph). Data expressed as the mean fold activation ± S E M . 41 from 0.1 to 10 n M (Figure 5a). The half-maximal level of peak luciferase activity was at 0.05 n M , and the peak level above 0.1 n M . Therefore, transfected cells were incubated with 0.05 and 1 n M D H T to demonstrate effects in the presence of half maximal and saturating levels of D H T . Cells transfected with the G R and exposed to a titration of D E X concentrations showed low levels of induction from 0.01 to 0.5 n M . Concentrations above 0.5 n M showed a steep increase in D E X induced luciferase activity until 10 n M , although a level of saturation was not reached. The sub-maximal level of luciferase activity from the range tested for D E X was at 1 n M (Figure 5b). Therefore, concentrations of hormone inducing half, sub, and maximal levels were used in combination with Aroclor 1242, Aroclor 1248, Aroclor 1254, Aroclor 1260, P C B 31, and P C B 42 titrations to detect changes in androgen and glucocorticoid driven luciferase activity. In order to measure for A R or G R agonist activity of P C B s , cells transfected with A R and G R were also exposed to titrations of the four Aroclors and two individual congeners in the absence of the known ligands. 3.1.2 Induction of Luciferase Activity Through A R by Aroclor 1254 In the absence of D H T , Aroclor 1254 (1 to 10 nM) did not induce an increase in luciferase activity. A slight increase was observed at 100 n M , followed by a marked increase in luciferase activity above 100 n M (Figure 6a). In the presence of 0.05 n M D H T there was no change in luciferase activity from 1 to 100 n M , but there was a pronounced antagonistic effect at 1000 and 10000 n M where luciferase activity is reduced by 66% and 93 % of the control value, respectively. (Figure 6b). A similar pattern of induction was found in combination with I n M D H T , where there was a gradual 42 120 A 0 -I , 1 , , 1 10 100 1000 10000 [Aroclor 1254], nM 0 A , , , , 1 10 100 1000 10000 [Aroclor 1254], nM 0 H 1 1 1 1 1 10 100 . 1000 10000 [Aroclor 1254], nM Figure 6. L N C a P cells transfected with A R and exposed to a) Aroclor 1254 titration + no. D H T b) Aroclor 1254 titration + 0.05 n M D H T c) Aroclor 1254 titration + 1 n M D H T . Cells transfected with G R and exposed to d) Aroclor 1254 titration + no D E X e) Aroclor 1254 titration + I n M D E X . Concentration of Aroclor 1254 in n M (b) + 0.05 n M D H T c) + 1 n M D H T e) + I n M D E X ) is plotted on the x axis, and the percent fold activation with maximal activation level set at 100% (a) and the percent fold activation with the control activation level set at 100% (b to e) plotted on the y axis. Luciferase activity was normalized relative to Renilla luciferase activity, and the maximal activity was set at 100% fold activation for assays without D H T , and control activity (activity before D H T -added) was set at 100% fold activation for assays with D H T (not shown on log graph). Data expressed as the mean fold activation ± S E M . 44 decrease in luciferase activity from 1 to 10000 n M , with a reduction by 64% of control activity (Figure 6c). These changes suggest that Aroclor 1254 acts as a weak androgen agonist in vitro. Aroclor 1254 did not alter luciferase activity either with or without D E X when G R was used to drive the reporter construct. In the absence of D E X , the luciferase induction levels are negative in cells exposed to Aroclor 1254, compared to the levels in the control where no P C B is added (Figure 6d). In the presence of 1 n M D E X , there are no distinct patterns in the trend of the graph (Figure 6e). A 40% drop relative to the control value in luciferase activity is seen at 100 n M , which increases 30% at 5000 n M and drops 30% at 10000 n M . The lack of change in activation of luciferase gene transcription by Aroclor 1254 in cells transfected with G R suggests specificity of Aroclor 1254 for the A R . 3.1.3 Alteration of Luciferase Activity Induction Through AR by Aroclor 1242 Aroclor 1242 did not have a significant effect on luciferase activity with A R in the absence of hormone. In the presence of 0.05 n M D H T there was an antagonistic effect demonstrated by a slight decrease in luciferase activity from 1 to 100 n M , followed by a distinct decrease in luciferase induction above 100 n M with a 98% reduction in luciferase activity at 10000 n M (Figure 7a). There was a similar effect in the presence of I n M D H T (Figure 7b) where there was no change in luciferase activity from 1 to 100 n M Aroclor 1242, a slight decrease in activity at 1000 n M followed by a steep decrease at 10000 n M by 92%. There was no effect observed with G R in the absence of hormone or in the presence of D E X . 45 10 100 1000 [Aroclor 1242], nM 10000 160 n 10 100 1000 [Aroclor 1242], nM 10000 Figure 7. L N C a P cells transfected with A R and exposed to a) Aroclor 1242 + 0.05 n M D H T b) Aroclor 1242 + 1 n M D H T . Concentration of Aroclor 1242 in n M (a)+ 0.05 n M D H T b) + I n M D H T ) is plotted on the x axis, and the percent fold activation with the control activation level set at 100% plotted on the y axis. Luciferase activity was normalized relative to Renilla luciferase activity, and control activity (activity before D H T added) was set at 100% fold activation (not shown on log graph). Data expressed as the mean fold activation ± S E M . 46 120 a. [Aroclor 1248], nM 0 -I , , , , 1 10 100 1000 10000 [Aroclor 1248], nM Figure 8. L N C a P cells transfected with A R and exposed to a) Aroclor 1248 + 0.05 n M D H T b) Aroclor 1248 + 1 n M D H T . Concentration of Aroclor 1248 in n M (a)+ 0.05 n M D H T b) + 1 n M D H T ) is plotted on the x axis, and the percent fold activation with the control activation level set at 100% plotted on the y axis. Luciferase activity was normalized relative to Renilla luciferase activity, and control activity (activity before D H T added) was set at 100% fold activation (not shown on log graph). Data expressed as the mean fold activation ± S E M . 47 3.1.4 Alteration of Luciferase Activity Induction Through AR by Aroclor 1248 Aroclor 1248 did not have a significant effect on luciferase activity with A R in the absence of hormone. In the presence of a half maximal level of D H T , there was no change in luciferase activity at 1 or 10 n M , but a strong antagonistic effect above 10 n M Aroclor 1248 was seen with a distinct decrease in activity at 100 n M (Figure 8a). This was followed by a further shallow decrease in activity from 1000 to 10000 n M Aroclor 1248 reaching a 99% decline in luciferase activity. A less striking effect was seen when a saturating level of D H T was used (Figure 8b). In this case, there was no decrease in activity from 1 to 10 n M . However, a slight decrease in luciferase activity was seen at 10, 100, and 1000 n M and a sharp decrease at 10000 n M at which point the luciferase activity had diminished by 69%. A change in luciferase activity was not observed with G R in the absence of hormone or the presence of D E X . 3.1.5 Alteration of Luciferase Activity Induction Through AR by Aroclor 1260 Aroclor 1260 did not have a marked effect on luciferase activity with A R in the absence of hormone. In the presence of 0.05 n M D H T there was an antagonistic effect shown by a slight decrease in activity at 1000 n M and a sharp decrease at 10000 n M by 88%o (Figure 9a). In the presence of saturating levels of D H T , Aroclor 1260 does not cause a distinct change in luciferase activity at 1 n M or 10 n M . Aroclor 1260 was a strong antagonist above 10 n M Aroclor 1260, as shown by an abrupt decline in luciferase activity at 100 n M that remains reduced by approximately 60% for all higher 48 120 0 -I , , , , 1 10 100 1000 10000 [Aroclor 1260], nM Figure 9. L N C a P cells transfected with A R and exposed to a) Aroclor 1260 + 0.05 n M D H T b) Aroclor 1260 + 1 n M D H T . Concentration of Aroclor 1260 in n M (a)+ 0.05 n M D H T b) + 1 n M D H T ) is plotted on the x axis, and the percent fold activation with the control activation level set at 100% plotted on the y axis. Luciferase activity was normalized relative to Renilla luciferase activity, and control activity (activity before D H T added) was set at 100% fold activation (not shown on log graph). Data expressed as the mean fold activation ± S E M . 49 concentrations (Figure 9b). There was not a discernible alteration of luciferase activity with G R in the absence of hormone or the presence of D E X . 3.1.6 Alteration of Luciferase Activity Induction Through AR and GR by PCB 42 P C B 42 did not have a detectable effect on luciferase activity with A R in the absence of hormone. P C B 42 had a profound antagonsitic effect with A R in the presence of 0.05 n M D H T as shown by a marked decrease in luciferase activity at 1000 n M P C B 42 (Figure 10a). This is followed by a less steep but certainly notable decrease in activity by approximately 98%. In the presence of 1 n M D H T there is no distinct change in activity observed from 1 to 1000 n M . However there is a very abrupt decrease in activity by 93% at 5000 and 10000 n M P C B 42 (Figure 10b). Using G R , in the absence of D E X there was no effect on luciferase activity. However, when 1 n M D E X was added to the cells, P C B 42 caused a sharp increase in luciferase activity from 1 to 100 n M , followed by a leveling off of luciferase activity at higher concentrations with a 150% increase (Figure 10c). 3.1.7 Alteration of Luciferase Activity Induction Through AR by PCB 31 P C B 31 had an antagonistic effect on luciferase activity with A R in the presence of 0.05 n M D H T . There was no change in activity from to 100 n M , however, there was a moderately distinct decrease by 60% in luciferase activity at higher concentrations (Figure 11). There was no effect on luciferase activity when I n M D H T or D E X was used. 50 0) 40 A 100 [PCB 42], nM 10000 120 -, b. [PCB 42], nM 0 -I , , , , 1 10 100 1000 10000 [PCB 42], nM Figure 10. L N C a P cells transfected with A R and exposed to a) P C B 42 + 0.05 n M D H T b) P C B 4 2 + 1 n M D H T , and transfected with G R and exposed to c) P C B 42 + 1 n M D E X . Concentration of P C B 42 in n M (a)+ 0.05 n M D H T b) + 1 n M D H T ) c) + 1 n M D E X is plotted on the x axis, and the percent fold activation with the control activation level set at 100% plotted on the y axis. Luciferase activity was normalized relative to Renilla luciferase activity, and control activity (activity before hormone added) was set at 100% fold activation (not shown on log graph). Data expressed as the mean fold activation ± S E M . 120 > ti ro o > 80 40 0 -1 , , , , 1 10 100 1000 10000 [PCB 31], nM Figure 11. L N C a P cells transfected with A R and exposed to P C B 31 + 0.05 n M D H T . Concentration of P C B 31 in n M + 0.05 n M D H T is plotted on the x axis, and the percent fold activation with the control activation level set at 100% plotted on the y axis. Luciferase activity was normalized relative to Renilla luciferase activity, and control activity (activity before D H T added) was set at 100% fold activation (not shown on log graph). Data expressed as the mean fold activation ± S E M . 52 3.2 Displacement of 3 H - D H T as A R L i g a n d by P C B s There are a number of mechanisms by which xenobiotics can alter steroid hormone action. One of these well established mechanisms is through modification of steroid hormone binding to the receptor. In order to determine i f the androgen receptor is a target for P C B action, a cell culture assay using a H e L a cell line with a stably transfected FLAG-tagged A R was employed (140). Cells were exposed to 3 H - D H T (specific activity =123 Cii /mmol), a known ligand of A R , in combination with an excess of each of the P C B s tested in the luciferase reporter gene assay. Once the A R binds to ligand, this complex translocates to the nucleus. Therefore, extraction of the nuclear fraction of these cells shows the location of the 3 H - D H T , and demonstrates the ability of the Aroclors and individual congeners to interfere with the some of the processes involved in androgen action, including ligand binding to the A R . These experiments were designed to provide insight into the complex mechanisms of P C B interference with steroid hormones, which has not been previously extensively examined at the cellular or molecular level in this manner. Figure 12 depicts the level of 3 H - D H T binding when a) unlabelled D H T is added as competitor, and b) Aroclor 1254, Aroclor 1242, Aroclor 1248, Aroclor 1260, P C B 42, and P C B 31 are competitors of 0.3 n M 3 H - D H T . There was a profound decrease in bound 3 H - D H T at 10000 n M for all P C B s except for P C B 31, which did not affect the amount of 3 H - D H T bound to the A R . A 50% decrease was observed at 1000 n M with Aroclor 1242 and P C B 42, and a 25% decrease at 1000 n M with Aroclor 1248, 1254 and 1260. A t this concentration, only P C B 42 caused a significant (p<0.05) reduction in bound 3 H - D H T . At 10000 n M , a 90-95% decrease in bound D H T was observed for Aroclor 1242, Aroclor 1254, and P C B 42, and 53 a 60% decrease for Aroclor 1248 and Aroclor 1260. P C B 42 and Aroclor 1254 reduced the bound 3 H-DHT significantly (p<0.05). 54 Figure 12. H e L a FLAG-tagged A R cells exposed to a) D H T titration + 0.3 n M 3 H - D H T and b) P C B titrations + 0.3 n M 3H-DHT. Bound 3 H - D H T normalized relative to A R concentration based on quantitative analysis of Western blotting, and control activity was set at 100% bound (not shown on log graph). Data expressed as the mean bound 3 H -D H T . * indicates statistical significance (p< 0.05). 55 3.3 Analysis of C A T Act iv i ty The in vitro cell culture studies demonstrated that P C B s have properties of potential endocrine disruptors. More specifically, the P C B mixture Aroclor 1254 was able to induce reporter gene activity through the A R in the absence of known ligand and also reduced reporter gene activity in the presence of the ligand D H T . This novel finding of an alteration in AR-driven activity was particularly notable above 100 n M Aroclor 1254 in the presence of half maximal levels of D H T . In addition, the ligand displacement assay showed that 10000 n M Aroclor 1254 prevents 90% of 3 H - D H T from being associated with the A R . These results, as well as reports of the ability of Aroclors to alter steroid hormone systems and reproductive development, warranted an examination of ih vivo effects of Aroclor 1254, particularly on the prostate, a sensitive target organ of androgens. In order to measure the influence of Aroclor 1254 on prostate development, the expression of the androgen-driven prostate specific L P B - C A T transgene and the effects on mouse organ weights were investigated. The organs examined were androgen-target and reproduction related organs such as testes, epididymis, and prostate; as well as liver, an established target for P C B effects as well as the site of steroid hormone metabolism. P C B s are known to bioaccumulate and are stored preferentially in adipose tissue, but are also present in serum, blood plasma, and milk (145). Consequently, P C B s can be transferred to offspring via the placenta and through breast milk (48, 146). Therefore, the offspring of exposed mothers, who are exposed to P C B s at early and sensitive stages of life can be more susceptible to the effects of P C B s than those exposed during adulthood. 56 Two studies were undertaken to evaluate the effects of Aroclor 1254 exposure during development. These studies were designed to look at chronic exposure to P C B s through different stages of development, rather than acute exposure which is tested in many toxicology studies. In the single dose level study, dams were dosed with 10 mg/kg/day from the time of mating, until the end of the lactation period at 3 weeks. Weanlings were then administered 10 mg/kg/day Aroclor 1254 until four weeks of age (prepubertal) and eight weeks of age (post pubertal), in order to assess the effects of Aroclor 1254 at different stages of development. In the dose response study, dams were treated with 10, 20, 40 mg/kg/day Aroclor 1254 from the time of mating, until the end of the lactation period at 3 weeks. Weanlings were then administered the corresponding dose until 8 weeks of age, to test the effects of different doses of Aroclor 1254 on the prostate. 3.3.1 Single dose level Group There was a reduction in C A T activity from the prostate extracts of the four and eight week treatment mice, which was statistically significant (p>0.05) in the eight week group (Figure 13). These findings show antiandrogenic action of the Aroclor 1254 treatment on the specifically androgen responsive L P B - C A T reporter. Due to this effect at a relatively low dose of 10 mg/kg/day, we chose to look at mice at eight weeks of age, treated from conception in the same manner, but also using treatments 2 and 4 fold higher. 57 Figure 13. Activi ty of C A T reporter transgene linked to the prostate specific probasin promoter from prostate tissue extracts of four and eight week old control (open bars) and treatment (closed bars) L P B - C A T mice in the single dose level study. M i c e were treated with 10 mg/kg/day Aroclor 1254 by gavage from conception until four or eight weeks of age. Data is presented as mean C A T activity. Error bars represent S E M . *p<0.05 (Student's f-test). 58 3.3.2 Dose Response Group The C A T enzyme activities in the dose response group did not show any significant difference between groups (p<0.05), despite the trend of a decrease in activity at higher doses. The activities of the prostate extracts from the control group were also remarkably lower than those in the single dose level study, and were lower than the 10, 20, or 40 mg/kg/day group C A T activities (Figure 14). 3.4 Alteration of Organ Weights In order to assess the effects of P C B s on the mice used in this study, changes in body weight and organ weight were useful parameters that were employed. Body weights of mothers being dosed, suckling pups, and weanlings were collected throughout both studies. Due to litter to litter differences, the treatment mice in the single dose level group were heavier than the control group. A n increase in body weight is not characteristic of P C B exposure, and in fact, a decrease in body weight has been reported as a health effect of these compounds (59). The differences in body weight in the mice in this study may be attributable to litter size. Some of the treatment mice were from smaller litters, which can yield larger pups. However, there was no trend in litter size observed when treatment and control groups were compared. This pattern was not seen in the dose response group, where there were no measurable differences seen between dose groups. Organ weight was corrected for individual body weight, due to the variation in body weight between individual animals in this study. 59 30 -j 25 -to © Control Low Medium High Figure 14. Activi ty of C A T reporter transgene linked to the prostate specific probasin promoter from prostate tissue extracts of eight week old L P B - C A T mice in the dose response study treated with vehicle only (control), 10, 20, and 40 mg/kg/day. M i c e were treated with 10, 20, 40 mg/kg/day Aroclor 1254 by gavage from conception until eight weeks of age. Data is presented as mean C A T activity. Error bars represent S E M . 60 3.4.1 L i v e r To determine i f there was an effect on the liver, a known target organ of P C B s , we examined the livers of treated mice for changes in weight, although the doses used were below levels shown to cause overt effects in other studies (144). There were no significant differences in liver weight/body weight between control and treatment mice in the single dose level study (Table 2). Liver weight/body weight in the dose response treatment mice were significantly higher for the 20 and 40 mg/kg/day dose groups than controls (Table 3). This is consistent with reports in the literature of P C B liver effects (44, 73, 74, 76-78). Increase in liver weights due to P C B exposure is usually associated with hepatocyte enlargement, which is not considered an adverse effect, unless certain biochemical changes along with notable histological alterations are also observed (59). 3.4.2 P r o s t a t e The prostates in the single dose level group did not show a significant difference in prostate weight to body weight ratio between treatment and controls for the single dose level group (Table 2). In the dose response study, prostate weight/body weight was altered by Aroclor 1254 treatment. Prostate weight/body weight for the 20 and 40 mg/kg/day dose groups were significantly lower than the control group (Table 3). 61 C O N T R O L (N=12) T R E A T M E N T (N=5) Prostate 2.99±0.20 2.7±0.25 Epididymis 7.69±0.26 7.2±0.22 Four Weeks Testis 27.15±1.38 28.4±0.70 Heart 51.84±0.94 49.16±3.3 Liver 537.42±5.25 625.66±11.49** Kidney 69.12±1.76 65.16±1.98 C O N T R O L ( N = l l ) T R E A T M E N T (N=8) Prostate 4.38±0.27 3.99±0.29 Epididymis 13.85±0.28 13.16±0.70 Eight Weeks Testis 42.17±2.12 43.87±1.81 Heart 50.58±1.40 46.61±2.14 Liver 494.10±15.00 540.05±21.88 Kidney 72.38±2.25 72.86±3.21 *p<0.05, **p<0.01 (Student's Mest) Table 2. Corrected organ weights for prostate, left epdidymis, left testis, heart, liver, and kidney (organ weight/body weight in grams/gram body weight x l O 4 ) o f L P B - C A T mice treated to four and eight weeks of age with vehicle only (canola oil) and 10 mg/kg/day Aroclor 1254. Values are m e a n ± S E M . C O N T R O L 10 M G / K G / D A Y 20 M G / K G / D A Y 40 M G / K G / D A Y (N=9) (N=5) ( N = l l ) (N=5) Prostate 4.36±0.25 3.91±0.23 3.50±0.18* 3.20±0.27** Epididymis 13.63±0.21 12.41±0.62 13.33±0.39 13.03±0.45 Testis 46.92±2.22 41.7U3.79 48.79±1.09 47.86±1.11 Heart 51 .30i l .27 47.21±0.80 49.05±0.38 51.48i2.41 Liver 496.53±9.67 536.39±14.54 617.58±13.52** 718.08±24.26** Kidney 74.72±2.44 75.63±2.59 75.67±1.70 78.30±2.72 *p<0.05 ,**p<0.01(analysis of variance and Dunnett's test) Table 3. Corrected (organ weight/body weight in grams/gram body weight x l O 4 ) for mice treated for eight weeks of age with canola oi l , Aroclor 1254 at 10, 20, 40 mg/kg/day. Values are mean ± S E M . 62 3.4.3 Serum Testosterone Levels The level of testosterone in the serum may provide insight into the effects of antiandrogens on the endocrine system. This has been seen in studies of antiandrogen drugs used in prostate cancer treatment (147-149), as well as in the treatment of aggression in convicted male sexual offenders (132). Patients treated with the antiandrogen flutamide, develop an increase in serum testosterone levels due to the inhibition of the negative feedback effects of androgens on the hypothalamic-pituitary-testes axis (150). Serum lutenizing hormone (LH) is also increased in a dose related manner with flutamide administration (151), and it is through negative feedback control by endogenous androgens of gonadotropin secretion in the hypothalamus that L H secretion is regulated (152). There was a high degree of variation in testosterone levels within treatment groups. However, this is consistent with other studies in which the measured serum testosterone levels ranged from 0.7 nmol/1 to 158 nmol/1 in post pubertal mice (153) ( M . Bowden, personal communication). Variations in serum testosterone levels in mice may be attributed to the release of L H from the pituitary in pulses (154). A s testosterone synthesis is dependent on L H , these fluctuations may be expected and data was therefore presented as a scatter plot of testosterone concentrations (Figure 15). There were three serum testosterone levels that were much higher than the other fifteen samples measured. Statistical analysis of the data set was done both with and without these three points which may have skewed the results. A N O V A tests of the data with and without the higher testosterone levels both showed no significant difference between treatment groups (p<0.05). The Kruskal-Wallis test was also used, as this test ranks the data and 63 40 35 ~ 30 o E c o (A O (A 0) 25 20 -15 -10 -5 0 • Control • 10 mg /kg /day A 20 mg /kg /day • 40 mg /kg /day 10 20 30 Aroclor 1254 Dose 40 50 Figure 15. Scatter plot of serum testosterone levels from eight week old L P B - C A T mice in the dose response study treated with vehicle only (control), 10, 20, 40 mg/kg/day. 64 uses the median of each group, rather than the mean and thus a valid result could be obtained and the inclusion of the three high testosterone levels would not be problematic to analysis. This test also found no statistical difference between treatment groups (p<0.05). Compromised liver function is a possibility when agents with known hepatic effects are used. Impaired hepatic function or inflammation with an increase in liver enzymes may reduce the liver's ability to remove aromatized testosterone metabolites from the serum. These metabolites are weakly estrogenic and reduce testosterone levels by negative feedback on the pituitary. Although a wide range of testosterone levels was observed, the testosterone levels are within a normal range (153). The results from the testosterone assay are not conclusive regarding the effects of Aroclor 1254 on liver function and hormone-metabolizing capacity; however, this does not negate the other notable findings in this set of studies. 3.5 Pathological Analysis of Histological Sections 3.5.1 Liver Histological analysis of livers of mice from each treatment group in the dose response study were performed to confirm that the alteration in liver weight was not a result of severe toxicological changes that might affect liver function. The livers of Aroclor 1254 treated mice showed changes in the distribution of hepatocellular vacuolization. Control and 10 mg/kg/day treated mice showed no vacuolization of hepatocytes around the centrilobular veins, whereas 9% of the medium dose mice and 80% of the high dose mice livers had centrilobular vacuolization (Figure 16). Control mice had diffuse or no obvious regional distribution of vacuoles in 89% (8 of 9) of the cases examined. L o w dose mice had 40%(2 of 5), medium dose mice had 9% (1 of 11), 65 Figure 16. Centrilobular vacuolization in the liver of an eight week old mouse treated with 40 mg/kg/day of Aroclor 1254. The hepatocytes surrounding the centrilobular veins (C) are highly vacuolated (arrows) compared to cells in other regions of the liver. This vacuolization is associated with increased intracytoplasmic glycogen accumulation. This histological finding was only present in mice dosed with 20 and 40 mg/kg/day. H & E staining. Bar= l O u M . Figure 17. Periportal vacuolization in the liver of an eight week old mouse treated with 20 mg/kg/day of Aroclor 1254. There is a high degree of vacuolization in the liver cells surrounding the periportal vein (P) relative to other areas of the liver (arrow). This type of vacuolization was observed in all treatment groups, and was most common in the liver of mice treated with 10 and 20 mg/kg/day. H & E . Bar= 10u.M. Figure 18. Accumulation of lymphocytes and small macrophages in periportal spaces (arrow) in the liver of an eight week old mouse treated with 20 mg/kg/day. The lesions that were observed were mild, although they were more common and more severe in the mice treated with 20 and 40 mg/kg/day. Ff & E . Bar= l O u M . 67 and high dose mice had no cases in which there was diffuse vacuole distribution. There was also vacuolization around the periportal areas (Figure 17). This was seen in 11% (1 of 9) of the control mice livers, 60% (3 of 5) in the low dose, 82% (9 of 11) in the medium dose, and 20% (1 of 5) in the high dose. There was also an increase of infiltration of lymphocytes and small macrophages in the periportal spaces of the livers of treated mice. This was observed in 33% (3 of 9) of control mice, 20% (1 of 5) of low dose mice, 72% (8 o f 11) of medium dose mice, and 100% (5of 5) of high dose mice (Figure 18). Although this type of change is often considered incidental in mice, the change was more common and more severe (based on pathologist's grading) in the medium and high dose groups which suggests an association with Aroclor 1254 treatment. This finding is consistent with that in the literature (44, 73, 74, 76-78). The cellular changes were mi ld and were not considered clinically significant (causing disease) by the pathologist. 3.5.2 Prostate Histologically, there was a change in the predominant structure present in the prostates of treated mice in the dose response study. There was an increase in the percentage of cases in which dilated acini was the predominant structure, as dose increased. Dilated acini which were considered acini that were filled with proteinaceous material, were lined with a single row of cuboidal epithelial cells with some intraluminal infolding. A single layer of myoepithelial cells surrounded these acini. Dilated acini (Figure 19) was the predominant structure in 44% (4 of 9) of control prostates, 60% (3 of 5)of low dose, 73% (8 of 11) of medium dose, and 80% (4 of 5) of high dose. In addition, there was a dose-related decrease in the percentage of cases in which non-68 69 Figure 19. Dilated acini of the prostate of an eight week old mouse control group mouse. This was found to be the predominant structure in the mouse prostates in an Aroclor 1254 dose related manner. H & E . Bar=^5uM. Figure 20. Non-dilated acini of the prostate of an eight week old control group mouse. This was found to be the predominant structure in the mouse prostates in an inverse Aroclor 1254 dose related manner. H & E . Bar= 5 u M . Figure 21. Infiltration of discrete aggregates of small lymphocytes (arrow) in the interstitial tissue of the prostate of an eight week old mouse treated with 40 mg/kg/day of Aroclor 1254. These infiltrates were present in the prostates of treatment mice but were not detected in the prostates of control mice. H & E . Bar= u M . 70 dilated acini was the predominant structure. Non-dilated acini (Figure 20) were smaller and contained less proteinaceous material than the dilated acini. They were lined by one to several rows of densely packed, low columnar to columnar, frequently disorganized, epithelial cells. There was a greater degree of infolding than in the dilated acini. One to three layers of myoepithelial cells surrounded the non-dilated acini. This structure was predominant in 56% (5 of 9) of controls, 40% (2 of 5) of low dose, 27% (3 of 11) of medium dose, and 0% (0 of 5) of high dose prostatic tissue examined. Histological analysis of the prostates also showed intravascular multifocal lymphocytic infiltration in the prostatic stroma of mice treated with Aroclor 1254 (Figure 21), while no such infiltration was seen in the prostates of control mice. This infiltration consists of random and multifocal infiltration of the interstitial tissue by discrete aggregates of small lymphocytes with undistinguishable cytoplasm. These aggregates were often in the lumen of small vessels. There was no obvious change in periductular focal lymphocytic infiltration in the prostatic stroma of the different groups. This was characterized by the presence of well-differentiated lymphocytic aggregates with scarce but distinguishable cytoplasm in the myoepithelial layers of the prostatic ductules. None of the lymphocytic infiltrates were associated with any detectable histological alterations of the glandular component of the prostate. Despite the lack of structural changes, the cause of the presence of the infiltrates has yet to be demonstrated. 3.5.2.1 Comparison of Stromal Areas in Prostate Tissue The dose related increase in dilated acini in combination with the dose related decrease in prostate weight suggested that this may indicate a loss of a specific structure type in the prostate due to Aroclor 1254 treatment. Therefore, the stromal and ductal 71 areas in the prostate tissue slides were measured using Adobe Photoshop 5.5. The number of pixels for each structure was counted in 3 random images o f the stained prostate slide for each mouse. There was no correlation between the treatment, and the percent stromal area (Figure 22). 3.5.3 Testes Testes of mice from the different dose groups were also examined and found to be histologically normal is all cases. The tissue consisted of numerous well-differentiated seminiferous tubules separated by the septula testis. The stroma contained a normal number of Leydig cells. Active spermatogenesis was found in all tubules, as characterized by the presence of numerous spermatogenic cells. 72 45 -i 40 -Control Low Medium High Figure 22. Percent ratio of stromal area to total area in prostate tissue from mice in the dose response study. C H A P T E R 4. D I S C U S S I O N The presence of xenobiotic chemicals in the environment has raised concerns in the scientific community regarding the endocrine modulating effects of these compounds in wildlife and humans and the detrimental effects that arise in fertility, sexual development, and reproductive tract cancers (9). The ability of environmental contaminants to alter the development and function of hormone target organs such as the prostate has shown the importance of testing chemicals in the environment for their endocrine disrupting potential. The role of P C B s and P C B metabolites as estrogens has been previously documented (86). However, the effects of P C B s (and other environmental chemicals) on androgen mechanisms has not been thoroughly investigated and has not been pursued by many investigators in this field of study. Bioassays using reporter genes, ligand displacement, and an androgen regulated prostate specific transgene were used in this study to gain new information pertaining to the capacity of P C B s to influence steroid hormone action through the androgen axis in vitro and in vivo. 4.1 Luciferase Reporter Gene Assay The ability of P C B s to alter D H T induced luciferase activity in the reporter gene system through the A R was tested at half maximal and saturation levels of D H T , as well as in the absence of D H T . This study showed that Aroclor 1254 can induce an activity that is similar to androgenic activity in prostate cells in culture in the absence of a natural A R ligand. In addition, in the presence of half-maximal D H T , Aroclor 1254, Aroclor 74 1242, Aroclor 1248, Aroclor 1260, P C B 31, and P C B 42 were antagonistic and reduced the activity of the androgen driven luciferase reporter gene. When saturating levels of D H T were used in combination with the P C B s tested, similar trends were seen. A l l the Aroclor mixtures tested and P C B 42 showed a reduction in luciferase activity at high concentrations when 1 n M D H T was present. The trend of these curves tended to be less steep compared to experiments in which the transfected cells were exposed to 0.05 n M D H T . This may be because when cells are exposed to saturating levels o f D H T there is more D H T for the P C B molecules to compete with. However, P C B 31 did not show any change in luciferase activity when saturating levels o f D H T were added to the cell culture medium. Similar experiments testing the effects of these P C B mixtures and congeners through the G R showed that one P C B , the individual congener 42 was able to induce transcription in the presence of half maximal D E X . P C B 42 did not interact with the G R to drive the reporter construct in the absence of D E X . This result may be due to altered cellular distribution of the steroid hormones. Changes in the import and export of steroid hormones such as glucocorticoids due to the presence of P C B 42 could cause the increase in reporter activity that is observed, in that an inhibition of steroid hormone export from the cell would increase cellular levels of these hormones. It is also possible that this congener is binding to serum binding proteins (SBPs) and causing an increase in the amount of D E X that is in the cell, as binding sites on the SBPs are occupied by P C B 42. This mechanism has been suggested for other endocrine disruptors influencing steroid hormone activity in the cell (15). Here the investigators found that 5-hexachlorocylohexane, methoxychlor, p ,p ' -DDT, p ,p ' -DDE, and atrazine all reduced the 75 binding of 3H-5fx-DHT to the A B P of rat epididymal cytosol by a statistically significant level. In addition, 5 and y hexachlorocyclohexane, o ' ,p ' -DDT, pentachlorophenol, and nonylphenol all significantly reduced the binding of 3 H -5a -DHT to sex hormone binding globulin in human postpartum serum. B y reducing the amount of steroid hormone bound to the binding proteins, changes in signal transduction can occur due to the increase in hormone that is free to enter the cell. Finally, P C B 42 may inhibit the cytochrome P450 enzymes needed to metabolize D E X , thereby resulting in an increase in D E X in the cell and higher levels of reporter protein activity, as has been seen with some P450 enzymes responsible for steroid hormone metabolism (155, 156). P C B 42 was the only compound in this study that induced transcription through the GR. This suggests that the other P C B compounds have activity in vitro that is specific to the A R in comparison to G R . 4.2 L i g a n d Displacement Assay A l l the Aroclor mixtures tested, as well as P C B 42, demonstrated an ability to reduce 3 H - D H T bound to the A R when the amount of labeled ligand bound to the F L A G -tagged A R from a stably transfected FteLa cell line was measured. However, P C B 31 did not alter H - D H T - A R binding. This finding, in combination with the lack of luciferase reporter gene activity in the presence of saturating levels of D H T suggests that P C B 31 does not have the same abilities as the other P C B s tested to act via the A R . It is also possible that this compound does have weak antiandrogenic properties and that the binding capacity was not detected in the test system used in this study. The ligand displacement assay results indicate that these P C B s can reduce natural ligand binding. The mechanisms that produce such changes are not clear. It is possible that P C B s bind directly to the A R and prevent the natural ligand from binding. Some hydroxylated P C B metabolites have been shown to bind to A R s in the kelp bass and 76 atlantic croaker causing 50% displacement of 3H-testosterone (157). Alternatively, P C B s may block D H T import into the cell or cause an increase of the rate that D H T is exported from the cell. There are certain limitations to working with complex industrial mixtures such as Aroclors, in that it is not clear which components are responsible for the results seen. The induction of transcription as well as the prevention of A R ligand binding may be due to action of certain congeners on their own, or a combined effect of many components of the Aroclor mixtures. It is also possible that the P C B s tested in both the in vitro experiments were metabolized by the cells in culture. P C B metabolites, particularly hydroxylated P C B s may be responsible for some of the changes in luciferase reporter induction, as well as 3 H - D H T association with A R . 4 . 3 C A T Assay The single dose level study of C A T enzyme activity in the prostate showed a significant decrease in activity at 8 weeks of age. This is the first time that P C B s have been shown to be antiandrogenic in a sensitive animal model system using an androgen regulated reporter gene. While there was a decrease in mean C A T activity in the prostates of the 4 week old mice, this was not significantly different than the control group. The dose response study showed a non significant dose related decrease in C A T activity from 10 to 40 mg/kg/day. In this study, the control group C A T activity was significantly lower than the treatment groups. The source of this reduction may be in tissue preparation. Alternatively, the vehicle used in this study, canola oi l , may contribute to the decrease in C A T activity. The canola oi l for the dose response group was obtained from a different source than that used for the single dose level group. 77 Although no effects were seen in the other parameters measured, it is possible that contaminating components of the oi l caused a decrease in C A T activity in the dose response group. Canola oi l is similar to vehicles used in studies of organochlorine toxicity and reproductive effects (44, 101, 114, 133) and is a normal dietary component. The decrease in C A T activity is not observed in the groups treated with canola o i l containing Aroclor 1254 and significant effects on C A T activity and reproductive function were not observed in other studies using canola o i l in our laboratory. Vehicle free controls may have added additional data to this study, however, this control was not considered as part of the initial experimental design. In addition, the levels of C A T activity seen in the control group of the dose response study is similar to the values reported by other investigators performing.a similar study of mice treated with an identical protocol using a different source of canola o i l (Dr. C. Butler, unpublished results). The C A T activity levels from the 8 week mice in the single dose level study were approximately 2 fold higher than in the dose response study. This is consistent with variability reported using this model in our laboratory, and in the literature (143). This suggests that extreme precautions are required in tissue preparation and analysis of C A T enzyme activity. 4.4 Organ Weights and Histopathology 4.4.1 Liver The dose related increase in liver weight in mice exposed until 8 weeks of age to 20 and 40 mg/kg/day Aroclor 1254 is similar to liver effects reported in other studies (44, 73-78). Histological changes were also observed in the livers of mice treated with higher, doses. These dose related alterations, such as increased vacuolization due to glycogen 78 accumulation (158) and lymphocytic infiltrates in perportal spaces, are considered pathologically mi ld and are not associated with organ dysfunction. The serum testosterone levels are within a normal range for all dose levels. Therefore, although mi ld changes were observed in the liver, the results do not suggest that liver function, such as the ability to metabolize testosterone, is significantly compromised by Aroclor 1254 treatment in this study. Studies of changes in the activity of C Y P enzymes involved in steroid hormone metabolism due to P C B exposure has been shown by other investigators (159-162). One of these studies found a significant decrease in C Y P 2c and in serum testosterone levels with exposure to 3,4,5,3',4',5'-hexachlorobiphenyl (161). However there was not a causal relationship between these two physiological changes (162), and alteration in testicular androgenesis has been suggested (160, 162). 4.4.2 Prostate A dose related decrease in prostate weight was observed in the dose response group. These findings are consistent with other studies of rodents treated with P C B s (114). This trend may be due to apoptosis, as is seen after castration, in which the lack of androgens produces rapid involution and apoptosis in the prostate. Another possibility is an alteration in prostate growth throughout development due to the influence of P C B s . The histological findings in the prostate showed that dilated acini were the predominant structure in the ventral prostates in a dose related manner. Similar results were found in a previous study, in which male rats treated with 8, 32, and 64 mg/kg of . Aroclor 1254 during lactation developed fewer acini, less folding of the mucosa, and flattened epithelial cells in the acini of the ventral prostate (114). These observations 79 were more pronounced in the rats exposed to the highest dose of Aroclor 1254. The observation of fewer acini may be due to the dilation of the acini and the presence of prostatic fluid within. A lesser degree of infolding of the mucosa would also be consistent with the dilation of the acini, which was observed by the pathologist in this study. It is not clear i f the presence of flattened rather than cuboidal epithelial cells is an indication of a group of cells that are less active or simply flattened due to the pressure of secretions that are fill ing the acini. Due to the function of this structure, the latter is a valid possibility. Lymphocytic infiltrates in the prostatic stroma of treated mice in a dose related manner was considered an abnormal histological finding. The presence o f these cells correlates with a decrease in prostate weight and have been found in rapidly involuting prostate tumour transplants of castrated rats (163) which are undergoing appptosis and supports apoptosis as a mechanism of prostate weight reduction. Other scenarios in which such cells are found include inflammation, infection and wound healing. These events are highly unlikely to occur in the prostates of the treated mice in this study which show no other evidence of such occurrences. However, the rate of apoptosis in the prostates of the mice in this study is below a level detectable by staining in situ, and thus the possibility of an alteration in some aspects of prostate development throughout mouse development remains a strong possibility. The reduction in prostate weight with higher doses of Aroclor 1254 does not imply a clear impact on prostate function in this study. A n analysis of prostatic fluid composition and ejaculated sperm morphology and motility may indicate changes in prostate function, although these tests were not included in this work. However, the 80 possibility still remains that Aroclor 1254 has the ability to alter the import and export of steroid hormones in prostate cells. Changes in the levels of steroid hormones such as testosterone and D H T would alter prostate cell signalling and impact organ growth. The predominance of dilated acini in the ventral prostates in a dose related manner is a relevant finding, but does not indicate functional effects. Assessment of changes in prostatic fluid proteins and proteinases were beyond the scope of this project, but could provide valuable information regarding prostate functional alterations due to P C B exposure. Although the prostates of the 20 and 40 mg/kg/day dosed mice maintain the structures necessary for prostate function, these organs are lower in weight and show the presence of abnormal cell types. These findings may hold valuable implications for wildlife and human reproductive health and species fertility. 4.5 Future Directions Further studies on P C B metabolism in the cell lines used would provide information regarding the action of the P C B s tested in this study and give more specifics about chemicals which have the capacity to interact with the A R . A greater understanding of the effects of C Y P changes on androgen levels and the impact seen in vitro and in vivo may clarify some of the mechanisms of P C B action. A congener-specific analysis of the mixtures tested would provide information regarding A R ligand binding characteristics and capabilities of individual components of these complex mixtures. Future in vitro studies examining the action of a broader spectrum of congeners in combination with ligand displacement assays would narrow the focus of the active congeners in vitro, and may help uncover the mechanisms of P C B alteration of hormone action in cellular systems. Incorporating experimental anaylses of 81 mechanisms involved in the import and export of steroid hormones and P C B congeners into the sensitive in vitro and in vivo assays that have already been established may help explain the changes in cellular steroid hormone levels in culture. These studies may also offer further information regarding pathways that lead to changes in hormone regulated organs such as the prostate. The evidence presented here demonstrates the ability of P C B s to influence androgen regulated processes in vitro at the molecular level in cell culture systems as well as in vivo in a mammalian androgen regulated transgene in the reproductive organ system. These results warrant further investigation of how these ubiquitous organochlorines impact on other androgen influenced systems in the body. For example, in addition to the prostate, the seminal vesicles are androgen regulated and therefore likely to be sensitive to compounds that alter androgen action. During development, androgens affect Wolffian duct development during embryogenesis and stimulate and maintain spermatogenesis. The influences of androgens are also required for the functional differentiation of the hypothalamus and erythropoeisis stimulation. Finally, androgens exert anabolic growth effects on skeletal muscle development and on bones by causing the epiphyses to ossify in long bones, which inhibits linear growth. 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