1 IN F LUENCE OF ORGANOCHLORINE PESTICIDES ON STEROID HOMEOSTASIS AND REPRODUCTIVE INDICES IN LARGEMOUTH BASS ( Micropterus salmoides ) By NICHOLAS JOHN DOPERALSKI A THESIS PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE UNIVERSITY OF FLORIDA 2009
2 2009 Nicholas John Doperalski
3 To my fiance Adele, whose support keeps me going through the most challenging of times
4 ACKNOWLEDGMENTS A number of individuals have been crucial to my completion of this work. First and foremost is my major advisor, David Barber, who always made himself available for discussion and advice regarding my projec t, my career and life in general. O f particular importance are Dr. Barbers vast knowledge and ability to solve even the most daunting of experimental design or technical problems. The experiences I enjoyed and knowledge that I gained during my time in his laboratory will benefit me for the rest of my days. I also thank my committee members, Dr. Nancy Denslow and Dr. Margaret James, for their support and advice during my years at the University of Florida. The c ompletion of my research would not have been possible without the assistance of a number of individuals at the Center for Environmental and Human Toxicology at the U niversity of F lorida and for that I owe great thanks. First, I extend my appreciation to my lab mates Dr. R. Joseph Griffitt, April Feswick and Noel Takeuchi, who maintained a supportive and friendly at mosphere in the laboratory and lent a hand whenever it was needed. Melinda Prucha taught me the finer points of fish gonadal culturing techni ques and was always more than happy to answer my many questions and for that I am grateful Dr. Chris topher Martyniu ks inspiring attitude and assistance made possible my foray into the world of molecular cloning Without the involvement of Kevin Kroll many of the complex sampling and exposure experiments which are included in my thesi s would not have been possible. And last (but certainly not least) I thank Dan iel Spade and Roxana Weil for their advice and assistance with molecular technique s and general laboratory procedures.
5 TABLE OF CONTENTS page ACKNOWLEDGMENTS .................................................................................................................... 4 LIST OF TABLES ................................................................................................................................ 7 LIST OF FIGURES .............................................................................................................................. 8 ABSTRACT ........................................................................................................................................ 10 CHAPTER 1 INTRODUCTION ....................................................................................................................... 12 Control of Reproductive Maturation and Behavior in Teleost Fish ......................................... 12 Overview of Endocrine Disruption ............................................................................................ 17 Organochlorine Pesticides (OCPs) in the Environment ........................................................... 18 OCPs as Endocrine Disruptors ................................................................................................... 28 2 MATERIAL AND M ETHODS ................................................................................................. 35 Ex vivo Largemouth Bass (LMB) Gonadal Tissue Culture Quantification of Gonadal Steroidogenic Potential ........................................................................................................... 35 Sampling of LMB from Lake Apopka North Shore Restoration Area and De Leon Springs, Florida ....................................................................................................................... 36 Laboratory Exposure of LMB to Individual Dietary OCPs ..................................................... 37 Ex vivo Exposure of LMB Gonadal Tissue to OCPs ................................................................ 38 Analysis of Ex vivo Synthesized Steroid Hormones ................................................................. 38 Analysis of Circulating Steroid Hormones ................................................................................ 39 Analysis of Gonadal Tissue Toxicant Burden ........................................................................... 40 Quantification of Ovarian Follicle Diameter ............................................................................. 42 Cloning of the LMB 18 Kilodalton (kD) Translocator Protein (Peripheral Benzodiazepine Receptor) ...................................................................................................... 42 Statistical Analysis ...................................................................................................................... 44 3 RESULTS .................................................................................................................................... 46 Accumulation of OCPs in LMB Gonadal Tissue from Mesocosm Exposure ......................... 46 Ex vivo Exposure of Gonadal Tissue to OCPs .......................................................................... 46 In vivo Exposure of LMB to OCPs: Lake Apopka North Shore Restoration Area ................. 47 In vivo Exposure of LMB to OCPs: Laboratory Exposure ....................................................... 48 Cloning of the LMB 18kD Translocator Protein (Peripheral Benzodiazepine Receptor) ...... 50 4 DISCUSSION .............................................................................................................................. 62
6 Overview ...................................................................................................................................... 62 Effects of Ex vivo Exposure of LMB Gonadal Tissue to OCPs ............................................... 62 Effects of Individual OCP Exposure on Gonadal Steroidogenesis, Sex Steroid Homeostasis, and Morphometric Parameters ........................................................................ 67 Dietary Exposure to Ethinylestradiol ................................................................................. 67 Dietary Exposure to Dieldrin .............................................................................................. 71 Dietary Exposure to Toxaphene ......................................................................................... 73 Dietary Exposure to P, p dichlorodiphenyldichloroethylene .......................................... 75 Dietary Exposure to Methoxychlor .................................................................................... 78 Effects of OCP Exposure in a Contaminated Environment on LMB Gonadal Steroidogenesis ........................................................................................................................ 79 Overview .............................................................................................................................. 79 Effects in Male LMB ........................................................................................................... 80 Effects in Female LMB ....................................................................................................... 80 Conclusions and Future Directions ............................................................................................ 82 APPENDIX A EXTRACTION OF ORGANIC COMPOUNDS FROM FISH MUSCLE TISSUE .............. 86 B VALIDATION OF 11 -KETOTESTOSTERONE ENZYME IMMUNOASSAY ................. 88 Sample Preparation and Analysis ............................................................................................... 88 Results and Validation ................................................................................................................ 89 LIST OF REFERENCES ................................................................................................................... 90 BIOGRAPHICAL S KETCH ........................................................................................................... 103
7 LIST OF TABLES Table page 2 1 Detection parameters utilized for gas chromatography -mass spectroscopy analysis of organochlorine pesticides and internal standard compounds in extracts of tissue and feed. ......................................................................................................................................... 45 2 2 Polymerase chain reaction primers utilized to clone the largemouth bass translocator protein messenger ribonucleic acid sequence. ...................................................................... 45
8 LIST OF FIGURE S Figure page 3 1 Ovarian organochlorine pesticide (OCP) concentrations from largemouth bass (LMB) collected at the Lake Apopka mesocosm in January 2008. Colored bars represent levels, in parts per million (PPM), of each of four OCPs from three individual female LMB *, not quantifiable. ...................................................................... 52 3 2 17 beta -estradiol (E2) produced in 20 hours by ovarian explants of LMB exposed to one of four OCPs. Each exposure group consisted of vehicle, vehicle + human chorionic gonadotropin (hCG), 100 micromolar (M) toxicant, and 100 M toxicant + hCG. Each bar represents E2 production as picograms (pg) E2 per milligram (mg) explant of 2 individual explants from 2 individual LMB. ................................................... 53 3 3 E2 produced in 20 hour s by ovarian explants of LMB exposed to increasing concentrations of methoxychlor (MXC) in culture media. Each bar represents E2 production as pg E2 per mg explant of 2 individual explants from 3 individual LMB. Significant differences between treatment groups are indicated by different letters (p < 0.05). Underlined letters indicate significant differences between treatment groups, while non underlined letters represent differences between treatment groups by condition (i.e. basal or hCG -stimulated). ........................................................................ 54 3 4 E2 produced in 20 hours by ovarian explants of LMB exposed to increasing concentrations of toxaphene (TOX) in culture media. Each bar represents E2 production as pg E2 per mg explant of 2 individual explants from 3 individual LMB. .... 55 3 5 E2 produced in 20 hours by ovarian explants of LMB collected from the Lake Apopka mesocosm or control LMB collected from De Leon Springs, FL and laboratory controls. Each bar represents E2 production as pg E2 per mg explant of 3 individual explants from 5 LMB from the first mesocosm collection, 5 from the second mesocosm collection, and 10 controls. ..................................................................... 56 3 6 E2 produced in 20 hours by testis explants of LMB collected from the Lake Apopka mesocosm or control LMB collected from De Leon Springs, FL and laboratory controls. Each bar represent s E2 production as pg E2 per mg explant of 3 individual explants from 6 LMB from the first mesocosm collection, 8 from the second mesocosm collection, and 5 controls. ................................................................................... 57 3 7 Mean ovarian follicle diameter in microns (m) versus percent increase in LMB ovarian E2 production in 20 hours over basal upon stimulation with 1 unit per milliliter hCG. Each point represents an individual anima l from one of two collections from the Lake Apopka mesocosm or control animals from De Leon Springs, FL and the laboratory. Percent increase from basal is the mean increase of E2 production of 3 independent ovarian explants upon stimulation by hCG over 3 ovarian explants under basal conditions from each fish. ..................................................... 58
9 3 8 Ovarian OCP concentrations from LMB exposed to OCPs in the diet for 60 days in the laboratory. Colored bars represent levels, in PPM, of each of four OCPs from four individual female LMB in each group. ......................................................................... 59 3 9 Gonadosomatic index of LMB exposed to individual OCPs via the diet for 60 days. ..... 59 3 10 Hepatosomatic index of LMB expose d to individual O CPs via the diet for 60 days. ...... 60 3 11 Circulating and gonadal production of E2 from LMB exposed to individual OCP s via the diet for 60 days. Significant differences within panes are indicated by different letters (p < 0.05). Underlined letters represent differences between treatment groups considering both conditions (i.e. basal and hCG -stimulated). ............................................. 60 3 12 Images taken at 2 times magnification of represent ative ovarian tissue from LMB. ....... 61 3 13 Circulating and gonadal production of testosterone from LMB exposed to individual OCPs via the diet for 60 days. Significant differences within panes are indicated by different letters (p < 0.05). Underlined letters represent differences between treatment groups considering both conditions (i.e. basal and hCG -stimulated). ............... 61 4 1 E2 produced under basal conditions in 20 hours by LMB ovarian explants exposed to increasing concentrations of TOX in culture media. Each bar represents E2 production as pg/mg explant of 2 individual explants from 3 LMB. .................................. 84 4 2 E2 produced in 20 hours by LMB ovarian explants exposed to increasing concentra tions of TOX in culture media. ............................................................................ 85 4 3 Percent increase in E2 production by LMB ovarian explants upon stimulation with hC G during a 20-hour incubation. ........................................................................................ 85
10 Abstract of Thesis Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Master of Science IN F LUENCE OF ORGANOCHLORINE PESTICIDES ON STEROID HOMEOSTASIS AND REPRODUCTIVE INDICES IN LARGEMOUTH BASS ( Micropterus salmoides ) By Nicholas John Doperalski August 2009 Chair: David S. Barber Major: Veterinary Medical Sciences Organochlorine pesticide (OCP) exposure has been linked to altered plasma sex hormone levels and poor reproductive performance in largemouth bass (LMB). Possible mechanisms for disruption of sex hormone homeostasis include direct effect s on gonadal steroi d synthesis. In th ese studies effects of in vivo and ex vivo OCP exposure on LMB steroidogenesis were investigated. Incubation of ovarian tissue with methoxychlor (MXC) decreased both basal and human chorionic gonadotropin (hCG) -stimulated 17 beta -estra diol (E2) synthesis by up to 46%. Incubation with toxaphene (TOX) had no effect on basal ovarian E2 production but reduced synthesis stimulated by hCG by up to 65 % To investigate the influence of in vivo exposure to OCPs on gonadal steroidogenesis, E2 production was measured in ovarian and testis explants from LMB which spent 2 months in a mesocosm contaminated with high OCP levels. Basal and hCG -stimulated E2 production by control ovarian explants was 3 .1 0.4 and 4.3 0.6 p icograms per m illigram (pg/mg) tissue, respectively. Similar production was observed by tissue from exposed LMB (3.3 0.3 and 4.8 0.5 pg/mg, respectively). Male gonadal production of E2 was higher following 2 -month mesocosm exp osure in winter compared to control animals or animals collected from the mesocosm in late spring (maximum 834% above controls). In further
11 studies, LMB were administered TOX, MXC, p,p dichlorodiphenyldichloroethylene (DDE) or dieldrin in the diet for 2 months. A significant reduction in ovarian E2 synthesis (p < 0.05) was observed only in female LMB exposed to ethinylestradiol (EE2, positive estrogenic control) and was associated with a non -significant reduction in plasma levels. T estosterone (T) produ ction was unaffected. C urtailed gonadosomatic indices (GSIs ) were induced in females exposed to EE2 and TOX (25 and 53% of control) and DDE exposure increased hepatosomatic index (HSI) v ersu s all other groups In male LMB, dietary exposure to MXC and DDE tended to reduce gonadal production and circulating levels of E2, while DDE, MXC, and EE2 significantly reduced testis production of T (p < 0.05) while only MXC tended to lower circulating levels. DDE induced a significant increase in male HSI while GSI was unaffected. These findings suggest that OCPs are capable of perturbing g onadal steroidogenesis directly and uniquely, however compensatory mechanisms may ameliorate this effect when exposure occurs in vivo
12 CHAPTER 1 INTRODUCTION Control of Reproductive Maturation and Behavior in Teleost Fish Reproductive success is the ultimate factor influencing species survivability. While many elements such as mate selection and proper development of juveniles through reproductive age affect population viability, the ability of sexually -mature individuals of a population to procreate is crucial (Connell et al., 1999; Guillette et al., 1994) Reproductive behavior and development is regulated by the hypothalamic pituitary gonadal axis (HPGa) in vertebrate species, which in teleosts is under the influence of endogenous circannual and environmental stimuli (Jalabert, 2005) Upon exposure to appropriate cues such as water quality, availability of food, and photoperiod, the hypothalamus is stimulated to enhance synthesis of the dec aneuropeptide gonadotropin releasing hormone (GnRH) which is secreted via the hypothalamohypophiseal portal circulation to target cells in the anterior pituitary gland (Berne et al., 2004; Weltzien et al., 2004) Upon binding of GnRH to plasma membra ne receptors on gonadotroph cells of the anterior pituitary, gonadotropin secretory granules release luteinizing hormone (LH) and follicle -stimulating hormone (FSH) (Melmed and Conn, 2005) -subunit and -subunits alone do not display significant biological activity (Bogerd et al., 2005) In teleosts, two distinct populations of gonadotropic cells exist within the pituitary, unlike mammalian pituitary cell s which often synthesize both gonadotropins. Cells which produce FSH surround the nerve ramification of the proximal pars distalis (PPD), near the somatotropic cells, while LH -producing gonadotrophs reside in the periphery of the PPD (Rosenfeld et al., 2007) It is important to note, for clarity, that the original terminology for gonadotropins in fish was gonadotropic hormone I
13 and II (GtH I and II) while the corresponding FSH and LH terms were developed in reference to mammalian species. While convention has in recent years turned to the exclusive use of the mammalian -based terms due to significant funct ional similarity, others in the field continue to utilize the older GtH terminology (Harris et al., 2001; Melamed and Sherwood, 2005; Miwa et al., 1994; Rosenfeld et al., 2007; Weltzien et al., 2004) LH and FSH travel in the blood to the gonad where they bind and activate their respective receptors on steroidogenic cells (FSH R and LH -R) thus stimulating the production of the sex estradiol, E2) and testosterone (T). FSH R and LH Rs are G protein -coupled receptors containing a classi c 7 -helical transmembrane domain and multiple extracellular leucine -sheets linked by a hinge region (Bogerd et al., 2005; Kobayashi and Andersen, 2008) Interacti on of ligand with the hydrophobic extracellular binding domain results in initiation of an intracellular signaling cascade involving the production of cyclic adenosine monophosphate (cAMP) by adenylyl cyclase, which participates in both the chronic and acu te regulation of steroidogenesis. Chronic regulation occurs via activation of cAMP response elements (CREs) which promote transcription of the genes encoding enzymes necessary for steroidogenesis. Acute regulation of steroidogenesis, i.e. stimulation of the rapid production and release of steroid hormones, is via cAMP activation of protein kinase A (PKA) (Kimura, 1986; Miller et al., 2006) Activated PKA phosphorylates protein targets within the cell, including the steroidogenic acute regulatory protein (StAR) (Stocco and Clark, 1996) and gonadotropic receptor activation results in de novo StAR protein synthesi s likely by transcriptional activation of CREs (Papadopoulos et al., 2007) Activated StAR is responsible for shuttling cholesterol across the outer mitochondrial membrane (OMM), the initial and rate limiting step in acute steroidogenesis (Clark et al., 1994;
14 Hu et al., 2001; Stocco and Clark, 1996) Following the arrival of cholesterol at the inner mitochondrial membrane (IMM), a cascade of enzymatic reactions involving members of the cytochrome P450 enzyme superfamily (CYP) and multiple hydroxysteroid dehydrogenases (HSDs) ensues. Primary steroidogenesis occurs in theca cells in females and Leydig cells in males. The initi al reaction, catalyzed by CYP11A (also known as the cholesterol side -chain cleavage enzyme) yields pre gnenelone from cholesterol, which is exported to the cytosol where the remainder of the steroidogenic reactions occur (Akgul et al., 2008; Villeneuve et al., 2007a) In the female, T is secrete d to the granulosa cell where it is aromatized to E2, while the entirety of the steroidogenic process occurs in the Leydig cell in the male (Takei and Loretz, 2006) E2 and T provide feedback to the g onadotroph cells in both males and females. Administration of sex steroids in vertebrates can blunt the response of gonadotroph cells to stimulation by GnRH, thus providing negative feedback for the continued synthesis of E2 and T. Conversely, E2 maintai ned at proper concentration and time course in females exerts a positive feedback effect on gonadotropin release, leading to advancement of follicular development (Berne et al., 2004; Melmed and Conn, 2005) In teleosts, positive feedback induced by E2 is necessary for maintaining plasma FSH levels in females which promotes follicular development in the ovary. In addition, elevated plasma E2 induces the hepatic synthesis and secretion of vitellogenenin (VTG), a glycolipoprotein egg yolk precursor necessary for egg development, and thus intera ction of the HPGa with the liver is crucial for proper coordination of reproductive events in fish (Villeneuve et al., 2007b) The pituitary gonadotropins aid follicular uptake of VTG by enhancing micropinocytotic activity (Wallace and Selman, 1981) however VTG itself is not incorporated into the developing egg and the relative influence of FSH and LH on VTG uptake var ies among species (Yaron and Sivan, 2006) Instead, VTG is proteolytically cleaved
15 into yolk proteins, eg. lipovitellin and phosviti n, within the developing oocyte. During seasonal reproductive maturation in female teleost fish, both go nadal somatic index (GSI, a method for expressing gonad mass normalized to body mass) and hepatosomatic index (HSI, ratio of liver mass to body mass) increase simultaneously, reflecting the coordinated physiological changes occurring within these tissues w hich prepare the animal for procreation (McMillan, 2007) Seasonal plasma levels of E2 and T have been studied in hatcheryreared largemouth bass (Micropterus sal moides ; LMB ) in Florida. In females, E2 levels begin to climb in January and peak in early February, at levels near 4000 p icograms (pg) E2 per milliliter (mL) plasma followed by a gradual decline reaching a plateau in mid -May. Levels of E2 in females ar e then maintained at a level < 1000 pg/mL until early fall when E2 begins gradually rising prior to the rapid increase in January. In males, E2 is maintained at levels < 1000 pg/mL with a minor peak in early March, followed by a dip to concentrations < 500 pg/mL for the remainder of the year. Circulating T is highest in January through April in females, exceeding levels of 2000 pg/mL, and drops to levels circulating T reaches a high p oint in early to mid March, approaching levels of 1500 pg/mL in plasma, and declines to levels under 1000 pg/mL by June until late fall when T concentration again begins to rise. Similar trends for regulation of plasma E2 have been noted for female LMB in a wild non -contaminated reference site in Florida (Lake Woodruff), however overall concentrations were reduced, with peak seasonal E2 levels near 2000 pg/mL (Gross et al., 2008) LMB are a multi -spawner fish species, in which the potential for > 1 spawning event exists during a given reproductive season. This multiple spawning capability is made po ssible by asynchronous ovarian development; the maintenance of oocytes at several stages of maturation simultaneously within the ovary. This is in contrast to other species which exhibit group -
16 synchronous development of oocytes that are ovulated and subse quently spawned as a single event, such as striped bass ( Morone saxatilis ) (Rosenfeld et al., 2007) D espite the wealth of knowledge regarding the function of gonadotropin ligands and receptors in mammalian species, inter -species variation in receptor specificity, plasma ligand maintenance, and receptor expression introduce a degree of complexity with rega rd to role of gonadotropins in spawning behavior in aquatic species (Bogerd et al., 2005; Jalabert, 2005; Kobayashi et al., 2008) For example, in most mammalian systems, the interactions between the gonadotropins and their respective receptors is highly specific, while this is not the case in many fish (Bogerd et al., 2005) In homologous assays investigating receptor binding, LH activated both FSH R and LH R in zebrafish (Danio rerio ), African catfish (Clarias gariepi nus ), and C oho salmon (Oncorhynchus kisutch) (Bogerd et al., 2001; Kobayashi et al., 2008; Yan et al., 1992) In contrast, LH exc lusively activated the LH R in C oho salmon (Miwa et al., 1994) Regulation of circulating GtH levels has been well -studied in synchronously -spawning species such as the sa lmonids, channel catfish (Ictalurus punctatus ), and rainbow trout ( Oncorhynchus mykiss), and involves high plasma FSH during oocyte development which diminishes upon maturation, followed by a peak in LH concentration during ovulation (Kobayashi et al., 2008; Kumar and Trant, 2004; Prat et al. 1996) However a stark difference in GtH expression and release has been r ecorded in asynchronously spawning species in which both FSH and LH transcripts increase throughout the reproductive season in females (Jackson et al., 1999; Kajimura et al., 2001) and expression of both GtH genes is abundant in males during spermiation (Weltzien et al., 2003) It has been postulated by a number of groups that d ifferential expre ssion of GtH receptors in follicles of asynchronously-spawning species may facilitate the simultaneous maintenance and development of several stages of oocytes in a single animal (Kobayashi et al.,
17 2008; Kwok et al., 2005) In support of this concept, Kobayashi and colleagues have recently shown FSH R expression is inv ersely related to follicle diame ter in the A tlantic halibut (Hippoglossus hippoglossus ), and absent in the ovulated egg. Expression of LH R appeared only slightly during the secondary growth stage (0.2 1.5 m illim eter [mm] follicle diameter) and strongly during maturation, but was absent during primary growth and following ovulation (2008) Recent data collected at the University of Florida indica tes that LMB exhibit seasonal changes in follicular expression of FSH R and LH R similar to those detected in other asynchronously-spawning species (C.J. Martyniuk, personal communication, April 10, 2009). Overview of Endocrine Disruption The idea that ant hropogenic compounds released into the environment may potentially induce physiological changes in wildlife by modulating the endocrine system was introduced in the early 1990s, and termed the endocrine -disrupting contaminants hypothesis (Guillette, 2006) En docrine disruption is defined as the process of altering endocrine homeostasis of an organism, and may occur via three major modes of action: 1) by interaction with and alteration of the function of hormone receptors by endocrine -disrupting chemicals (EDCs ); 2) alteration of the steroidogenic process; and 3) modulation of the metabolism and subsequent elimination of hormones (Garcia Reyero et al., 2006; Rhind, 2002; Villeneuve et al., 2008) The mode of endocrine disruption which has gained the most attention in recent decades is that of estrogenicity elicited by xenoestrogen compounds, which was initially stimulated by the discovery and study of adverse effects of diethylstilbestrol (DE S) on embryonic and neonatal humans and subsequent investigations using rodent models (Guillette, 2006) Despite this early focus, EDCs comprise a rather diverse number of compounds which have the capability of influencing physiology via multiple mechanisms. For example, induction (or repression) of estrogenic effects in vertebrates need not solely occur via direct interaction with estrogen
18 receptors ( ERs ) rather modulation of steroidogenic enzyme activity or expression or metabolic processes may also perturb E2 levels (Guillette, 2006; Harvey and Joh nson, 2002; Sanderson and vandenBerg, 2003; van Duursen et al., 2003) The complex control mechanisms comprising the HPGa as described above make evident the many opportunities for perturbation of the system, such as modulation of synthesis or secretion of hypothalamic neuropeptides (Gore, 2002) agonistic and antagonistic influence of steroid hormone receptors and GtH receptors as well as influence on expression of these proteins (Blum et al., 2008; Khan and Thomas, 2001) interference with the steroidogenic pathway including possible enhancement (Akgul et al., 2008; Sanderson et al., 2000; S anderson and vandenBerg, 2003) and effects on biotransformation leading to dysregulation of plasma hormone half life (Thibaut and Porte, 2004; van Duursen et al., 2003) Organochlorine Pesticides (OCPs) in the E nvironment The development of the organochlorine pesticides (OCPs) was considered a tremendous breakthrough in the areas of human disease control and agriculture worldwide. Pesticides available prior to the adoption of OCPs were generally less effective and carried high collateral toxicities to both plants and animals, including humans. An example of a compound in common use during the late 19th and early 20th century is Paris Green, (copperII acetoarsenite, also known as Vienna Green or Schweinfurt Green), which gained popularity and name for use against rats in Parisian sewers (Stapleton, 2004) but was also used heavily in the United States (US) to control agricultural pests such as the cotton worm and Colorado potato beetle (Hall, 1909; Kedzie, 1876) Paris Green was also utilized to prevent the spread of malaria and typhus dur ing the second World War (WWII). However, in the fall of 1942, the chemical firm Geigy of Switzerland provided samples of a new insecticidal compound to the US Britain, and Germany. Testing performed in the US confirmed the efficacy of this novel substa nce against lice and
19 mosquito larvae, and large -scale production ensued in the spring of 1943. This new compound, first synthesized in 1874 but not recognized as an insecticide until decades later, was dichlorodiphenyltrichloroethane, more commonly know n as the organochlorine pesticide DDT (Stapleton, 2004) Due to the incredible usefulness of DDT against a myriad of pest species coupled with its low acute toxicity to mammals, DDT was heralded as a panacea of its time res ulting in its creator, Paul Meller, receiving the Nobel Prize in Physiology or Medicine in 1948 (Bate, 2007; Knsli, 1966) The success of DDT during WWII stimulated ubiquitous use of the compound in agricultural and general pestilence applications both commercially and privately in the US and abroad. In 1944, US production of DDT was 4,366 tons, utilized solely for military purposes, and climbed nearly four -fold the following year. In August of 1945, DDT was made available for commercial sale in the US and widespread agricultural adoption began in 1946 (World Health Organization, 1979) Although the contribution of DDT to the eradication of malaria in developed countries is considered minor, the OCP played a large role in controlling spread of the disease in areas of Africa, Southeast Asia, the Balkans, an d others and was crucial in the campaign of the World Health Organization (WHO) to completely eradicate the disease (Gladwell, 2001) However, complications arose quickly, most importantly resistance of insect vectors to DDT, and the eradication goal of the WHO was eventually abandoned and replaced by a more conservative management practice in 1969, involving greater emphasis on preventing human exposure to insect carriers rather than large -scale eradication (Chapin and Wasserstrom, 1981) Concern regarding the ecological impact of DDT usage emerged even prior to its release to the general public. The United States Department of Agriculture ( USDA ) Bureau of
20 En tomology published findings of an experiment performed on May 23, 1945 during which 5 pounds per acre DDT were applied to a 1,200 acre oak forest in Pennsylvania infested with gypsy moths (Lymantria dispar ). Re sults indicated no detection of live gypsy moth caterpillars hours following application. However, at least 4,000 birds died in the same forest within 8 days of DDT spraying. Another unforeseen negative outcome was the killing of ladybug populations in t he forest, which naturally control the outbreak of aphids. Without the ladybugs, aphids, unaffected by DDT, nearly destroyed the forests foliage, and were only stopped by a rainfall event. Other tests demonstrated the efficacy of only 1 pound DDT per ac re against gypsy moths, a dose which had far fewer effects on bird populations, but was still found harmful to aquatic organisms. The persistence of DDT and bioaccumulation in mammals was also known as early as 1944 via experiments with dogs, demonstratin g deposition in fatty tissue. In the spring of 1945, the USDA summarized two years of testing of DDT, calling the new pesticide a twoedged sword. Despite these findings, the pesticide was released in the US for nearly unrestricted use in late summer o f 1945 (Davis, 1971) As early as 1957, the US government began imposing regulatory actions as t he USDA Forest Service banned application of DDT near aquatic environments on lands owned or controlled by the entity. An important issue was the great environmental persistence of DDT residues, initially viewed as an attribute of the compound, but quickl y implicated in negative effects (Environmental Protection Agency, 1975) Due to the hydrophobicity of DDT and its primary metabolites (log octanol -water partition coefficients of 6.91, 6.51, and 6.02, respectively for the p,p isomers of DDT, dichlorodiphenyldichloroethylene [DDE], and dichlorodiphenyldichloroethane [DDD]), the majority of the chemical adsorbs to soil in the environment and is resistant to degradation by microbes, leading to a half life of up to 30 years.
21 In addition, DDT residues are not well metabolized or eliminated following entry into an organism and instead accumulate in tissues, especially hydrophobic sites such as adipose depots. This results in bioconcentration in many organisms, i.e. toxicant concentrations i n the organism being higher than those present in the surrounding environment. Furthermore, predator -prey relationships lead to biomagnification of these compounds, with the highest cumulative levels occurring in apex predators. Biomagnification of DDT a nd metabolites led to the well publicized population declines in many raptorial species, most notably the bald eagle (Haliaeetus leucocephalus ), due to weak or insufficient egg shell formation (Department of Health and Human Services, 2002b; Lincer, 1975; Stokstad, 2007) Accumulation of DDT residues in humans is significant; the average concentration in adipose tissue of sampled US populations was found to be 12 parts per million ( PPM ) during 1959, the year of peak use (Crathorne et al., 2001) when 80 million pounds were utilized, primarily for agricultural purposes. Eventually broader legislation was adopted in the US as the negative impacts o f DDT became more clear, leading to a complete ban in 1972 (Department of Health and Human Services, 2002b) The example of DDT use and residue deposition in multiple ecosystems is not unique. Many chlori nated pesticides or metabolites thereof utilized in the US from the 1940s until the 1980s and beyond, such as methoxychlor (MXC), toxaphene (TOX), dieldrin (DIEL), chlordane, heptachlor, and others, are persistent in both environment and organism. Followi ng the cancellation of Environmental Protection Agency ( EPA ) registration of DDT in the early 1970s, other chlorinated pesticides were adopted One example is MXC a compound very similar in structure to DDT, but possessing two methoxy groups covalently bound to the dual phenyl rings in place of the chlorine atoms of DDT. MXC was first synthesized in 1893 which, like DDT, was decades prior to the discovery of its insecticidal properties (Schneller and Smith, 1949)
22 MXC was registered by DuPont in 1 948 as a replacement for DDT and use of the newly licensed compound climbed quickly following the cancellation of DDT in the early 1970s. 300,000 to 500,000 pounds of MXC were applied per year during the 1990s in the US, largely for agricultural activitie s although the OCP was also utilized on fruit and shade trees, directly on cattle, for crop seed pre treatment, in greenhouses, in home gardens, and for landscape management. MXC is considered less persistent in the environment than DDT, as the compound i s dechlorinated and demethylated by soil microbes under both aerobic and anaerobic conditions, and breaks down in the presence of sunlight (California Environmental Protection Agency, 1999) resulting in a maximal half -life of < 1 year in most media H owever, MXC binds tightly to sediment due to its hydrophobicity and has been detected in topsoil at least 18 months following experimental application (Golovleva et al., 1984) In the Lower Fraser Valley farm areas of British Columbia, Canada, MXC was detected at < 0.02 to 0.40 PPM dry weight in crop soil between 2002 and 2003, even though MXC use had been banned for greater than 30 years in the area (Wan et al., 2005) In addition to the perceived biodegradation advantage of MXC over DDT, MXC was also seen as a less toxic alternative since it is hydroxylated very rapidly in most mammal s includi ng humans by the CYP class of enzymes, leading to elimination from the body within 24 hours of exposure, thus MXC does not bioaccumulate in tissues to the same degree as DDT. Most fish species, while possessing very low median lethal dose values compared to mammalian and bird species, also metabolize MXC rapidly and thus do not appreciably accumulate the compound. An example is the channel catfish, found to metabolize MXC to both mono and bis hydroxy metabolites by CYP hydroxylation via study of intestinal and hepatic microsomes. The velocity of the enzyme -catalyzed reaction ( Vmax) for mono -hydroxy metabolite formation by liver
23 microsomal prep aration s was found to 131 53 p ico mol es (pm) per minute per milligram ( mg ) of protein, which was increased nearly 2 -fold upon pre -treatment of fish with the classic inducer of CYP1A, 3 -methylcholanthrene, 5 days prior to microsomal preparation (Stuchal et al., 2006) In addition, more recent work has demonstrated that glucuronidation of MXC metabolites in channel catfish is rapid, typically limited by the concentration of uridine diphosphate glucuronic acid available in intestine and liver (James et al., 2008) Other organisms, such as algae, bacteria, snails, clams and mussels do bioaccumulate MXC (Department of Health and Human Services, 2002c) with mussels and snails exhibiting the highest bioaccumulation factors (BCFs) of 12,000 and 8,570, respectively. In addition, accumulatio n occurs in some fish species and is highly variable, as the sheepshead minnow (Cyprinodon variegatus variegates ) exhibits a BCF of 138 while that for fathead minnow (Pimphales promelas ) is 8,300 (Howard, 1991) Two additional OCPs, TOX and DIEL, are cyclodiene compounds which antagonize the gamma aminobutyric acid receptor (GABA R) (Bradbury et al., 2008; Zhao et al., 2003) TOX is a mixture composed of at least 670 congeners of chlorinated terpenes, produced by bubbling chlorine gas through technical camphene or pinene. TOX was manufactured and used in the US from 1947 until 1990 for agricultural and other purposes, particularly to protect cotton crops, but also for production of corn, fruit, vegetables, grains, and directly on livestock (Department of Health and Human Services, 1996) In addition, TOX was used as a piscicide to eradicate populations of fish and other aquatic organisms c onsidered undesirable for sport fishing, mostly in Canada and the northern US. In 1975 TOX was reported to be the most heavily used pesticide in the US, when 27 million kilograms were produced (Department of Health and Human Services, 1996) TOX is persistent in the environment and binds sediment, with a half -life of up
24 t o 14 years (US EPA, 1999) and has been detected in rain water in areas surrounding the Great Lakes, likely due to atmospheric transport from the southern US (Muir et al., 2006) Degradation in the environment may occur via the action of anaerobic microbes in aquatic environments which utilize the chlorine of particular TOX congeners as a final electr on acceptor in an electron transport reaction. T hus so -called aged TOX extracted from sediment and accumulated in aquatic species may exhibit altered congener ratios versus technical grade TOX (Young et al., 2009) DIEL, also a GABA R antagonist, was used directly as a pesticide but also results from the environmental or metabolic breakdown of the pesticide aldrin (Department of Health and Human Services, 2002a) Aldrin/Dieldrin were first produced in the US in 1948 by Shell Chemical in Denver, CO. The t otal mass of aldrin/dieldrin produced in the US is unknown, as the US Tariff Commission did not include these compounds in production reports until 1968. 20 million pounds per year was estimated to have been used in the US during the mid1960s, of which a pproximately 10% by mass was DIEL and 90% aldrin. DIEL was more expensive to produce versus aldrin, which explains the relative use by mass, and was recommended for application primarily to corn, hay, wheat, rye, barley, oats, orchards, vegetables, tobacc o, cotton, and citrus. DIEL was also utilized to permanently treat wood against termites and was impregnated into wool fabrics to prevent moth damage (Jorgenson, 2001) In 1970, all uses of bo th aldrin and DIEL were revoked following recognition of the persistence and associated human health risks (i.e. acute toxicity and possible carcinogenicity), although DIEL was re -registered for limited use against termites in 1972, with all uses again rev oked in 1987 (Department of Health and Human Services, 2002a)
25 DIEL is a persistent OCP which binds to soil and sediment and does not readily break down. Volatilization and soil migration are the primary modes of loss from contaminated environments. Once volatilized, DIEL may react with hydroxyl radicals in the air, causing degradation within 3 to 30 days. Degradation of DIEL in soil is dependent on the media in which it is deposited, and some microbes are capable of degrading DIEL to inactive metabolites (Department of Health and Human Services, 2002a) Waste water sludge was found to degrade 55% of DI EL within 9 days under aerobic conditions (Kirk and Lester, 1988) In soil at an application rate of 1.1 3.4 kilo g rams /hectare, the half -life of DIEL was determined to be 2.5 years with 95% loss in 8 years (Freedman, 1989) yet another investigation demonstrated a loss of biological activity of 75 100% in 3 years (Jury W. A. et al., 1987) In contrast, an area near the foundation of a home treated with DIEL for termite control was found to contain 10% of the applied mass of DIEL 21 years following the application (Department of Health and Human Services, 2002a) As well as environmental persistence, bioaccumulation and bioconcentration of DIEL occur. In aquatic species such as fish and snails, variable BCFs have been described. Rainbow trou t reportedly exhibit a BCF of 2.3 relative to wet weight, yet other studies report BCF values ranging from 2,700 to 6,145 in fish and 61,657 to 114,935 in snails. Studies of channel catfish determined that steady -state exposure to 13 part s per trillion (PPT) DIEL resulted in equilibrium between absorption and elimination after 56 days of exposure, while exposure to 49 PPT avoided equilibrium for up to 70 days of exposure (Departme nt of Health and Human Services, 2002a) Although nearly all registrations of OCPs have currently been revoked, significant concentrations of these compounds may be detected yet today in numerous areas across the world (Yedla and Dikshit, 2005) A site in the southeastern US significantly contaminated with
26 DDT metaboli tes as well as MXC, TOX, and DIEL is Lake Apopka in Florida. Lake Apopka is the third largest lake in the state, located northwest of Orlando in Orange and Lake counties near the cities of Apopka and Clermont, and is the headwaters of the Ocklawaha chain of lakes. Marshland surrounding the lake was drained in the 1940s and converted into so-called muck farms, heralded for their exceptional nutrient content. Agricultural activity continued in the area into the 1980s, a timeframe which encompassed peak usa ge of OCPs in the US. Crops grown in the Lake Apopka area included corn, carrots, radishes, spinach, and others, and were often planted more than once per year (Marburger et a l., 2002) As an example of the magnitude of OCP usage in the area, one crop of sweet corn required 13.3 pounds of TOX 68 pounds of DDT, and 0.4 pounds of chlordane per acre for production, and 13,000 acres were farmed in the North S hore Area of Lake Apopka in 1967 alone, illustrating the source of much of the residual OCP contaminatio n ( Letter addressed to C.W. Sheffield, Chairman Lake Apopka Technical Committee, Orlando, FL, from Henry F. Swanson, County Agent, Orange County A gricultural Center, Orl ando, FL July 24, 1967, p.3) Another source of OCP contamination in the Lake Apopka area originated from the Tower Chemical Company, which was situated on the southern neck of the lake near Clermont. Tower Chemical produced pesticides from 1957 through 1981 and used DDT in a synthesis reaction to produce dichlorobenzil, t he byproducts of which contain approximately 15% DDT. Liquid manufacturing waste was either sprayed onto an irrigation field or stored in an unlined retention pond at the site. A large rainfall event in 1980 caused the retention pond to spill over and contaminate the marsh area which drains into Lake Apopka with DDT. Despite cleanup effort s, soil, sediment, and wildlife at the site retain significant concentrations of OCPs versus other aquatic environments in the area (such as Lake Woodruff and Lake Orange), with the most
27 prevalent OCPs being the DDT metabolite DDE and TOX (Guillette et al., 2000; Kristensen et al., 2006) In 1992, the St. Johns River Water Management District (SJRWMD) began restoration of the Lake Apopka muck farms to wetlands. This was accomplished by flooding the previously farmed areas to an elevation of 17.81 meters above sea level, the 20-year average for the nearby Lake Griffin. Stocking of the habitat with game fish, including the apex predator LMB, began the same year under the control of the Florida Fish and Wildlife Conservation Commission (FFWCC) (Marburger et al ., 2002) Not surprisingly, bioaccumulation of OCPs by LMB stocked into the restoration areas occurred quickly. However, consumption and concentration of OCPs by LMB in the Lake Apopka area was known prior to the restoration effort of 1992. In 1985, DDT burdens in LMB fillets were determined to be in the range of 0.01 0.34 PPM and 3.3616.51 PPM in fat. More recent investigation has shown tissue concentrations of 13.287 4.393 and 0.167 0.073 PPM DDT in fat and muscle, respectively, from LMB resid ing in a reclaimed muck far m just east of Lake Apopka for approximately 3 years (Marburger et al., 2002) This same study also detected significant accumulation of chlordan e, DIEL, and TOX. Following the flooding of previously-farmed areas surrounding the lake, numerous bird species began to return to the area, so much so that Lake Apopka became a popular destination for bird -watching. The return of birds brought with it t he first evidence of significant pesticide induced ecological harm in the Lake Apopka area. In November of 1998, 5 months following the flooding of a 6,000 acre area on the northeast side of Lake Apopka (designated Unit 2), birds began dying on site. Dur ing the four months that followed, 441 American white pelicans, 58 great blue herons, 43 wood storks, 34 great egrets and smaller numbers of 20 other bird species died at Unit 2. Testing of carcasses eventually revealed toxic concentrations of OCPs, which
28 were surmised to have been ingested via the consumption of fish dwelling in the area which had accumulated the compounds from the local sediment. To prevent further ecological damage, draining of the site began in the late fall of 1998 and was complete by mid -February 1999 (Industrial Economics Incorporated, 2004) OCPs as Endocrine D isruptors Detrimental effects of endocrine disruptors are often not outwardly visible or easily detectable until significant harm to a population has occurred. Since wildlife in many cases is exposed more directly to potential EDCs (especially in aquatic environments) and in general have greater generational turn -over, animal populations may provide early indication of problems prior to effects manifesting in humans (Colborn, 1995) Strong associative evidence for the potential for OCPs to act as EDCs in aquatic species surfaced during the 1980s. During that decade, a significant decline in the American alligator (Alligator mississippiensis ) population was noted in the OCP -contaminated Lake Apopka area of Florida, while populations in many other areas of the US were climbing as a result of recently enacted species pr otections (Guillette et al., 1994) The cause seemed to be reproductive failure, as viability rates of alligator eggs collected from th e Lake Apopka area were significantly reduced compared to non-contaminated areas. Levels of OCP contaminant, specifically DDE, were shown in the mid 1980s to be elevated in Apopka alligator eggs at a level surpassing the threshold known to adversely affec t avian eggs, although the mechanism of disrupted viability was not elucidated (Heinz et al., 1991) Findings published in 1994 indicated elevated E2/T ratios in male and female juvenile alligators from Lake Apopka (Guillette et al., 1994) In the same study, males from Apopka were also found to be much more sensitive to LH administration versus animals from the control site (Lake Woodruff, FL), the response being exaggerated production of E2. In addition, several males from Apopka were lacking distinguishing external
29 sex characteristics, and every Apopka female exhibited polyovular follicles and polynuclear oocytes, a histological trait that was completely absent in animals from the control site. Furthermore, testes from Apopka males exhibited an unidentified cell type/morphology not seen in control males. Subsequent investigations of the Apopka juvenile ma le alligator population unveiled a 70% reduction in circulating testosterone levels associated with a 24% average decrease in penile size versus other populations, along with elevated concentrations of DDE in adipose tissue (Guillette et al., 1996) Exposure of alligator eggs reared in the laboratory to as little as 100 micrograms ( g ) DDE per egg at tempera tures which normally induce the male sex yielded decreased circulating levels of T and lack of external male sex characteristics in neonates, confirming the role of this OCP in endocrine disruption in the alligator (Gross et al., 1994) Evidence for endocrine disruption by OCPs also exists in reference to fish species. For example, reproductive problems have been noted in populations of white sturgeon ( Acipenser transmontanus ) from impounded se ctions of the Columbia River in the northwest US Comparison of fish from these impounded areas to those from a reference aquaculture facility yielded a negative correlation between DDE accumulation and circulating levels of T in both sexes, and 11 ketote stosteron e (11 -KT ) in males (Foster et al., 2001) Koch and colleagues (2 006) discovered that intersexual male shovelnose sturgeon ( Scaphirhynchus platorynchus ) from the Middle Mississippi River, US, exhibited higher accumulated levels of organochlorine compounds in the brain-hypothalamic -pituitary complex than mature males and GSI was negatively correlated with accumulation of organochlorines in several tissues including gonad. A decade ago, reproductive difficulties of fish stocked into pesticide -contaminated bodies of water in Florida were associated with abnormal levels of plasma sex steroid hormones (Johnson,
30 1999) In addition, reproductive problems of LMB stocked into reclaimed muck farms in the Lake Apopka area were noticed shortly after the reclamation project began (Marburger et al., 2002) suggesting that OCPs may be mediating the endocrine disrupting effects in these fish. Recent research has found blunted plasma E2 and VTG concentrations and reduced GSI in female LMB residing in an area of the St. Johns River contaminated with polychlorinated organic compounds (Palatka, FL) and GSI and c irculating 11 -KT were reduced in male fish from this location (Sepulveda et al., 2002) Clear evidence of the ability of OCPs to modulate sex hormone homeostasis in LMB was provided by Garcia -Reyero and colleagues in 2006. Typical female LMB exhibit substantially higher plasma concentrations of E2 versus 11-KT, while the inverse is true for male animals. However, upon a 120-day laboratory exposure to DDE at the highest leve l of 45.9 PPM in feed, females exhibited reduced circulating E2 and elevated 11 -KT. Dietary exposure to DIEL at a maximal dose 0.810 PPM in feed severely blunted plasma concentrations of both steroids across sexes. The possible mechanisms by which OCPs an d their metabolites elicit endocrine disrupting effects are many. To begin, p,p DDT, DDE MXC (and the mono and bis -hydroxylated metabolites of MXC ), DIEL, and TOX are known to bind ERs, although their estrogenic potency is debated but agreed to be muc h less than E2 in all cases (Blum et al., 2008; Danzo, 1997; Department of Health and Human Services, 2002b, d; Garcia Reyero et al., 2006; Kelce et al., 1995; Rama moorthy et al., 1997; Scippo et al., 2004; Soto et al., 1994) DIEL, DDE and MXC also have anti androgenic properties, and the bis -hydroxy MXC metabolite 2,2 bis (p hydroxyphenyl) 1,1,1 trichloroet hane (HPTE) has been shown to be 10-fold more potent as an anti androgen than the parent MXC (Andersen et al., 2002; Maness et al., 1998; Ramamoorthy et al., 1997; Sohoni and Sumpter, 1998; Soto et al., 1994) Using recombinant human ERs and
31 progesterone receptors (PRs) Scippo and colleagues established relative receptor binding a ffinities (RBAs) for a number of OCPs (2004) DD E, MXC, DIEL, and TOX exhibited RBAs of 0.001, 0.002, 0.002, and 0.002% respectively (with E2 binding being 100%). OCPs bound to the human recombinant PR at levels one to two orders of magnitude higher versus the ER, with the exception of DIEL, which did not bind the PR. P, p -DDT bound with the highest affinity to the progesterone recept or, with a half maximal inhibitory concentration of 2 micromolar ( M ) and 2.5% RBA. The ability of OCPs to interact with ERs has possible implications not solely for increased estrogenic potential, but also for receptor antagonism via competition. This competitive anti-estrogenic effect may interfere with proper E2 signaling, possibly resulting in reduced gonadotroph stimulation and blunted estrogenresponsive gene tr anscription. Of great importance to proper oocyte development in females is VTG gene transcription and subsequent protein synthesis in the liver, which requires E2 signaling. Thus interference with HPGa -liver coordination has the potential to greatly aff ect fecundity (Villeneuve et al., 2007b) Interestingly, the rece nt investigation performed by Garcia Reyero and colleagues demonstrated an induction of VTG transcription in female LMB in a dose -dependent manner upon dietary exposure to DDE, and a maximal 4-fold increase in VTG transcript was induced upon exposure to or al DIEL (2 006) Conversely, gene array investigation of multiple gene expression changes upon exposure of female LMB to DDE showed greater than 2 -fold downregulation of VTG 1 and 2 as well as choriogenin 2 (Larkin et al., 2002) demonstrating the need for further investigation of the E2 response element -mediated effects of OCPs. Another opportunity for disruption of the HPGa is via modulation of hypothalamic GnRH -synthesizing cells. While perturbation of circulating sex hormone levels should affect
32 GnRH production via a feedback loop, direct effects on hypothalamic cells are als o plausible. Data published by Gore (2002) indicates a repressive effect of direct MXC exposure on GnRH transcription in a hypothalamic cell line. An important finding of the study is the lack of complete ablation of t he MXC -induced effect by the ER antagonist ICI 182,780 (pharmaceutical designation) and nonparallel effects of E2 on the cell line, indicating a novel direct mechanism of OCP influence on GnRH -secreting cells not involving ERs. Another neural effect of O CPs is mediated by the cyclodiene pesticides TOX and DIEL, which have been recognized for over two decades for their ability to bind and antagonize the GABA -R, leading to increased neural excitability (Carr et al., 1999; Lawrence and Casida, 1984) The involvement of the GABA R in modulation of GtH release from the pituitary has been demonstrated in a number of species including goldfish (Carassius auratus )(Kah et al., 1992) rainbow trout (Maanosa et al., 1999) and Atlantic croaker ( Micropogonias undulatus )(Khan and Thomas, 1999) Research utilizing the goldfish model has demonstrated GABA -R mediated stimulation of GnRH release from the hypothalamus, LH release from the pituitary, an d inhibition of dopamine turnover, together resulting in a stimulatory effect on reproduction (Basu et al., 2009) In an investigation utilizing zebrafish, 2 -week dietary exposure of female animals to concentrations ranging from 0.0235 to 2.2 g TOX/g ram (g) body mass/day resulted in a significant dose d ependent inhibition of oviposition, although the mechanism of this altered reproductive behavior was not elucidated (Ree and Payne, 1997) Steroidogenesis is a logical target for perturbation of sex hormone homeostasis, and thus warrants mechanistic investigation with respect to OCP exposure considering the preponderance of data indicating perturbed cir culating sex steroid levels in exposed aquatic organisms (Garcia Reyero et al., 2006; Guillette et al., 1994; Guillette et al., 1996) The enzymatic cascade
33 involved in sex hormone synt hesis in the gonad consists of several CYP enzymes and HSDs, thus illustrating the broad opportunity for dy sregulation. The first step in steroidogenesis, the importation of cholesterol across the OMM is dependent on the activity of StAR, the expression of which has been shown to be sensitive to DDE and DIEL. Following 120 days exposure to DDE in feed, StAR message was mildly elevated in female LMB while male expression was ~4 fold higher than contro ls. Induction by dietary DIEL averaged around 2 -fold over control in females and < 1 fold in male LMB (Garcia -Reyero et al., 2006) CYP11A catalyzes the conversion of cholesterol to pregnenelone, the first reaction in steroidogenesis, and is a target of the MXC metabolite HPTE. Akgul and colleagues (2008) utilized cultured rat ovarian cells to demonstrate a dose -dependent inhibition of CYP11A enzymatic activity with concentrations of HPTE as low as 50 M by measuring progesterone (P) production. Importantly, this inhibition of P synthesis was not associated with changes in gene expression of CYP11A, StAR, a drenodoxin reductase or adrenodoxin (required for CYP11A activity) or protein levels of CYP11A, strongly suggesting effects on enzyme activity. Additional experiments also excluded effects mediated via ERs Aromatization of T to E2 is accomplished by CYP19 (aromatase) in the gonad and other tissues. Halogenated organics such as tetrachlorodibenzo -p -dioxin (TCDD) and certain polychlorinated biphenyl compounds (PCBs) inhibited CYP19 activity in a choriocarcinoma cell line, while other PCB congeners and TOX had no effect. However, the maximal concentration of TOX utilized in this study was 3.0 M, raising questions about possible effects at higher doses (Drenth et al., 1998) In vivo studies performed at the University of Florida utilizing LMB yielded reduced gonadal CYP19 expression in both males and females by DDE, while DIEL
34 induced gonadal CYP19 in females, but blunted expression in males (Garcia Reyero et al., 2006) Despite findings regarding effects on expression of steroidogenic enzymes and in vitro culture assays examining stero id synthesis, investigation of the ability of gonadal tissue from LMB exposed to OCPs in the wild or via the diet under controlled conditions has not been performed. Borgert and colleagues performed the first investigation evaluating the direct effects of OCP exposure on LMB ovarian tissue (2004) Concentrations of MXC and DDE in media to which ovarian explants were exposed were 0.01 to 10,000 parts per billion ( PPB ) which equals a maximum concentration in media of 28.931 and 31.445 M MXC and DDE. F ollowing 48 hour exposure, both compounds were found to elicit a dose -dependent reduction in T production, while E2 synthesis was unaffected. The authors also report that simultaneous exposure of ovarian explants to DDE and MXC suggested an antagonistic e ffect of both compounds upon suppression of T production, indicating different disruptive modes of action for DDE and MXC (Borgert et al., 2004) Quantification of circulating sex steroid hormones and gonadal synthetic potential from the same animal would enable concl usions regarding the contribution of direct effects of OCPs on gonadal steroid production to perturbations in circulating sex steroid homeostasis in LMB. In addition, direct effects of TOX and DIEL on LMB gonadal steroidogenesis have not been investigated The hypotheses guiding the current research are a) OCPs are capable of disrupting gonadal steroidogenesis in LMB by direct actions on gonadal tissue, and b) upon exposure to OCPs in the diet (mimicking exposure in the wild), direct disruptive effects of OCPs on gonadal steroidogenesis correlate with pertur bed plasma sex hormone levels in LMB thus establishing direct gonadal effects as a mode of endocrine disruption in LMB by OCPs. Thus, this research
35 was guided by three goals: 1) Examine the influence of ex vivo exposure of LMB gonadal tissue to OCPs on steroid production with and without stimulation by human chorionic gonadotropin (hCG ) to determine if direct effects on steroidogenesis are occurring. 2) Assess the potential for environmental exposure of LMB to OCPs to perturb basal and stimulated gonadal steroidogenesis. 3) Investigate the relationship between circulating sex steroid hormone levels (E2 and T) and gonadal steroidogenic potential from LMB exposed to environmentally relevant concentrati ons of DDE, MXC, TOX, and DIEL individually in the diet under controlled conditions in the laboratory. Exposure to single contaminants will allow for conclusions regarding unique effects of each compound, unlike the wild study (goal 2) in which animals ar e being exposed to an OCP mixture. CHAPTER 2 MATERIAL AND METHODS Ex v ivo Largemouth Bass ( LMB) Gonadal Tissue C ulture Quantification of Gonadal Steroidogenic P otential All procedures involving laboratory animals were approved by the Institutional Animal Care and Use Committee at the University of Florida. Procedures described here were utilized to assess ex vivo gonadal production of sex steroid hormones from LMB expose d to OCPs naturally in the environment and in the laboratory, as detailed below. Reproductively -mature male and female LMB were netted and eu thanized by immersion in 100 mg per L iter (L) tricaine methane sulfonate (Tricaine -S, Western Chemical, Inc., Fer ndale, WA) buffered with sodium bicarbonate. Body mass was recorded and blood was collected from the ventral tail vein, placed in heparinized glass tubes and immediately centrifuged for plasma collection. Gonad from male and female LMB was excised via ve ntral incision, weighed, minced and placed in cool L15 medium (Leibovitz; +L -glutamine, -phenol red; Sigma -Aldrich, St. Louis, MO) + 1% antibiotic antimycotic (Sigma -Aldrich) (herein referred to as media). In
36 the laboratory, gonadal tissue was further m inced to ~25 mg pieces, weighed, and placed into individual wells of a 24 -well culture plate containing fresh media. After a 2 -h ou r pre incubation period, media was aspirated to remove detritus and replaced with media alone or media plus 1 unit (U) per mL hCG (Sigma -Aldrich). Six gonadal explants from each fish were utilized, 3 exposed to hCG and 3 unexposed. Explants were flash-frozen in liquid nitrogen and media was collected for sex hormo ne analysis following 20-hour incubation at room temperature in the dark with gentle rocking. Additional gonadal tissue from each fish was stored at 20 degrees centigrade (oSampling of LMB from Lake Apopka North Shore Restoration Area and De Leon Springs, Florida C) until analyzed for OCP content. To address the fir st research goal of evaluating the potential for environmental OCPs to affect LMB gonadal steroidogenesis, 25 adult (> 1 year of age) male and female LMB and 25 juvenile LMB were captured from De Leon Springs, Florida by FFWCC personnel using electrofishing equipment. These fish were stocked into a 1/10 acre mesocosm in the North Shore Restoration Area of Lake Apopka, known to contain high sediment concentrations of OCPs, with cooperation from SJRWMD personnel. To differentiate newly stocked bass from those that may be present in the mesocosm, adult bass were implanted with radiofrequency identification chips while j uvenile fish were fin -clipped. The mesocosm also contained established populations of forage species including bluegill ( Lepomis macrochirus ), shad, and tilapia that were present in the at the time of stocking. On January 15th, 2008, following a 2 -month exposure period, the mesocosm was partially drained and LMB were collected by seine net for analysis of gonadal steroidogenic pot ential as detailed above. Additional LMB dwelling in a known uncontaminated environment, De Leon Springs, Florida, were collected by FFWCC via electro fishing on February 4th, 2008 for comparative analysis of gonadal steroidogenic potential. During the sa me
37 collection event, additional LMB from De Leon Springs were stocked into the Lake Apopka North Shore Restoration Area mesocosm and sampled approximately 2 months later on April 1st, 2008. Laboratory Exposure of LMB to Individual D ietary OCPs To address the second research goal of investigating the effects of individual OCPs on sex hormone homeostasis and gonadal steroidogenesis in LMB, male and female animals were acquired from American Sport Fish Hatchery, LLC (Montgomery, AL). Upon arrival, fish were treated with potassium permanganate (2 mg/L bath, 1/week, 3 weeks) and oxytetracycline (10 mg/L bath, 10 day course) to ablate any existing parasitic or bacterial infection. At least 30 days were allowed to elapse following treatment prior to the initiati on of experiments. Ten LMB of each sex were segregated to each of seven 147 -gallon fiberglass indoor flow through tanks supplied with dechlorinated tap water at the University of Florida Center for Environmental and Human Toxicology Aquatic Toxicology Fac ility (CEHT -ATF). Temperature was maintained at 25oC with a photoperiod of 16 h light:8 h dark. Following acclimatization for 48 hours, animals were exposed to one of DDE, MXC, DIEL, TOX, ethinyl estradiol (EE2), flutamide (FLUT), or vehicle via the diet (Chem Service, West Chester, PA). Contaminated feed was prepared by dissolving toxicants in Menhaden oil (Sigma International, St. Petersburg, FL) and mixing with Silvercup Trout Chow (Zeigler Brothers, Inc., Gardners, PA) ) at a final concentration of 1% oil. Feed was stored at 4oC until use. Nominal concentrations of toxicant in feed were confirmed via gas chromatography -mass spect romet ry (GC -MS) analysis prior to initiation of exposures. LMB were fed at approximately 1% of their body weight per day f or 60 days at which time all animals were euthanized and studied as above.
38 Ex vivo Exposure of LMB Gonadal T issue to OCPs Unexposed male and female LMB maintained at the CEHT -ATF were euthanized as above and gonadal tissue was excised, weighed, and minced into ~25 mg pieces in cool media. Minced tissue was weighed upon addition to individual wells of a 24-well culture plate containing fresh media. Following a 2 -hour pre -incubation period, media was aspirated and replaced with media only or media + 0, 2.5, 10, 25, 50, or 100 M of MXC, TOX, DDE, or DIEL (Chem Service). OCPs were dissolved in DMSO (A.C.S. Grade, Sigma -Aldrich) and added to media to yi eld a final concentration of 0.5 % DMSO in all solutions. Two explants from each of 4 fish were exposed to ea ch level of toxicant in individual wells of the culture plate, and each toxicant exposure was performed with and without stimulation with 1 U/mL hCG, yielding a total number of 4 independent exposed explants at each level of toxicant for each fish studied. Following a 20h ou r exposure, culture media was collected and stored at 80oC for hormone analysis and explants were flash -frozen in liquid nitrogen and subsequently stored at 80oAnalysis of Ex vivo Synthesized Steroid H ormones C for potential evaluation of gene expression. Ex vivo synthesized E2 and T was determined via commercially available enzyme -linked immunoassay ( ELISA ) kits (Cat. # IB79103 and IB79106, respectively, ImmunoBiological Laboratories, Inc., Minneapolis, MN). Media in which ti ssue explants were incubated was analyzed in duplicate and standard solutions used for the generation of plate standard curves were prepared in media from E2 and T stock solutions in ethanol (Sigma -Aldrich, St. Louis, MO). Plates were analyzed with a spec trophotometer (SpectraMax 250, Molecular Devices, Sunnyvale, CA) at 450 nanometers ( nm ) within 10 minutes following addition of stop solution, as per manufacturer instructions. Spectrophotometric data were analyzed using SOFTmax PRO Life Sciences Edition s oftware, v. 4.0 (Molecular Devices). Media spiked with known
39 concentrations of steroid was used to validate the ELISA kits utilized. Amount of sex steroid produced is expressed as pg steroid per mg of individual tissue explant. Analysis of Circulating Steroid H ormones Plasma samples were extracted and reconstituted in buffer prior to determination of steroid concentration by radioimmunoassay (RIA). Briefly, 20 100 microliters ( L ) of LMB plasma or steroid standards prepared in charcoal -stripped LMB pla sma were added to individual borosilicate glass tubes along with 150 L assay buffer (50 millimolar [ mM ] sodium phosphate, 0.1% gelatin, pH 7.6). The mixture of buffer and sample was extracted twice with 1.0 mL of 1 chlorobutane (99%+ pure, Acros Organics ) by vortex mixing for 30 seconds followed by centrifugation at 500 times gravity ( x g ) for 5 minutes to separate hydrophilic and hydrophobic phases. The organic phase was transferred to a new tube, and the combined extract from each sample was dried unde r a gentle stream of nitrogen at 37oC. Extracted samples were reconstituted overnight at 4oReconstituted samples were analyzed for steroid hormone levels using RIA adapted from the protocol of Jensen and colleagues (2001) with significant modification. 100 L of reconstituted plasma extract was added to borosilicate glass tubes along with 100 L of 50 nanocuries ( nCi ) per mL tritiated steroid solution ( E2 or T specific activity of 44 and 73 curies [Ci ] per millimole [ mmol ], respectively; Amersham Radiochemicals, now GE Healthcare Bio Sciences Corp., Piscataway, NJ) and 100 L of antibody solution (1:20,000 dilution of anti -E2 or anti T ; catalog number 20 -ER06 and 20TR05T, respectively; Fitzgerald Industries International, Concord, MA). Control re actions were also included as follows: B C with sample buffer (50 mM sodium phosphate, 0.1% gelatin, 0.1% RIA grade Fraction V bovine serum albumin, pH 7.6). 0 (assay buffer, tritiated steroid, and antibody solution), total activity (assay buffer and tritiated steroid only), and nonspecific binding (assay buffer and tritiated steroid only). Following incubation at room
40 temperature for at least 2 hours, tubes were placed in an ice bath for 15 minutes followed by the addition of 400 L of charcoal -dextran solution (50 mM sodium phosphate, 0.1% gelatin, 0.5% charcoal, 0.05% dextran) with the exception of the total activity tubes which received 400 L of assay buffer. Tubes were mixed by vortex briefly and incubated again on ice for 15 minutes. Following centrifugation at 1500 x g for 15 minutes at 4oAnalysis of Gonadal Tissue Toxicant B urden C, 500 L of supernatant from each tube was transferred to a 5 -mL liquid s cintillation vial to which 4.5 mL of scintillation fluid was added (ScintiSafe 30%, Fisher Scientific, Waltham, MA). Samples were counted with a Beckman LS 6000IC scintillation counter in the tritium window for 3 minutes (Beckman Coulter, Inc., Fullerton, CA). Concentration of sex steroid in plasma is expressed as pg steroid per mL plasma. OCP concentrations in tissue were determined by the method of Gelsleichter and colleagues (2005) with modification. Approximately 2.5 g of LMB gonadal tissue was weighed and placed in a Teflon -capped glass vial to which 7.0 mL n-hexanes (HEX; A.C.S. Grade, Fisher Scientific Company) and 100 L of internal standard solution containing 3 PPM d10 phenanthrene (PHEN; SPEX CertiPrep, Metuchen, NJ) and 9 PPM Ring 13C124,4 dichlorodiphenyldichloroethylene (13C12DDE; Cambridge Isotope Laboratories, Inc., Andover, MA) in cyclohexane (H igh P erformance L iquid Chromatography [HPLC] grade, Fisher Scientific) were added. Tissues were homogenized with a mechanical h omogenizer (Tekmar Tissumizer, Tekmar, Cincinnati, OH) and centrifuged for 5 minutes at 300 x g. Organic phase was transferred to a borosilicate glass tube via aspiration with a Pasteur pipet. Two additional extractions were performed by vortex mixing for 30 seconds upon addition of 3.0 mL of HEX, followed by centrifugation and collection of organic phase as above. The combined organic extract (~13 mL) from each sample was dried under a gentle stream of nitrogen at 37oC and
41 reconstituted in 3.0 mL acetonitrile (ACN; American Chemical Society Grade, Fisher Scientific Company). Further sample extraction was accomplished via the use of solid -phase chromatography columns. Extract in ACN plus a 1.0 mL ACN rinse of the extract tube was eluted through a pre conditioned C18 column (500 mg, 6.0 mL; Agilent Technologies, Santa Clara, CA). The column was washed with 500 L ACN once all extract had eluted. This combination of primary eluant, tube rinse eluant, and column rinsate was eluted through a pre -conditioned NH2 column (500 mg, 3.0 mL; Varian, Inc., Palo Alto, CA) followed by a 1.0 mL ACN rinse of the tube and 500 L wash of the column. Total eluant was dried under a gentle stream of N2 at 37oC. The dried extract was transferred to an amber HPLC vial via two 700 L washes with cyclohexane. The combined volume of cyclohexane was evaporated as above and the extract was reconstituted in a final volume of 100 L cyclohexane. Vials were sealed and stored at 4oGC -MS analysis of toxicant concentration was performed by comparing sample measurements to a mixed standard curve containing DDE, MXC, DIEL, and TOX in 3 PPM PHEN and 9 PPM 13C12DDE in cyclohexane. Mass/charge (m/z) spectrographic methods for the detection of each individual compound were developed by comparison of individual standards to known spectra followed by selection of characteristic m/z peaks, the intensity of which varied predictably with concentration. Parameters utilized for individual c ompounds are detailed in Table 2 1 A Shimadzu 17A gas chromatograph (Shimadzu Scientific I nstruments, Columbia, MD; HP 5MS column of 29 m eters x 250 microns [ m ] using helium as carrier gas) coupled with a Shimadzu QP 5000 mass spectrometer was utilized for quantification of toxicant C to await analysis via GC -MS. Recovery of toxicant from 2 individual control tissue samples spiked with an OCP mixture yielded 149 and 100% normalized to the 13C12DDE internal standard.
42 concentrations. One L of sample in cyclohexane was injected manually via Hamilton syringe into the splitless inlet. Initial temperature of oven, injector, and interface was 100, 250, and 280oC, respectively, and was maintained for at least 24 hours prior to sample analysis. The analytical program utilized a sampling time of one minute and initial oven temperature of 100oC held for 2.5 min utes followed by ramping to 190oC at a rate of 15 oC/minute, then to 250oC at 5 oC /min, and 290oC at 20oQuantification of Ovarian Follicle D iameter C /minute followed by a 5 -minute hold (total program time = 27.50 minutes ). Column pressure was maintained at 100 kilopascals and column flow was 1.4 mL/min ute Selective ion monitoring (SIM) was utilized to record spectra during set retention windows and data are expressed as the ratio of m/z for each target ion to m/z for se lected ions of the internal standards PHEN or 13C12DDE. Data are expressed as PPM toxicant in wet tissue. A minimum of 3 individual portions of ovarian tissue (~100 mg) from all LMB studied were imaged at 2 ti mes magnification utilizing an Olympus EX51 compound microscope and Olympus DP70 camera (Olympus America Inc., Center Valley, PA) coupled to a PC running imaging software (Olympus). Follicle diameter quantification was accomplished using ImageJ software v ersion 1.40g (National Institutes of Health, available for download at http://rsbweb.nih.gov/ij/ ). Briefly, the software was calibrated for measurement utilizing a scale bar included on each image and the diameter of at least 30 follicles from each animal was recorded. Each measured follicle was also marked to avoid repeated sampling. Diameter for each animal is expressed as the mean follicle diameter in m. Cloning of the LMB 18 Kilodalton (kD) Translocator Protein (Peripheral Benzodiazepine Receptor) Messenger ribonucleic acid (mRNA) sequences encoding the 18 kilodalton ( kD ) translocator protein (TSPO ; formerly known as the peripheral benzodiazepine receptor) for
43 several fish species were located utilizing the National Center for Biotechnology Information (NCBI) Nucleotide database ( http://www.ncbi.nlm.nih.gov/ ) and aligned using ClustalW2 software ( http://www.ebi.ac.uk/Tools/clustalw2/index.html ). Several areas of significant sequence homology were identified between species, and were thus ut ilized for the creation of polymerase chain reaction (PCR) primer sets utilizing Primer3 software (http://frodo.wi.mit.edu/ ). Three PCR primer sets were created to attempt to amplify the TSPO gene from mixedtissue L MB c opy deoxyribonucleic acid (DNA) (Table 2 2 ) and custom synthesized by Integrated DNA Technologies (Coralville, IA). PCR was carried out utilizing the following reagents (per reaction): 18.75 L nuclease -free water, 2.50 L 10X PCR Buffer (200 mM Tris HCl [pH 8.4], 500 mM KCl), 0.75 L MgCl2 (50 mM), 0.50 L mixed deoxynucleotide triphosphates (10 mM), 0.50 L forward and reverse primers (10 M), 0.20 L Taq DNA polymerase (all reagents: Invitrogen Corporation, Carlsbad, CA), and 1.00 L mixed tissue LM B c opy DNA. Amplicons of the correct length were identified from PCR reactions via agarose gel electrophoresis. PCR product was ligated into the pGEM T sequencing vector overnight (Promega Corporation, Madison, WI) following manufacturer instruction. E. coli cells (One Shot T OP10 chemically competent cells; Invitrogen Corp.) were transformed with the pGEM T vector containing the target insert a nd grown on a Luria Bertani ( LB )agar plate (Fisher Scientific) with ampicillin and 5 -bromo 4 -chloro3 indolyl D galactopyranoside (X G al ; Promega Corporation) for at least 24 hours. Select individual colonies were grown up in ~3.0 mL LB -broth + ampicillin overnight and plasmid preparations were created by cell lysis and isopropanol DN A precipitation (Invitrogen Corp). Sequencing of plasmid inserts was performed by the Interdisciplinary Center for Biotechnology Research Sequencing Core at the University of Florida.
44 Statistical Analysis All data are expressed as mean standard error th e mean (SEM) unless otherwise indicated. Prior to analysis, all data sets were evaluated for normality using the Kolmogorov Smirnov test and equal variance by calculating the Spearman rank correlation between the absolute values of the residuals and the o bserved value of the dependent v ariable (SigmaPlot version 11.0; Systat Software, Inc., Richmond, CA). Circulating sex steroid hormone data were evaluated by Kruskal -Wallis One -Way Analysis of Variance (ANOVA) on Ranks. Ex vivo gonadal steroid production data from males and females upon individual dietary OCP exposures and from the mesocosm were analyzed by 2-way ANOVA with location or OCP and stimulation condition (basal or hCG -exposed) as covariates, coupled with a Student -Newman -Keuls post -hoc test to d etect differences between individual group means. Steroid hormone data from ex vivo MXC exposures were analyzed utilizing 2 -way ANOVA with dose and stimulation condition (basal or hCG -exposed) as covariates, coupled with a Student Newman -Keuls post -hoc te st. Data from ex vivo TOX exposures displayed equal variance but were non -normal, thus the data set was transformed by simple log method prior to analysis by 2 -way ANOVA. Morphometric data from 2 -month in vivo OCP exposures were analyzed via 1 -way ANOVA.
45 Table 2 1. Detection paramet ers utilized for gas chromatography -mass spectroscopy analysis of organochlorine pesticides and internal standard compounds in extracts of tissue and feed. Compound M/Z Retention Time (Sec) Detection Start Time (Sec) Detectio n End Time (Sec) DDE 246.05 17.27 17.00 17.40 MXC 227.15 21.58 21.10 21.80 DIEL 262.95 17.49 17.35 17.55 TOX 159.10 18.75 18.60 19.00 PHEN 188.15 11.95 11.40 12.30 13C12 DDE 258.10 17.27 17.00 17.90 Table 2 2. P olymerase chain reaction p rimers utilized to clone the largemouth bass translocator protein m essenger ribonucleic acid sequence. Primer Denotation Sequence (5 3) LMB TSPO Forward 1 CCTACCTGGTGTGGAAGGAG LMB TSPO Forward 2 GTGGCTGCCTATGATTGGA LMB TSPO Forward 3 AGGAGCTGGGAGGTTTCACT LMB TSPO Reverse 1 GGTTGTCTCTCCATATGCAGTAGTT LMB TSPO Reverse 2 ACCAGTGCATCCTCAGTGAA LMB TSPO Reverse 3 AGTTGAGAGAGGTGGCGATG
46 CHAPTER 3 RESULTS Accumulation of OCPs in LMB Gonadal Tissue from Mesocosm E xposure Previous investigation has demonstrated that LMB dwelling in a small (1/10 acre) mesocosm in the North Shore Restoration Area of Lake Apopka accumulate significant levels of OCPs in body tissues. Whole -body analysis performed by Pace Analytical Services, Inc. for the SJRWMD of animals living in the North Shore Restoration Area for 4 months in the fall of 2007 demonstrated significant OCP deposition, with average concentrations of 16.233 8.652, 0.081 0.002, 0.500 0.061, and 8.200 1.308 PPM wet weight of whol e carcass of DDE, MXC, DIEL, and TOX. To investigate the accumulation specifically in reproductive tissue, ovaries from n = 3 female LMB that were housed in the same mesocosm for 2 months were collected and analyzed for OCP content. Significant accumula tion of DDE, MXC, DIEL, and TOX was found in gonadal tissues from these animals, with maximal levels of 18.03, 0.05, 0.28, and 13.64 PPM of DDE, MXC, DIEL, and TOX, respectively ( Figure 3 1 ). Ex vivo Exposure of Gonadal T issue to OCPs The current dataset demonstrates that LMB exposed to OCPs in the wild accumulate significant concentrations of these toxicants within gonadal tissue. Thus, the opportunity for OCPs to elicit direct effects on the gonad exists. To determine the influence of direct exposure of gonadal tissue to OCPs on steroidogenic potential, an ex vivo culture method was utilized in which tissues were exposed in vitro to a range of concentrations of toxicant for 20 hours. Evaluation of the potential for these tissue explants to synthesize sex hormones was accomplished by analyzing, via ELISA, the amount of steroid present in the culture media at the termination of the incubation pe riod and expressing production as a ratio of steroid to explant mass (pg/mg). A preliminary survey of ex vivo production revealed effects of two OCPs: MXC
47 and TOX ( Figure 3 2 ). MXC elicited a dose-dependent inhibitory effect on ovarian E2 production unde r both basal and stimulated conditions (p = .002 for overall e ffect of dose; Figure 3 3 ) with maximal reduction from vehicle -control of 46% upon exposure to 100 M MXC. Ex vivo gonadal production returned to control levels at the lowest concentration test ed (1.34 0.10, 1.64 0.23, 1.69 0.23, 1.88 0.21, 2.29 0.29, and 1.93 0.17 pg/mg at 100, 50, 25, 10, 5, 2.5, and 0 M MXC). Direct exposure of ovarian tissue to TOX produced a different effect on steroid production. Basal synthesis of E2 by ov arian explants was not perturbed by the presence of TOX in culture media even at 100 M (production from vehicle control and during 100 M exposure was 2.39 0.42 and 1.79 0.19 pg/mg, respectively). However, hCG -stimulated production was blunted in a d ose dependent fashion (Figure 3 4 ). Vehicle -control synthesis was 7.06 3.67 pg/mg and was reduced to 2.59 0.81 pg/mg upon exposure to 100 M TOX. However, this affect was ameliorated at lower concentrations of TOX, with production of 5.57 2.52 pg/m g upon exposure to a 2.5 M solution. Because of a high degree of variability likely due to inconsistent mean follicle diameters in fish studied, this interesting biological effect did not attain statistical significance. In vivo E xposure of LMB to OCPs : Lake Apopka North Shore Restoration Area We have shown that OCPs accumulate in gonadal tissue in the wild and the ex vivo exposure experiments described above establish the capability of OCPs to directly perturb normal sex hormone production by LMB gonada l tissue. However, in vivo control of steroid homeostasis is a complex process mediated by the HPGa, and therefore investigation of effects on gonadal synthesis following in vivo OCP exposure is necessary. Two stocks of LMB were introduced into the Lake Apopka mesocosm for 60 days each, the first during mid-winter and the second during late -winter 2008. Statistical analysis revealed no difference in gonadal production of E2 by ovarian tissue in animals from either visit to the mesocosm compared to control animals
48 (Figure 3 5 ). However, disparate production of E2 by male gonad was detected, with control animals and those from the first and second mesocosm samplings synthesizing 88.3 46.9, 356.1 121.4, and 28.9 13.8 fem tograms (fg) E2 per mg explant in 20 hours under basal conditions, respectively, and 76.7 38.1, 639.6 232.5, and 35.5 28.7 fg/mg, respectively, upon stimulation with hCG (Figure 3 6 ). Under basal incubation conditions, E2 production from LMB gonad from the first mesocosm sampling ex hibited a strong trend for elevation versus the second mesocosm collection (p = 0.059), while upon stimulation with hCG, E2 from the first mesocosm sampling was significantly higher than the second sampling and control gon a dal production (p < 0.001). Due to reports of differential expression of GtH receptors in follicles at various stages of maturation in asynchronously-spawning species, influence of mean follicle diameter on the ability of the gonad to enhance steroid synthesis upon stimulation with hCG was investigated. While no differences were d etected between OCP exposure groups and controls, follicle diameter was directly correlated with responsiveness to stimulation by hCG (Figure 3 7 ). For example, w hen hCG response (expressed as percent chang e from basal) was normalized to mean follicle diameter, the resulting linear correlation from control animals yielded an R2In vivo E xposure of LMB to OCPs : Laboratory E xposure value of 0.5055 and p-value of 0.0212, indicating a highly significant relationship. These data may provide clues regarding receptor expression in the follicles of LMB. Exposure of LMB to OCPs in the wild represents an experimental challenge. Lack of controlled consumption, variable weather conditions, exposure to a mixture of OCP s, and difficulty identifying appropriate control populations at similar maturation stages all introduce complications. In addition, the Lake Apopka North Shore Reclamation Area is contaminated with a number of industrial and agricultural chemicals aside from the four OCPs specifically
49 addressed by this research. To alleviate these issues and to identify which specific OCPs are capable of perturbing gonadal steroid synthesis in LMB, dietary exposure to environmentallyrelevant bioavailable doses of OCPs w as performed for 60 days in the laboratory. Following exposure to one of DDE, MXC, TOX, DIEL, FLUT or EE2 in the feed, morphometric parameters, circulating steroid levels, and gonadal steroidogenic potential were evaluated as well as OCP burden in gonadal tissue. Ac cumulated levels of OCPs in LMB ovary were similar to values observed for fish harvested fro m the mesocosm (Figure 3 8 ), indicating environmentally relevant exposures. Gonad burdens of DDE ranged from 4.160 to 8.491 PPM, .043 to 0.178 PPM MXC, 0.080 to 0.284 PPM DIEL, and 1.540 t o 2.770 PPM TOX (n=4 in each group). Results indicated chemical -specific effects on morphometric parameters. In females, a reduction in GSI was induced upon exposure to T OX and EE2 (p < 0 .05; Figure 3 9 A ) while no ef fect on GSI was detected in males, likely due to the high degree of variability of these data (Figure 3 9 B). A significant effect of DDE exposure was detected upon evaluation of HSI in both sexes (Figure 3 10). DDE enhanced HSI in females versus all other groups (p < 0.05), and while DDE increased HSI in males, statisti cal significance was detected versus MXC, CON, and DIEL groups only (p < 0.05). The ability of the gonad to synthesize steroid was compared to circulating steroid levels. Circulating se x steroid concentrations in LMB exposed to dietary OCPs for 60 days often mirrored the gonads ability to synthesize sex hormones. Plasma E2 concentrations in female LMB showed a trend for elevation upon exposure to DIEL, FLUT, DDE, and MXC (52, 112, 27, and 34% increase from control, respectively), while exposure to EE2 reduced circulating levels by 78% from control (Figure 3 11). Correspondingly, gonad from female LMB exposed to EE2
50 was reproduct ively under developed (Figure 3 12) and ex vivo E2 production was significantly reduced under both basal and stimulated conditions (p < .001). Compared to control animals, basal ovarian production of E2 from animals exposed to DIEL, FLUT, DDE, and MXC was elevated 10, 19, 11, and 8%, while exposure to TOX seemed to enhance basal but blunt stimulated production. Circulating levels of T from the same animals showed similar tre nds for effect as E2 (Figure 3 13). Once again, DIEL, FLUT, DDE and MXC tended to increase plasma T concentration (42, 122, 74, and 85% from control), however TOX also slightly elevated plasma T (25% from control), and EE2 showed no trend for effect. Influence of OCP exposure on ovarian production of T was less pronounced and more poorly correlated than for E2. DDE and MXC enhanc ed basal synthesis by 7 and 6% respectively, however all other exposures elicited a slight decrease in production. In male animals, 60 -day dietary exposure to DDE and MXC elicited a trend for reduction in circulating levels of E2 (58 and 50% reduction, p = 0.198 and 0.284, respectively), which mirrored blunted stimulated production by the gonad from the same animals (Figure 3 11). EE2 and FLUT increased plasma E2 concentration over control (61 and 33%), while DIEL and TOX reduced E2 by 12 and 20% respecti vely, none of which was correlated with gonadal production. Circulating T in male animals was elevated 31, 40, 18, and 21% by EE2, DIEL, TOX, and DDE, unchanged by FLUT, and depressed 34% from control by MXC (Figure 3 13 ). EE2, DDE, and MXC significantly reduced overall testis production of T (32, 22, and 21% reduction in basal synthesis from control; p < 0.05), while TOX curtailed stimulated production (p < 0.05). Cloning of the LMB 18kD Translocator Protein (Peripheral Benzodiazepine Receptor) Recognizi ng the association between StAR and the TSPO in cholesterol transport to the IMM evaluation of expression of the TSPO gene in gonadal tissue is desirable. As a first step in evaluating possible changes in expression of the TSPO gene, an attempt was made to clone a
51 partial sequence of the LMB TSPO to facilitate the creation of real -time q uantitative PCR primers. Using the custom -synthesized LMB TSPO Forward 2 and Reverse 2 PCR primer set (Table 2 2 ), a 234bp sequence of the LMB TSPO mRNA was cloned: 5 CGATTGTGGCTGCCTATGATTGGATTCACTGCCCTGCCACACCTGGGAGGGCTC TATGGCGGTTACATCACACGCAAAGAGGTGAAGACCTGGTACCCAACCCTACA GAAACCATCTTGGCGCCCACCAAATGCAGCGTTCCCTGTGGTGTGGACCTGTC TGTACACAGGCATGGGATATGGCTCCTACCTGGTGTGGAAAGAGCTGGGAGGT TTCACTGAGGATGCACTGGTA 3 Use of the NCBI Basic Local Alignment Search Tool (BLAST) revealed the sequenced region from LMB c opy DNA is significantly homologous with TSPO mRNA sequences from several aquatic species, including Hippoglossus hippoglossus, Salmo salar, Oncorhynchus mykiss, and Esox lu cius
52 PPM Toxicant in Ovarian Tissue 0.00 0.05 0.10 0.15 0.20 0.25 0.30 2.00 4.00 6.00 8.00 10.00 12.00 14.00 16.00 18.00 Female 1 Female 2 Female 3 DDE DIEL TOX MXC Figure 3 1. Ovarian organochlorine pesticide (OCP) concentrations from largemouth bass (LMB) collected at the Lake Apopka mesocosm in January 2008. Colored bars represent levels, in parts per million ( PPM ), of each of four OCPs from three individual female LMB. *, not quantifiable.
53 Vehicle 0.0 0.5 1.0 1.5 2.0 2.5 3.0 Vehicle+hCG DDE17-Estradiol (pg/mg gonadal explant) DDE+hCG Vehicle Vehicle+hCG MXC MXC+hCG Vehicle Vehicle+hCG DIEL DIEL+hCG Vehicle Vehicle+hCG TOX TOX+hCG Figure 3 2. 17 beta -e stradiol (E2) produced in 20 hours by ovarian explants of LMB exposed to one of four OCPs. Each exposure group consisted of vehicle, vehi cle + human chorionic gonadotropin (hCG ) 100 mi cromolar ( M ) toxicant, and 100 M toxicant + hCG. Each bar represents E2 production as p icograms (pg) E2 per milligram (m g ) explant of 2 individual explants from 2 individual LMB.
54 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 Basal hCG Stimulation Vehicle 100uM 50uM 25uM 10uM 2.5uMa a e 3 d d d d b ab ab ab c c d f e f e f e f e f17-Estradiol (pg/mg gonadal explant) Figure 3 3. E2 produced in 20 hours by ovarian explants of LMB exposed to increasing concentrations of methoxychlor ( MXC ) in culture media. Each bar represents E2 production as pg E2 per mg explant of 2 individual explants from 3 individual LMB Significant differences between treatment groups are indicated by different le tters (p < 0.05). Underlined letters indicate significant differences between treatment groups while non underlined letters represent differences between treatment groups by condition (i.e. basal or hCG -stimulated).
55 0 2 4 6 8 10 12 Basal hCG Stimulation 17-Estradiol (pg/mg gonadal explant) Vehicle 100uM 50uM 25uM 10uM 2.5uM Figure 3 4. E2 produced in 20 hours by ovarian explants of LMB exposed to increasing concentrations of toxaphene ( TOX ) in culture media. Each bar represents E2 production as pg E2 per mg explant of 2 individual explants from 3 individual LMB.
56 0 1 2 3 4 5 6 7 Mesocosm Collection 1 Mesocosm Collection 2 Controls 17-Estradiol (pg/mg tissue)Basal hCG Stimulation Figure 3 5 E2 produce d in 20 hours by ovarian explants of LMB collected from the Lake Apopka mesocosm or control LMB collected from De Leon Springs, FL and laboratory controls Each bar represents E2 production as pg E2 per mg explant of 3 individual explants from 5 LMB from the first mesocosm collection 5 from the second mesocosm collection and 10 controls.
57 0.0 0.2 0.4 0.6 0.8 1.0 Mesocosm Collection 1 Mesocosm Collection 2 Controls 17-Estradiol (pg/mg tissue)Basal hCG Stimulation p=.059 p<.001 Figure 3 6 E2 produced in 20 hours by testis explants of LMB collected from the Lake Apopka mesocosm or control LMB collected from De Leon Springs, FL and laboratory controls. Each bar represents E2 production as pg E2 per mg explant of 3 individual explants from 6 LMB from the first mesocosm collection 8 from the second mesocosm collection and 5 controls.
58 400 500 600 700 800 900 1000 1100 1200 -50 0 50 100 150 200 250 Mesocosm Collection 1 Mesocosm Collection 2 Controls Percent Increase from BasalMean Follicle Diameter (m) Figure 3 7 Mean ovarian follicle diameter in microns ( m ) versus percent increase in LMB ovarian E2 production in 20 hours over basal upon stimulation with 1 unit per m illiliter hCG. Each point represents an individual animal from one of two colle ctions from the Lake Apopka mesocosm or control animals from D e Leon Springs, FL and the labora tory. Perc ent increase from basal is the mean increase of E2 production of 3 independent ovarian explants upon stimulation by hCG over 3 ovarian explants under basal conditions from each fish.
59 PPM Toxicant in Ovarian Tissue 0.0 0.1 0.2 0.3 2.0 4.0 6.0 8.0 DDE DIEL TOX MXC Figure 3 8 Ovarian OCP concentrations from LMB exposed to OCPs in the diet for 60 days in the laboratory Colored bars represent levels, in PPM, of each of four OCPs from four individual female LMB in each group CTRL EE2 FLUT DDE MXC TOX DIELGonadosomatic Index 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 CTRL EE2 FLUT DDE MXC TOX DIELGonadosomatic Index 0 1 2 3 4 b ab ab a ab ab b A B Figure 3 9 Gonadosomatic index of LMB exposed to individual OCPs via the diet for 60 days. A) Female LMB. Significant differences are indicated by different letters (p < 0.05) B) Male LMB.
60 p CTRL EE2 FLUT DDE MXC TOX DIELHepatosomatic Index 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 b b b a ab ab ab p CTRL EE2 FLUT DDE MXC TOX DIELHepatosomatic Index 0 1 2 3 4 b b b b b b a A B Figure 3 10. Hepatosomatic index of LMB exposed to individual OCPs via the diet for 60 days A) Female LMB. B) Male LMB. Significant differences are indicated by different letters (p < 0.05) 0.0 0.2 0.4 0.6 2.0 3.0 4.0 Basal hCG b a a a a a a c d f g c d c d c d c d f g f g f g f g f g e EE2 0.0 0.2 0.4 0.6 0.8 Basal hCG FLUT CTRL DDE MXC TOX DIEL 0 100 200 300 400 500 b ab a a a ab abEE2 FLUT CTRL DDE MXC TOX DIEL 0 20 40 60 80 100 120 a b ab ab ab ab bEE2 FLUT CTRL DDE MXC TOX DIEL Circulating E2 (pg/ mL ) Gonadal Production (pg E2/mg explant ) Figure 3 11. Circulating and gonadal production of E2 from LMB exposed to individual OCPs via the diet for 60 days. Significant differences within panes are indicated by different letters (p < 0.05) Underlined letters represent differences between treatment groups considering both conditions (i.e. basal and hCG -stimulated).
61 Figure 3 12. Images taken at 2 times magnification of representative ovarian tissue from LMB. A) Representative ovarian tissue from a fish exposed for 60 days in the laboratory to dietary ethinyl estradiol B) Representative image of a laboratory control fish collected on the same day Scale bar in each image is 512 m. Testosterone as pg/mg gonadal explant 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 Basal hCG 0 1 2 3 4 5 Basal hCG b a b b ab ab ab c d c d c d d d dEE2 FLUT CTRL DDE MXC TOX DIEL 0 200 400 600 800 1000 0 100 200 300 400 500 EE2 FLUT CTRL DDE MXC TOX DIEL Gonadal Production (pg T/mg explant ) Circulating T (pg/ mL ) Figure 3 13. Circulating and gonadal production of testosterone from LMB exposed to individual OCPs via the diet for 60 days. Significant differences within panes are indicated by different letters (p < 0.05) Underlined letters represent differences b etween treatment groups considering both conditions (i.e. basal and hCG stimulated).
62 CHAPTER 4 DISCUSSION Overview The goal of the work presented was to determine the impact of exposure to OCPs on gonadal steroid production and steroid homeostasis in LMB and to attempt to ascertain whether any effects observed were due to direct effects on the gonad. To this end, experiments were performed to evaluate the effects of ex vivo exposure of LMB gonadal tissue to OCPs on steroidogenesis, to determine if exposur e to individual OCPs at environmentally relevant concentrations in feed could perturb gonadal steroidogenesis and circulating sex hormone levels, and finally to assess the effects of OCP exposure at a contaminated aquatic environment on gonadal sex steroid production in LMB. Effects of Ex vivo Exposure of LMB Gonadal T issue to OCPs Data from these studies show that MXC and TOX elicit direct inhibitory effects on E2 synthesis by female LMB gonadal explants. MXC reduced gonadal production under both basal and hCG -stimulated conditions in a dose -dependent fashion, with production returning to near control levels a t the lowest concentration evaluated (2.5 M MXC in media). The inhibition of both basal and GtH -stimulated E2 production by MXC may suggest effects on the synthesis pathway itself. CYP11A, the enzyme responsible for the conversion of cholesterol to preg nenelone during steroidogenesis in the gonad, is sensitive in mammalian systems to HPTE, the bis -hydroxy metabolite of MXC. Akgul and colleagues exposed primary cultures of rat theca interstitial cells to increasing concentrations of HPTE in media (maxima l 500 nanomolar [ nM ]) for 24 hours resulting in a dose dependent reduction in CYP11A activity as measured by release of cleaved tritiated side chain from radiolabeled 25-hydroxy cholesterol. There was no apparent saturation of the inhibitory response over the HPTE concentration range evaluated. It is
63 important to note that use of 25-hydroxy cholesterol does not require the activity of StAR for transport across the OMM, thus excluding transport as a possible mechanism of inhibition (2008) In the current studies, ovarian explants were exposed directly to parent MXC in culture media, however biotransformation to HPTE cannot be rul ed out. In the mouse, surface epithelium of the ovary possesses CYP enzymatic activity, and expression of CYP2C29, which is capable of hydroxylating MXC, is sensitive to induction by 14 day exposure to as little as 3 M MXC in media, an effect blocked by the ER antagonist ICI (Symonds et al., 2006) Since concentrations of HPTE in tissue or culture media were not quantified in the current study, a distinction between the influence of MXC versus mono or bis -hydroxy metabolites cannot be made. In addition to MXC metabolite effects, potential cytotoxicity of MXC exposure was not evaluated in the current study. Cytotoxic effects following a 24 hour exposure of PLHC 1 cells (fish he patoma cell line) to MXC indicated a midpoint cytotoxicity (NR50) value of 50 M in solution (Babich et al., 1991) In a study evaluating the effects of direct MXC exposure on primary cultures of pig granulosa cells, a 48 -h ou r exposure to 10 M MXC had no effect on cell number or cAMP levels compared to unexposed culture, however inhibitory effects on progesterone synthesis were noted (Chedrese and Feyles, 2001) Therefore, while toxicity data specifically related to fish gonadal tissue is not available, the possibility of MXC induced cytotoxicity to LMB gonadal explants at the high end of the MX C concentration curve utilized in the current study must be acknowledged. However, even at a concentration unlikely to induce cytotoxicity (2.5 M), hCG -stimulated production of E2 by LMB ovarian explants was significantly reduced versus vehicle -control ( p < 0.05), suggesting a genuine effect on steroidogenesis.
64 The current data are in contrast to work performed by Borgert and colleagues which indicated no perturbation of LMB ovarian E2 production upon exposure to MXC in culture media (2004) Howe ver, the highest concentration of MXC used for those investigations was 10,000 PPB, which is equivalent to 28. 931 M, m uch lower than the maximal 100 M concentration used in the current investigation, which may have prevented recognition of a dose respons e. In addition, Borgert utilized a 48 hour incubation period with approximately 100 mg of tissue per well, while the current study included approximately 25 mg of ovarian tissue and incubated for 20 hours. Lastly, the previous work utilized the Hill mode l to represent dose response relationships for each of three animals individually, while the current investigation used the mean of E2 production from 3 individual fish, resulting in detection of a statistically significant influence of MXC concentration on ovarian E2 production. The effect of ex vivo ovarian exposure to TOX was different than that elicited by MXC. TOX, even at a concentration of 100 M in media, did not significantly inhibit basal ovarian E2 synthesis (Figure 4 1 ). However, TOX inhibited an hCG -stimulated increase in E2 production in a dose -dependent manner. Stimulation by hCG occurred at concentrations of 10 and 2.5 M TOX in culture media, while significant inhibition of hCG -stimulation occurred at concentrations of 100, 50, and 25 M. This inhibitory effect on stimulated ovarian E2 production was variable among the animals included in this investigation Explants exposed to TOX from two of the three LMB studied had very similar responses in which neither basal nor hCG -stimulated E2 production was significantly affected (Figure 4 2 A ). However, hCG -stimulated E2 production from ovarian explants of the third animal was sensitive to inhibition by TOX, and it was this animal which drove the overall observed trend for the dose -dependent inhi bitory action of TOX (Figure 4 2B ).
65 Explants from t he two a nimals which were affected similarly by TOX expo sure responded poorly to hCG ad mi ni stration, while the third animal was highly responsive to GtH stimulation (as indicated by an increase in E2 production) The nonres ponder animals were found to have mean ovarian follicle diameters of 605 and 670 m with maximum observed diameters of 1050 and 1177 m, respectively while the third animal possessed a mean follicle diameter of 917 m and maximum of 1391 m Body weights of all animals were similar, at 514, 404, and 402 g, while GSI was 2.2, 2.3, and 3.9 for the two unresponsive fish and the responder, respectively. Graphical c omparison of the potential for TOX to inhibit hCG stimulation of gonadal E2 production as a function of mean follicle diameter of each animal indicates a possible follicle diameter dependent sensitivity to TOX (Figure 4 3 A). These data suggest ovarian E2 s ynthesis may be uniquely sensitive to TOX during later stages of reproductive m aturation when ovarian follicles are more responsive to stimulation by LH receptor ligands such as hCG. A lack of strong response of ovarian tissue to stimulation by hCG was no ted during ex vivo exposure to MXC as well. Mean follicle diameters from fish utilized for MXC exposures were 573, 674, and 622 m with maximum observed follicle diameters of 976, 1014, and 942 m, respectively. Body weights of these animals were 304, 370, and 440 grams while GSI was 2.4, 2.7, and 2.3, respectively. When expressed graphically, it is obvious that stimulated ovarian E2 production is blunted upon ex vivo exposure to MXC (Figure 4 3 B) however ovarian explants from these animals appeared to respond to hCG to a slightly greater degree than animals utilized for TOX exposures with similar follicle diameters. As was demonstrated by a subset of the current data, LMB mean ovarian follicle diameter seems to be positively correlated with the ability of ovarian tissue to respond to hCG stimulation. This
66 finding is discussed in detail below (see Effects of OCP Exposure in a Contaminated Environment on LMB Gonadal Steroidogenesis ). In the work of Borgert and colle agues, exposure of LMB ovarian explants to 3.33 U/mL hCG demonstrated a trend for stimulation of sex steroid production, but this effect was not found to be significant (2004) The current studies utilized a concentration of 1 U/mL hCG, a concentration deemed by preliminary data to be most appropriate for gene expression assays (M. S. Prucha personal communication, January 15, 2008), however this concentration may be inappropriate for assays evaluating maximum steroidogenic potential of gonadal tissues. Other studies investigating gonadal steroidogenesis in fish species have used a wide range of hCG concentrations for stimulation of steroid production, from 5 U/mL hCG (Benninghoff and Thomas, 2006) 10 U/mL (Lister et al., 2008) and up to 20 U/mL (Singh and Joy, 2009) Since only stimulated ex vivo ovarian E2 production was affected by direct exposure to TOX this suggests effects on GtH signaling and/or the induction of the acute steroidogenic regulatory system. The intracellular signaling cascade f rom receptor to mitochondrion involves cAMP production, however work of Kodavanti and colleagues (1988) found no effect of TOX exposure on adenylate cyclas e or phosphodiesterase activity in rat brain synaptosomes thus suggesting effects may not be mediated via cAMP. Integral for the acute stimulation of steroidogenesis is the function of StAR, which has been demonstrated to require the function of the TSPO (formerly known as the peripheral benzodiazepine receptor). The TSPO is crucial for the recruitment of StAR to the OMM and responds to the presence of cholesterol, and reductions in expression of TSPO in Leydig cells have been shown to attenuate hormone induced transport of cholesterol across the OMM and subsequent steroid synthesis, independent of the presence of StAR protein. In addition, de novo StAR protein synthesis was not induced by hCG
67 administration to normal Leydig cells when a TSPO antagonist was co administered (Papadopoulos et al., 2007) Some l igands whi ch bind and antagonize the GABA -R also bind the TSPO, and TOX is well -recognized for the ability to bind and inhibit the GABA R in fish and other species (Carr et al., 1999; Lawrence and Casida, 1984) Thus, TOX may antagonize the TS PO resulting in inhibit ion of hCG -stimulated steroidogenesis, whereas basal steroid production is not affected. To enable investi gation of TSPO expression, a 234base pair segment of the LMB TSPO mRNA was cloned. This sequence will be utilized to create real time PCR primers for quantification of TSPO mRNA in LMB gonadal tissues upon exposure to OCPs and for seasonal expression determination. Effects of Individual OCP Expo sure on Gonadal Steroidogenesis, Sex Steroid Homeostasis, and Morphometric P arameters LMB were exposed in vivo via the diet to one of four OCPs, including DDE, MXC, DIEL, and TOX or vehicle (menhaden oil) In addition, b ecause some OCPs have been shown t o possess estrogenic and anti androgenic properties, two additional control exposures were included in this study. One group of LMB was exposed to the synthetic estrogen EE2 in feed at a concentration of 2 PPM while a second group was exposed to the anti androgen FLUT at 750 PPM in feed. Dietary Exposure to Ethinylestradiol In female LMB, dietary exposure to 2.0 PPM EE2, included as a positive estrogenic control, induced a significant reduction in GSI an d inhibited gonadal maturation (Figure 3 12), reduced circulating E2 and significantly reduced gonadal production of E2, but did not affect circulating or ex-vivo g onadal production of T. EE2, commonly used as an active ingredient in human birth control pills, acts in mammals to inhibit the release of GtHs from the pituitary (Strobl, 2003) Thus the observed reduction in GSI and both circulating and ex vivo gonadal
68 production of E2 by EE2 may be via inhibitory effects on pituitary release of FSH and LH. Hepatic VTG production is known to be elevated upon exposure to synthetic estrogens such as EE2 (Denslow et al., 1997; Peters et al., 2007) however foll icular VTG uptake is stimulated by GtHs (Wallace and Selman, 1980; Walla ce and Selman, 1981) therefore reduced plasma GtH levels coupled with unknown receptor expression in imma ture follicles may have prevented normal follicular VTG incorporation Since the majority of mass gained by developing oocytes is from VTG protein uptake (Wallace and Selman, 1981; Yaron and Sivan, 2006) follicles would be expected to remain pre -vitellogenic and GSI would remain low, as observed in the current study. A ne gative effect of EE2 on GSI similar to that indicated by the current data was shown in female zebrafish as soon as 6 days following exposure to 25 ng/L EE2 in water, and 12 days upon exposure to 10 ng/ L (Van den Belt et al., 2002) In the same investigation, a lack of mature vitellogenic oocytes was observed in ovaries from exposed fish relative to controls. The reduction in GSI persisted for 24 days during EE2 exposure, and was reversible upon cessation. Similarly, a 2 -month exposure of female Japanese medaka ( Oryzi as latipes ) to 10 or 100 ng/L EE2 in water followed by a 6 -week recovery significantly reduced GSI and the percentage of females carrying eggs in a dose -dependent fashion (Scholz and Gutzeit, 2000) The lack of effect of dietary EE2 on female circulating or gonadal production of T observed in the current study may be due to modulation of CYP 19. The proposed reduction in circulating gonadotropins as a result of EE2 exposure may have caused diminished activity of this enzyme which is responsible for aromatization of T to yield E2. Curtailed CYP19 activity may thus contribute to the E2 deficit observed while increasing levels of T, as less T would aromatized. Following this logic, it may be possible for net steroidogenesis to be reduced while normal levels of T are maintained. It has been known for decades that FSH stimulate s induction
69 and activation of CYP19 in mammalian species and, via the androgen receptor, T can itself stimulate synthesis of CYP19 in the granulosa cell (Davis and Heindel, 1998) In human granulos a cells in culture, FSH stimulates CYP19 activity, transcription, and protein synthesis. In addition, CYP19 activity diminishes in a time -dependent fashion when FSH is removed (Steinkampf et al., 1987) Further supporting evidence for the role of CYP19 in the response to EE2 exposure comes from the female Japanese medaka, as a 7 -day water expo sure to 500 ng/L EE2 induced a 5 -fold down regulation of ovarian CYP19A expression (Zhang et al., 2008) Dietary exposure of male LMB to EE2 did not change GSI, however plasma E2 and, to a greater extent, T showed trends for elevation. In addition, EE2 elicited a significant reduction in gonadal T production but did not affect production of E2. A similar effect of EE2 on male animals was observed by Tilton and colleagues; a significant increase in circulating E2 was detected 7 days following intraperito neal (IP) injection of channel catfish with 1 mg/kg EE2 (2001) Peters and coauthors observed an increase in GSI in male mummichog ( Fundulus heteroclitus ), a species whose reproductive strategy is similar to LMB, following a 21 day exposure to 100 ng/mL EE2 in water (2007) Conversely, a series of s tudies utilizing male fathead minnow (FHM) indicate d a dose -dependent reduction of GSI and circulating sex hormones upon 21-day water exposure to a maximal dose of 40 ng/L EE2, which the authors explain is likely due to a negative feedback mechanism on the pituitary (Salierno and Kane, 2009) However, the reproductive strategy of the FHM is quite different from that of LMB Female FHM are fractional spawners capable of producing large clutches (50 100 eggs) year round every 3 5 days Associated with this frequent spawning behavior are fast and large changes in circulating sex hormone levels. For example, plasma E2 in females has been found to peak at 1 -day post -spawn during a 4 -day spawning interval maintained in the la boratory,
70 reaching levels of approximately 10 ng/mL and dwindling to around 4 ng/mL by day 4 following spawning. Similar changes in magnitude of circulating T were also observed. Significant f luctuations in male plasma sex hormone levels do not occur during the short spawning interval (Jensen et al., 2001) Therefore, differences in reproductive behavior and ma intenance of circulating sex steroids between FHM and LMB may preclude direct comparison of the effects of endocrine disrupting compounds among these species A possible cause of increased plasma E2 and T levels in male LMB upon dietary exposure to EE2 w ithout an associated increase in gonadal production may involve modulation of metabolic and elimination processes. Exposure to natural and synthetic estrogens has been shown to alter the activity and expression of enzymes involved in steroid metabolism in several fish species (Pajor et al., 1990; Stegeman et al., 1982) CYP3A enzymes in particular are involved in the metabolism of steroids (Miranda et al., 1989) and thus alteration of expression or activity of these isozymes may alter proper steroid elimination and thus homeostasis For example, e xposure of male Japanese medaka to a maximal IP dose of 100 g E2 /g fish caused downregulation of CYP3A isoform protein s in the liver at 48 hours post injection, however mRNA expression corresponding with these CYPs was unaffected. In the same study, male medaka were co cultured with female animals for 7 days resulting in a water borne concentration of 43 ng/L E2 (versus 3.3 ng/L in male -only tanks) produced by steroid elimination from these untreated animals which induced similar male CYP3A protein suppression to that observed via IP E2 injection, and was associated wi th increased serum E2 levels. Finally, e xposure of male medaka to EE2 in water (highest concentration of 100 ng/L) for 24 hours significantly increased serum E2 concentration (Kashiwada et al., 2007) Furthermore, season al changes in CYP3A
71 isoforms have been detected in LMB, with females exhibiting up to 4 times the expression of CYP3A 68 versus males during t he peak of the reproductive season (Barber et al., 2007) Another possible explanation for the observed increase in E2 observed in males following dietary exposure to EE2 may involve CYP19. E xpression of CYP19 in testicular tissue of medaka has been shown to be increased by 9.5 -fold upon exposure to 50 ng/L EE2 and 21.5-fold when exposed to 100 ng/L EE2 in water (Kashiwada et al., 2007) thus increased expression of CYP19 in LMB testicular tissue may account for increased circulating E2 levels compared to T observed in the current study however the lack of incr eased gonadal E2 production makes this scenario unlikely. Dietary Exposure to Dieldrin Dietary expo sure to DIEL for 60 days (2.95 PPM in feed) induced a minor and nonsignificant increase in plasma T levels in both sexes, although gonadal production was un affected. Gonadal E2 production was increased slightly but not significantly, in both sexes, while DIEL appeared to increase circulating levels in females and possibly reduce levels in males. No significant effect on GSI or HSI was noted in either sex. Exposure of LMB to dietary DIEL (0.810 PPM in feed) for 120 days has been previously shown to induce StAR expression in the gonad of male and female LMB (Garcia Reyero et al., 2006) and thus this may account for the slightly enhanced gonadal production of E2 in both sexes observed currently. However, the same investigation indicated an obvious reduction of both circulating E2 and 11-KT in female fish and reduction in 11 -KT in males, which was not observed in the current data set. The difference in outcomes of these studies may be related to duration of exposur e, as current experiments were performed for 60 days only. Studies performed by Johnson and colleagues demonstrated no consistent effect on circulating E2 or 11 -KT in male or female LMB upon 30day exposure to a maximal concentration of 0.810 PPM DIEL in feed. Additionally 120 days of
72 exposure to dietary DIE L did not affect plasma E2 in males however 11-KT appeared to be reduced in a dose -dependent fashion. In females exposed for 120 days to the same, 11 -KT was unchanged, but DIEL reduced circulating E 2 at all doses evaluated (2007) These studies make clear the variability in response of LMB to DIEL exposure which may involve differences in dose, duration of study, or time of year during which the study was performed. Previou s work has indicated induced CYP19 expression in female LMB but repression in males upon dietary exposure to 0.810 g/g DIEL for 120 days (Garcia -Reyero et al., 2006) Correspondingly, the current dataset indicates a possible increase in female plasma E2 and a slight reduction in male levels however these findings were not significant nor did they correlate with gonadal production. Lack of correlation between circulating T levels and gonadal production of T observed in the present study may be explained by suppression of metabolic activity, leading to elevated plasma concentrations Changes in expression of metabolic CYP enzymes in LMB have been noted following 120day exposure to 0.810 PPM DIEL in feed (Barber et al., 2007) Specifically, hepatic expression of CYP3A68 was induced by 2.1 -fold whi le CYP3A69 was repressed by 2.5 -fold in female LMB after a 120 day exposure to dietary DIEL, while changes in expression of CYP3A isozy mes were not detected following 30 day DIEL exposure In male fish, DIEL was not found to modulate CYP3A expression even after 120 days In addition, evaluation of E2 sulfotransferase (SULT) activity from hepatic cytosol of male and female LMB indicated no modulation upon exposure to 0.810 PPM DIEL in the diet for 120 days compared with control animals (D.S. Barber, personal communication, October 9, 2008). Therefore, while changes in steroid metabolism may explain a slight elevation in plasma T levels of female LMB, the literature does not support this mechanism in males. Importantly, care must be taken to prevent over interpretation of the current data set, as no changes discussed
73 in the context of DIEL exposure were found to be statistically signific ant. The lack of GSI changes in DIEL -exposed animals in the current study seems to corroborate the absence of significant changes in circulating sex hormone levels. Dietary Exposure to Toxaphene Sixty -day dietary TOX exposure (25.78 PPM in feed) had a mi ld non -significant diminishing effect on circulating E2 in male and female LMB, but appeared to enhance basal gonadal production of E2 while reducing relative stimulated synthesis. M ale stimulated gonadal production of T was significantly reduced versus control. Circulating T in both sexes was similar to control values, but appeared slightly elevated. As demonstrated upon ex vivo exposure to TOX in the current investigation production of E2 by ovarian explants was inhibited only upon stimulation with hCG. Interestingly, the only statistically significant effect of TOX noted upon in vivo exposure is a reduction of hCG -stimulated E2 production. This strengthens the postulation that TOX may be acting directly on the gonad to inhibit processes involved in GtH mediated acute steroidogenesis. Even though TOX was not added to the culture media during evaluation of steroidogenic potential following dietary exposure, significant levels of TOX have been shown via the current data set to accumulate in the gonad upon dietary exposure with a mean value of 2.14 PPM or 5.16 M in ovary following 60 day dietary exposure in the laboratory, and thus the opportunity for TOX to elicit a direct effect on the gonad following excision exists. Data from ex vivo exposure of ovary to TOX indicates a possible inhibitory effect of direct exposure to as low as 2.5 uM TOX in media, although this cannot be stated with certainty due to variability in the response among fish. In other work, TOX has been shown to directly inhibit st imulated steroidogenesis in adrenal tissue in rats. Mohammed and colleagues found a significant reduction of adrenocorticotropic hormone (ACTH) induced corticosterone production by adrenal cortical cells from rats exposed to 1.2 PPM TOX in the diet for 5 weeks.
74 In the same investigation, direct exposure of adrenal cells from control animals to TOX was also found to inhibit ACTH stimulated corticosterone production, with a half -maxi mal inhibition concentration of 11.7 PPM in media (1985) TOX was the only OCP currently studied which significantly reduced GSI compared to control animals. The effect on GSI may have multiple causes, including inhibition of GtH stimulated steroid production or GtH -mediated uptake of VTG into follicles. In addition TOX has been shown via numerous investigations to act as a xenoestrogen (Lemaire et al., 2006; Ramamoorthy et al., 1997; Soto et al., 1994) The reduction in GSI in female animals observed in the current study is similar to the effect induced by EE2, which suggests TOX may be eliciting this effect via an estrogenic influence. However, significant reductions in both gonadal production of E2 and circulating levels were observed in females follow ing dietary exposure to E E2, while TOX did not induce these changes. Therefore, TOX may act via a mechanism independent of its estrogenic potential. While it is possible that TOX may also be affecting cAMP regulation within the cell leading to perturbed intracellular signaling and thus altered acute steroidogenesis, Kodavanti and colleagues (1988) found no effect of TOX exposure on adenylate cyclase or pho sphodiesterase activity in rat brain synaptosomes thus sug gesting effects of TOX are not mediated by direct modulation of cAMP levels While stimulated production of T by male gonad was significantly reduced versus control upon TOX exposure in the current study circulating T was similar to control levels. E2 production under basal conditions was increased slightly in both sexes, while basal T synthesis was mildly inhibited. The differential levels of T and E2 production may suggest the involvement of CYP19. However, the literature does not support this postulation as Drenth and colleagues found no effect of exposure of a choriocarcinoma cell line to 3.0 M TOX for 18
75 hours on aromatase activity (1998). However, evaluation of the effects of direct exposure of fish gonadal tissue to TOX on CYP19 activity or expression has not been performed. Dietary Exposure to P, p di chlorodiphenyldi chloroeth ylene Dietary exposure to 78.84 PPM DDE in feed for 60 d ays had little effect on ovarian E2 production and elevated circulating levels slightly (27% increase over control) In males, DDE reduced E2 production by the greatest extent of all compounds studied, and reduced circulating levels of E2 below the limit of detection. DDE had a minimal effect on ovarian T synthesis although circulating levels seemed elevated compared to control. Circulating T in males was very similar to controls, but overall ex vivo production and stimulated production by gonad w as sign ificantly reduced. Past studies have shown a number of effects of DDE in LMB. Upon 120day exposure in the diet (45.9 PPM in feed), female LMB exhibited reduced circulating E2 and elevated 11 -KT, while males were largely unaffected (Garcia -Reyero et al., 2006) However, an additional study which investigated the effects of 30 and 120day exposure of LMB to several concentrations of DDE in feed (maximal 136 PPM) showed no effect on circulating E2 or 11 -KT in females upon 30day exposure, and a reduction of plasma E2 at day 120 which was inconsistent among animals No effects on E2 or 11-K T were seen at the 30 day timepoint in males, however a slight (though non-significant) positive dose dependent effect of DDE concentration in feed on circulating E2 levels was detected at 120 days. No effects on GSI compared to control were detected in any group, supporting the lack of chan ges in circulating sex hormones (Johnson et al., 2007) Still other work shows decreases in circulating E2 and 11-KT in male and female LMB upon 50-day dietary exposure to 5 PPM DDE (Muller et al., 2004) Data from the current exposure of LMB to 78.8 PPM in feed for 60 days do not agree, as DDE dramatically reduced plasma E2 in males but did not affect levels in females. In support of this current finding, values
76 of GSI similar to control were recorded f or females exposed to DDE, indicating normal gonadal development which suggests physiological plasma E2 concentrations. GSI values for males were too variable to glean meaningful relationships. Thus, duration of exposure and other unknown factors may inf luence effects on sex steroid homeostasis in LMB as multiple trends for maintenance of circulating levels have been recorded upon exposure to DDE. The work of Garcia Reyero also indicated that dietary DDE exposure repressed CYP19 expression in males and females and StAR message was mildly elevated in female LMB while male expression was ~4 -fold higher than controls (2006) Repression of CYP19 expression in males may partially account for reduced circulating levels of E2 and gonadal production upon DDE exposure in the current study and may account for the g reater relative inhibition of E2 production versus T However the current data provide no indication of enhanced steroidogenesis, as would be suggested by Garcia -Reyeros finding of induced gonadal StAR expression. The finding of normal T plasma levels w hile gonadal p roduction was inhibited in male LMB may indicate systemic compensation to maintain normal T, i.e. reduced T metabolism. Conversely, diminished male plasma E2 concentration may be due to an inability to compensate for reduced gonadal synthesi s by inhibition of metabolic processes. In other studies, one month exposure to DDE decreased expression of CYP3A68 and 3A69 in females, while males exhibited a slight upregulation of CYP3A69. In the same work, a 4 -month dietary DDE exposure was capable of upregulating male hepatic CYP3A68 to a greater extent than females suggesting greater metabolic capacity in male LMB at that timepoint (Barber et al., 2007) which does not support the current findings. In addition, previous work has demonstrated that 120day exposure to 46 PPM DDE in the diet had no effect on hepatic E2 SUL T activity in male or female LMB
77 (D.S. Barber, personal communication, October 9, 2008) eliminating modulation of phase II sulfonation as an explanation for maintenance of normal circulating T levels in males. In males, the currently observed trends for changes in gonadal E2 and T product ion following 60-day dietary DDE exposure are similar to changes induced by 60 day dietary exposure to EE2, whereas FLUT had an opposite influence on gonadal production. This may suggest the DDE is acting as a xenoestrogen in male LMB upon 60-day dietary exposure. Certainly DDE has been shown to possess estrogenic properties in a number of other studies (Department of Health and Human Services, 2002b) However, in female LMB, g onadal steroid production was unchanged from control upon dietary DDE exposure. This effect is similar to that observed for ovarian T production upon EE2 administration, however EE2 exposure in females significantly curtailed E2 production in the current study, an effect not mirrored by DDE exposure. These data therefore indicate male LMB may be more sensitive to the endocrine disrupting effects of DDE versus females during the reproductive season. A number of investigations have detected differential responses to OCP exposure between sexes. Differential regulation of genes involved in steroidogenesis, steroid me tabolism, and encoding ERs has been detected in male versus female LMB (Barber et al., 2007; Garcia -Reyero et al., 2006) DDE was the only compound evaluated which affected HSI, elevating it in both sexes (78 and 53% from control in male and female LMB, respectively) Since circulating E2 levels were not elevated, an E2 mediated influence on HSI is unlikely. DDE, known to be mildly estrogenic, may induce VTG production with an associated increase in HSI, and presence of pl asma VTG in male animals is a biomarker of xenoestrogen exposure (Denslow et al., 1997) However, DDE is also known to bind the constitutive androstane receptor (CAR) in mammalian species and upregulate CYP2B inducing hepatomegaly (Wyde et al., 2003) Although a CAR has not been
78 identified in fish species (Hinton et al., 2008) the possible interaction of DDE with CAR or another nuclear receptor mediating CYP expression cannot be dismissed (such as the pregnane X recep tor). Dietary Exposure to Methoxychlor In vivo MXC exposure (7.46 PPM in feed) for 60 days elicited effects that were similar to those of DDE with one exception. While DDE exposure did not elicit a reduction in circulating T levels in male LMB, MXC reduc ed plasma T to the lowest concentration of all compounds studied (34% depression from control) HTPE, the MXC metabolite, has been shown to inhibit CYP11A activity in mammalian systems, and thus could account for the inhibitory effects on testis productio n by directly inhibiting gonadal production (Akgul et al., 2008) In support of this concept, current data has demonstrated the ability of d irect exposure of LMB ovarian tissue to MXC to inhibit steroidogene s is in a dose -dependent fashion. However, no effect on ovarian E2 production upon dietary MXC exposure for 60 days was observed. These data suggest a differential response of male and fem ale LMB to MXC exposure, similar to that currently observed upon exposure to dietary DDE. Importantly, no analysis of MXC metabolism was performed in the present study although f ish species are known to rapidly metabolize MXC in vivo to compounds which have been suggested to have greater endocrine -disrupting potential than MXC (Department of Health and Human Services, 2002c; James et al., 2008) While differential biotransformation of MXC to bioactive metabolites between sexes ha s not been demonstrated in fish species, this phenomenon may account for effects observed in male versus female LMB. MXC, like DDE, has been shown in a number of studies to possess estrogenic properties (Blum et al., 2008; Department of Health and Human Services, 2002c) and fittingly, in males, produce s effects on gonadal production of E2 and T similar to those observed upon dietary
79 administration of EE2 Also similar to DDE exposure, gonadal steroid production in males was dissimilar to that induced by FLUT administration via the diet suggesting a la ck of significant anti androgenic influence However, as observed with DDE, the effect of MXC on circulating T and E2 in male LMB did not mirror effects of EE2. Thus while MXC may interfere directly with gonadal steroidogenesis, the current data may also suggest the possibility of an inhibitory effect on male gonadal CYP19, since reduction of gonadal E2 synthesis was greater than reduction of T DDE, structurally similar to MXC, has been shown to repress expression of CYP19 upon dietary exposure in male LMB as discussed above (Garcia -Reyero et al., 2006) and thus MXC or on e of its metabolites may act via a similar mechanism to inhibit production of E2 and circulating E2 levels Effects of OCP Exposure in a Contaminated Environment on LMB Gonadal S teroidogenesis Overview These experiments showed that LMB exposed in the wild at a contaminated site for a relatively short duration of 2 months accumulate significant levels of OCPs within gonadal tissue during the reproductive season, illustrating the opportunity for these compounds to elicit direct e ffects on gonad function. S teroidogenic potential of gonadal tissue from exposed male and female LMB was examined by measuring the amount of E2 synthesized ex vivo in 20 hours. While no differences in production by ovary were detected among exposure grou ps, statistically different levels of production by testes were observed. Control males exhibited E2 production similar to animals sampled from the Lake Apopka mesocosm on 4/1/2008, while males from the earlier mesocosm sampling (1/15/08) produced signifi cantly more E2 under both ba sal and stimulated conditions.
80 Effects in Male LMB E2 production by male LMB gonad may be the influence of seasonal variation in response to OCPs. Ci rc ulating levels of E2 have been shown to fall slightly from January to Apri l in male LMB in the wild, which may account for the elevated E2 production by mesocosm fish sampled on 1/15/08 compared with the later mesocosm sampling and control animals (Gross et al., 2008) however the large differences currently no ted seem to suggest otherwise. T he effect of season on response of LMB to OCP exposure has been noted in previous st udies, with exposures in late spring having less influence on circulating hormone levels compared to exposures in the fall (Johnson et al., 2007) Thus the exaggerated synthesis of E2 by male gonad in exposed fish collected during January may represent a maturation sta ge -dependent influence of OCPs on gonadal steroidogenesis. Effects in Female LMB Variability in the average follicle size of female LMB which is related to maturation stage, was noted. To compensate for differing stages of reproductive maturation in females, ex vivo E2 production (as percent change from basal upon stimulation with hCG) was normalized to the mean follicle diameter for each fish. While significant differences between exposure groups were not present influenc e of mean follicle diameter on response to hCG stimulation was observed. The most sensitive follicles were of larger diameter ( in E2 production exceeding 200% for these animals, while ovaries containing smaller follicles wer e relatively unresponsive. It is possible that this effect is due to specific sensitivity of larger or more mature follicles to hCG stimulation. In mammals, hCG binds most strongly to the luteinizing hormone/choriogonadotropin receptor (LHCGR ) and is responsible for maintenance of the corpus luteum and associated progesterone production required to sustain fetal growth. Cross activation between GtH receptors in humans occurs only at very high concentrations of
81 ligand (Fradkin et al., 1989; Vassart et al., 2004) In fish, the function of LH receptor activation is thought primarily to be induction of final oocyte changes required for ovulation, although promiscuous binding by FSH and LH o ccurs in many species (Bogerd et al., 2005) hCG, however, is thought to bind most strongly to LH receptors in fish species and has been used successfu lly for years to induce maturation and ovulation in fish. Inyang and Hettiarachchi discovered that the efficacy of hCG to promote ovulation was dependent on follicle diameter in the African catfish Clarias gariepinus and Clarias anguillaris and the dosag e of hCG required to induce ovulation was inversely proportional to oocyte diameter (1994) In concordan ce, the current data indicate greater responsiveness of larger follicles in LMB to stimulation by hCG. It has been suggested that regulation of oocyte maturation and ovulation in asynchronously spawning species such as LMB involves differential expression of FSH and LH receptors as opposed to rapidly fluctuating GtH levels in the plasma, such as occurs in synchronously spawning species (Kobayashi et al., 2008; Kwok et al., 2005) Thus the present data may suggest a receptor regulated maturation mechanism in LMB. Very recently, seasonal expression data of the GtH receptor genes in LMB was collected at the University of Florida which supports this postulation ( C.J. Martyniuk, personal communication April 10, 2009). The inability to detect differences in gonadal E2 synthesis between exposed and unexposed females may reflect of number of influences. First, for all of these experiments, gonad was excised on site and placed in cool media for transport back to the laboratory, a trip of approximately 2 hours. As a consequence, disturbance during transport and simple time delay may have affected gonadal tissue. Secondly, control animals were collected from an independent site and on a date separated by up to 2 months from mesocosm collections, introducing confounding variables from seasonal variation in reproductive stage Lastly, while
82 follicle maturation stage was quasi -normalized by mean follicle diameter, histological staging was not performed to de termine the precise stage of follicle maturation, which may also influence response to endocrine disrupting compounds (Kime, 2000) Conclusions and Future Directions These data indicate a direct role for OCPs in the perturbation of gonadal sex steroid production and homeostasis in the LMB. Two OCPs, MXC and TOX, elicited d irect inhibitory effects on ovarian E2 production by what appear to be different mechanisms The ability of OCPs to elicit direct effects on gonadal steroidogenic tissue is relevant to the LMB model since, as demonstrated by the current data, OCPs accumulate in these tissues both at contaminated sites and upon environmentally relevant dietary exposure in the laboratory. These experiments showed that d ietary exposure to OCPs during the reproductive season is capable of perturbing both gonadal steroid production and circulating levels of sex steroids in LMB. However, the effect of each OCP on the LMB reproductive endocrine system seems to be unique and male and female animals are affected differently Some e ffects on gonadal steroid production induced by direct OCP exposure ex vivo were mirrored upon dietary exposure yet t he mechanisms of OCP induce d endocrine disruption in vivo are likely complex and involve more than direct gonadal inhibition alone, as plasma steroid concentrations did not always correlate with gonadal production. Additional possible effects induced by dietary OCP exposure may inc lude direct actions on other levels of the HPGa such as hypothalamic synthesis and release of GnRH and pituitary secretion of GtHs. In addition, compensatory mechanisms in vivo may ameliorate the influence of some OCPs. These mechanisms could include pos itive or negative feedback regulation of the hypothalamus and pituitary changes in GtH or estrogen and androgen receptor expression, and possible modulation of metabolic processes which are responsible for biotransformation and elimination of sex hormones.
83 The current investigations showed no significant effect of natural exposure to OCPs in a contaminated site on LMB ovarian steroidogenesis whil e a possible influence on male gonadal production was noted. These data should be interpreted carefully, however, as collection times from the contaminated and control sites varied by months, which may confound comparison of data points. Several other va riables may have contributed to uncertainty in these data, including the presence of other industrial and agricultural com pounds at the contaminated site which may also accumulate in LMB tissue. It was for these reasons that controlled laboratory investig ation of exposure to individual OCPs in the diet which mimicked environmental ly bioavailable levels was performed. Further experiments which may help elucidate mechanistic details of endocrine disruption by OCPs in LMB are many. Evaluation of the activi ty of individual enzymes involved in the steroidogenic cascade in gonadal tissue upon exposure to OCPs would identify specific points of uncoupling of the steroidogenic process. Assays evaluating activity of hepatic enzymes involved in steroid elimination would add a level of detail to what is currently known regarding changes in expression of these enzymes following in vivo OCP exposure. Analysis of VTG levels in the plasma of fish exposed to dietary OCPs may provide clues regarding estrogenic effects of these compounds, as would histological examination of hepatic tissue. Finally, co -exposure of gonadal tissue to OCPs and agonists/ anta gonis ts of estrogen and androgen receptor s may clarify possible effects mediated via these hormone receptors
84 0.0 0.5 1.0 1.5 2.0 2.5 3.0 17-Estradiol (pg/mg gonadal explant) DMSO 100uM 50uM 25uM 10uM 2.5uM Figure 4 1 E2 produced under basal conditions in 20 hours by LMB ovarian explants exposed to increasing concentrations of TOX in culture media. Each bar represents E2 production as pg/mg explant of 2 individual explants from 3 LMB
85 0 5 10 15 20 Basal hCG Stimulation 17-Estradiol (pg/mg gonadal explant) DMSO 100uM 50uM 25uM 10uM 2.5uM A B 0.0 0.5 1.0 1.5 2.0 2.5 3.0 Basal hCG Stimulation 17-Estradiol (pg/mg gonadal explant) DMSO 100uM 50uM 25uM 10uM 2.5uM Figure 4 2 E2 produced in 20 hours by LMB ovarian explants exposed to increasing concentrations of TOX in culture media. A) Ovarian production from two LMB with mean follicle diameters of 605 and 670 m Each bar represents mean production from 2 individual explants from each fish. B ) O varian production from one LMB with mean follicle diameter of 917 m Each bar represen ts mean production of 2 individual explants. 550 600 650 700 900 910 920 930Percent Increase from Basal -100 0 100 200 300 400 500 Vehicle 100 M TOX 550 575 600 625 650 675 700 -40 -20 0 20 40 60 80 Vehicle 100 M MXC Mean Follicle Diameter (m) A B Figure 4 3 Percent increase in E2 production by LMB ovarian explants upon stimulation with hCG during a 20 -hour incubation. A) Stimulated E2 production with and without exposure to 100 M TOX versus mean follicle diameter. B) Stimulated E2 production with and without exposure to 100 M MXC versus mean follicle diameter.
86 APPENDIX A EXTRACTION OF ORGANI C COMPOUNDS FROM FISH MUS CLE TISSUE Muscle was excised from LMB carcasses dorsally just posterior of the opercle and dermal tissue was removed. Organic extraction of muscle samples was accomplished utilizing a 2 -day procedure. 5 6 g of muscle from each animal were frozen in liquid nitrogen and pulverized using a BioSpec BioPulverizer cryopulverization devic e (BioSpec Products, Inc., Bartlesville, OK) and transferred to a Teflon -capped glass extraction vial. 7.0 mL HEX was added to each vial along with 100 L of internal standard solution containing 3 PPM PHEN (SPEX CertiPrep ) and 9 PPM 13C12-DDE ( Cambridge Isotope Laboratories, Inc. ) in cyclohexane (HPLC grade, Fisher Scientific). Primary extraction was performed by vortex mixing at high speed for 60 seconds. Samples were then placed in a bath sonicator for 30 minutes, centrifuged at 500 x g for 5 minutes, and the top organic layer was transferred to a borosilicate glass vial which was sealed and stored at 4oC. An additional 7.0 mL HEX was added to each extraction vial, mixed by vortex for 60 seconds, and placed on an over the top rotator overnight. The following morning, samples were again centrifuged as above and the organic layer was added to that collected on the first day of extraction. The combined organic extract was evaporated under a gentle stream of N2 at 37oAd ditional sample clean up was performed using solid -phase chromatography columns. Extract in ACN plus a 1.0 mL ACN rinse of the extract tube was eluted through a pre conditioned C18 column (500 mg, 6.0 mL; Agilent Technologies). The column was washed with 500 L ACN once all extract had eluted. The combinati on of primary eluant, tube rinsate and column rinsate was eluted through a pre conditioned NH2 column (500 mg, 3.0 mL; Varian, Inc.) followed by a 1.0 mL ACN rinse of the tube and 500 L wash of the c olumn. Combined eluant was dried under a gentle stream of N C an d reconstituted in 3.0 mL ACN. 2 at 37oC. The dried extract was reconstituted in
87 cyclohexane and transferred to an amber HPLC vial followed by evaporation and a final reconstitution in 100 L cyclohexane.
88 APPENDIX B VALIDATIO N OF 11 -KETOTESTOSTERONE ENZYME IMMUNO ASSAY Sample Preparation and Analysis Samples of charcoal -stripped LMB plasma were spiked with 11 -KT to yield final concentrations of 100, 10 and 1 pg/mL. Standard solutions of 11-KT were created using stock 11KT solution included in a commercially availabl e enzyme immunoassay (EIA) kit (catalog no. 582751, Cayman Chemical Company, Ann Arbor, MI) per manufacturer instructions to yield 100, 50, 25, 12.5, 6.25, 3.13, 1.56, and 0.78 pg/mL in sample buffer (50 mM sodium phosphate, 0.1% gelatin, 0.1% RIA grade Fraction V bovine serum albumin, pH 7.6). 125 L of each sample or standard were added to a borosilicate tube along with 50 L of assay buffer (50 mM sodium phosphate 0.1% gelatin, pH 7.6) and 62.5 L of 50 nCi/m L tritiated T (specific activity of 73 Ci/mmol; Amersham Radiochemicals, now GE Healthcare Bio -Sciences Corp ) to enable evaluation of extraction recovery. All samples were extracted twice with 1.0 mL of 1 chlorobutane (99%+ pure, Acros Organics) by vorte x mixing for 30 seconds followed by centrifugation at 500 x g for 5 minutes to separate hydrophilic and hydrophobic phases. The organic (top) phase was transferred to a new borosilicate tube, and the combined extract from each sample was dried under a gen tle stream of nitrogen at 37oC. Extracted samples were reconstituted overnight at 4oAll reconstituted samples were analyzed in duplicate via EIA (catalog no. 582751, Cayman Chemical Company ). In addition, unextracted kit 11-KT standards were loaded onto the EIA plate. T he stock solution of tritiated 11 -KT was analyzed at 1:1 and 1:10 dilutions to determine if antibody binding of the tritiated steroid is comparable to nonlabeled 11 -KT, and the stock solut ion of tritiated T was analyzed to evaluate cross reactivity with the 11 -KT antibody used in the EIA All manufacturer instructions were followed, and the EIA plate was developed with C with 125 L sample buffer
89 Ellmans Reagent for 60 minutes prior to analysis by a plate spectroph otometer (Sp ectraMax 250, Molecular Devices. ) at 412 nm. Results and Validation The standard curve generated by plotting the absorbance at 412 nm against the log of concentration of 11-KT in each ex t racted standard solution yielded an R2These results indicate the Cayman 11-KT EIA kit performs very well when analyzing concentrations of 11-KT in LMB plasma following organic extraction procedures. In addition, the current stock of tritiated 11-KT solution reac ts with Cayman 11 -KT antibodies similarly to unlabeled 11-KT, while tritiated T does not cross react to an appreciable degree. value of 0.999. U sing this standard curve, the samples of charcoal -stripped LMB plasma spiked with 11 -KT yielded 0.67, 6.74, and 104.56 pg/mL for samples intended to be 1, 10, and 100 pg/mL following subtraction of the value ascertained for the charcoal -stripped plasma al one. Extraction recoveries for plasma samples spiked with 11 -KT was calculated by transferring 10 uL of the reconstituted samples to a 5 -mL liquid scintillation vial t o which 5.0 mL of scintillation fluid was added (Sc intiSafe 30%, Fisher Scientific ). Sa mples were counted with a Beckman LS 6000IC scintillation counter in the tritium window for 3 minutes (Beck man Coulter, Inc. ). Calculated extract ion recoveries for spiked plasma samples ranged from 91 to 101% However, extracted standard solutions were not spiked with tritiated T solution, therefore recoveries were not used to correct observed values of 11-KT in extracted samples. Concentrations determined for the stock solution of tritiated 11 -KT were 107.72 and 15.77 pg/mL for the 1:1 and 1:10 dilutions, respectively. Finally, the stock solution of tritiated T resulted in a very small cross reactiv ity, indicating a value of 2.32 pg/mL.
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BIOGRAPHICAL SKETCH Nicholas John Doperalski grew up in rural Wisconsin, where he enjoyed long hours in the outdoors. Following completion of his Bachelor of Science degree, Nick relocated to Madison, Wisconsin where he worked in the field of neuroscience for 3 years. He then migrated south to Florida where he spent another 3 years working in neuroscience, more specificall y respiratory neurophysiology. In 2007, he decided to pursue graduate school and began wo rk on his Master of Science in veterinary medical s ciences in the lab of David S. Barber at the Center for Environmental and Human Toxicology. He successfully defended his Master of Science work in April of 2009.