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An Evaluation of dosing methods and effects of p,p'-DDE and dieldrin in Florida largemouth bass (Micropterus salmiodes f...

University of Florida Institutional Repository
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PAGE 1

AN EVALUATION OF DOSING METHODS AND EFFECTS OF p,pDDE AND DIELDRIN IN FLORIDA LARGEMOUTH BASS ( Micropterus salmoides floridanus) By JENNIFER KEENE MULLER A THESIS PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLOR IDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE UNIVERSITY OF FLORIDA 2003

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Copyright 2003 by Jennifer K. Muller

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ACKNOWLEDGMENTS I would like to extend my gratitude to Dr. Timothy S. Gross for providing resources, funding, staff, and guidance over the course of my studies. This research could not have been accomplished without the aid of the entire Ecotoxicology Lab at USGS-BRD-CARS (Gainesville, FL). Special thanks go to Carla Weiser, Nikki Kernaghan, Shane Ruessler, Jon Weibe, Janet Buckland, Travis Smith, Jessica Grosso and my fellow graduate students Jessica Noggle, Eileen Monck, Kevin Johnson, Heath Rauschenberger, and Sekeenia Haynes. I would like to thank my other committee members, Dr. Maria Seplveda and Dr. David Barber, for their invaluable assistance in revising this thesis. Their comments greatly improved my writing. Also, I would like to acknowledge Dr. Christopher J. Borgert for study design and editorial assistance. This research was made possible through a grant to Dr. Gross and Dr. Borgert from the American Chemistry Council. Lastly, I would like to thank my husband, Josh, and my family for all of the love, support, and most of all, patience they have given me over the past two years. My life and my ability as a scientist benefits from their presence. iii

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TABLE OF CONTENTS page ACKNOWLEDGMENTS.................................................................................................iii LIST OF TABLES............................................................................................................vii LIST OF FIGURES.........................................................................................................viii ABSTRACT......................................................................................................................xii CHAPTER 1 INTRODUCTION........................................................................................................1 Background...................................................................................................................1 Upper Ocklawaha River Basin..............................................................................1 Chemical Contamination in the Upper Ocklawaha Basin.....................................2 Selection of Animal Model...........................................................................................3 Largemouth Bass Reproductive Cycle.........................................................................4 OCP Contaminant Effects on Fish Reproduction.........................................................6 Research Significance...................................................................................................9 2 EXPOSURE OF LARGEMOUTH BASS TO p,p-DDE AND DIELDRIN BY SLOW-RELEASE MATRIX PELLET......................................................................11 Introduction.................................................................................................................11 Specific Aims......................................................................................................12 Null Hypothesis...................................................................................................12 Alternative Hypothesis........................................................................................12 Materials and Methods...............................................................................................12 Experimental Animals.........................................................................................12 Experimental Design...........................................................................................13 Chemicals and Dosing.........................................................................................13 Blood and Tissue Collection...............................................................................14 Gonad Histology..................................................................................................15 Analysis of Circulating Sex Steroid Hormones...................................................15 Analysis of Largemouth Bass Tissues for OCPs.................................................16 Analysis of Largemouth Bass Blood Plasma for OCPs......................................17 Statistical Analysis..............................................................................................19 Results.........................................................................................................................20 iv

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Largemouth Bass.................................................................................................20 Health Parameters: K and HSI............................................................................20 Gonad Histology..................................................................................................20 Reproductive Parameters: GSI and Circulating Sex Steroids.............................20 GSI...............................................................................................................20 17-Estradiol................................................................................................21 11-Ketotestosterone......................................................................................21 In Vivo Treatment Dosing Consistency...............................................................22 Health Parameters: K and HSI (Regression).......................................................24 Reproductive Parameters: GSI and Circulating Sex Steroids (Regression)........25 GSI...............................................................................................................25 17-Estradiol................................................................................................25 11-Ketotestosterone......................................................................................26 Discussion...................................................................................................................26 3 ACCUMULATION OF DIETARY p,p-DDE AND DIELDRIN BY LARGEMOUTH BASS: A PILOT STUDY..............................................................63 Introduction.................................................................................................................63 Specific Aims......................................................................................................64 Null Hypothesis...................................................................................................64 Alternative Hypothesis........................................................................................65 Materials and Methods...............................................................................................65 Experimental Animals.........................................................................................65 Chemicals and Feed.............................................................................................66 Feeding Rate........................................................................................................67 Day 30 and Day 50 Endpoints.............................................................................67 Analysis of Largemouth Bass Tissue for OCPs..................................................68 Analysis of Circulating Sex Steroid Hormones...................................................68 Statistical Analysis..............................................................................................69 Results.........................................................................................................................69 Largemouth Bass.................................................................................................69 Feeding Rate........................................................................................................70 30-Day Exposure.................................................................................................70 p,p-DDE accumulation..............................................................................70 Dieldrin accumulation..................................................................................71 GSI and sex steroids.....................................................................................72 50-Day Exposure.................................................................................................72 p,p-DDE accumulation..............................................................................73 Dieldrin accumulation.................................................................................74 GSI and sex steroids.....................................................................................74 Discussion...................................................................................................................75 4 GENERAL CONCLUSIONS / FUTURE RESEARCH............................................93 General Conclusions...................................................................................................93 Future Research..........................................................................................................95 v

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LIST OF REFERENCES...................................................................................................96 BIOGRAPHICAL SKETCH...........................................................................................105 vi

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LIST OF TABLES Table page 2-1 GC-MS results of tissue samples for bioaccumulation measurements at Day 30 of exposure. Blue and green numbers highlight fish treated with p,p-DDE and dieldrin, respectively................................................................................................46 3-1 Ten-day increment totals of feed eaten per tank. The total feed eaten by each tank was totaled at Day 30 and used to estimate the average amount of pesticide ingested by each fish................................................................................................82 3-2 Ten-day increment totals of feed eaten per tank (floating only after Day 30). The total feed eaten by each tank was totaled at Day 50 and used to estimate the average amount of pesticide ingested by each fish..................................................83 vii

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LIST OF FIGURES Figure page 2-1 Female body condition over time. Box plots of weight (a), length (b), and condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 63). Box plot contains the 25 th to the 75 th quartile, line in box indicates the median, whiskers extend to the minimum and maximum value............................................36 2-2 Male body condition over time. Box plots of weight (a), length (b), and condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 93). Box plot contains the 25 th to the 75 th quartile, line in box indicates the median, whiskers extend to the minimum and maximum value............................................37 2-3 Day 60 HSI. There was no significant difference in HSI among treatment group for females (a). Males in the Dieldrin 1.0 mg treatment group showed a significant increase in HSI over Sham and DDE 2.5 mg treatments (b). Sample size per treatment ranged from 7 to 11 for females and 7 to 16 for males. Treatments with the same lower case letter were not significantly different...........38 2-4 Histological section of stage 4 vitellogenic female gonad viewed at 40X. CA = cortical alveoli, GV = germinal vesicle, YV = yolk vesicle....................................39 2-5 Day 60 GSI. There was no significant difference in GSI among treatment group for females (a) or males (b). Sample size per treatment ranged from 7 to 11 for females and 7 to 16 for males. Treatments with the same lower case letter were not significantly different................................................................................40 2-6 Female circulating estradiol at Days 0 (a), 30 (b), and 60 (c). Placebo and DDE 2.5 treatments showed consistently higher plasma E 2 concentrations than all other treatments. Sample size per treatment ranged from 7 to 11. Treatments with the same lower case letter were not significantly different..............................41 2-7 Male circulating estradiol at Days 0 (a), 30 (b), and 60 (c). DDE 5.0 mg treatment showed a consistently higher mean than all other treatments. Sample size per treatment ranged from 7 to 16. Treatments with the same lower case letter were not significantly different.......................................................................42 2-8 Female circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Treatment groups demonstrated high variability compared to controls. By Day 60, no difference existed between treatment groups and controls. Sample size per treatment viii

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ranged from 7 to 11. Treatments with the same lower case letter were not significantly different...............................................................................................43 2-9 Male circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Placebo and DDE 2.5 mg treatment groups demonstrated consistently higher 11-KT. At Day 30 and 60 all dieldrin treatments and higher DDE treatments demonstrated decreased 11-KT. Sample size per treatment ranged from 7 to 16. Treatments with the same lower case letter were not significantly different............................................44 2-10 Day 60 mean female gonadal dose. p,p-DDE (a) and dieldrin (b) concentration SD for each treatment group (n = 6) compared to the target dose for each treatment. Also included is the mean gonadal concentration of either OCP found in the Eustis property of EMCA.................................................45 2-11 Regression analysis of gonad (a, d), liver (b, e) and muscle (c, f) DDE concentrations against blood DDE concentrations in females (a c; n = 9) and males (d f; n = 6). Significant linear relationships exist between blood plasma and tissue DDE concentrations, except in male muscle...........................................47 2-12 Regression analysis of gonad (a, d), liver (b, e), and muscle (c, f) dieldrin concentrations against blood dieldrin concentrations in females (a c; n = 4) and males (d f; n = 6). Significant linear relationships exist between blood plasma and tissue dieldrin concentrations, except in female muscle...................................48 2-13 Linear regression of ELISA DDE (a) or dieldrin (b) results of pooled blood samples against GC-MS results (n = 12). A highly significant and linear relationship exists between the two methods...........................................................49 2-14 Linear regression of HSI against blood plasma DDE concentrations. No significant correlation was found for females (a, n = 37) or males (b, n = 53)........50 2-15 Linear regression of HSI against blood plasma dieldrin concentration. No significant correlation was found for females (a, n = 34) or males (b, n = 51)........51 2-16 Linear regression of GSI against blood plasma DDE concentration. No significant correlation was found for females (a, n = 37) or males (b, n = 53)........52 2-17 Linear regression of GSI against blood plasma dieldrin concentration. No significant correlation was found for females (a, n = 34) or males (b, n = 51)........53 2-18 Day 0 plasma hormone concentrations. Mean SD (a) female, n = 59 (b) male, n = 88........................................................................................................................54 2-19 Day 30 Hormones Female p,p-DDE treated (n = 30). Linear regression against DDE dose (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30. Significant negative relationships exist between DDE dose and circulating E 2 and percent change in E 2 .............................55 ix

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2-20 Day 30 Hormones Male p,p-DDE treated (n = 47). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to day 30 against DDE dose. No significant relationships were found...................56 2-21 Day 60 Hormones Female p,p-DDE treated (n = 35). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 30 to Day 60 against p,p-DDE dose and (c) ciculating E 2 to GSI. A significant negative relationship was found between DDE dose and circulating E 2 .................57 2-22 Day 60 Hormones Male p,p-DDE treated (n = 51). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 30 to Day 60 against dieldrin dose and (c) circulating 11-KT against GSI. No significant relationships were found........................................................................58 2-23 Day 30 Hormones Female dieldrin treated (n = 34). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30 against dieldrin dose. No significant relationships were found..............59 2-24 Day 30 Hormones Male dieldrin treated (n = 42). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30 against dieldrin dose. A significant negative relationship exists between dieldrin dose and percent change in 11-KT from Day 0 to Day 30...........60 2-25 Day 60 Hormones Female dieldrin treated (n= 34). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 30 to Day 60 against blood plasma dieldrin and (c) circulating E 2 against GSI. A significant negative relationship between blood plasma dieldrin and circulating E 2 was found.............................................................................................................61 2-26 Day 60 Hormones Male dieldrin treated (n = 51). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30 against dieldrin dose and (c) circulating 11-KT against GSI. Significant negative relationships between circulating 11-KT and dieldrin dose, as well as, between percent change in E 2 from Day 30 to Day 60 were found.........................62 3-1 Tank setup for preliminary feeding study. Each set of four tanks had a divided well water and air supply line and separate drainage lines......................................81 3-2 Day 30 whole-body concentration SD of p,p-DDE (A) and dieldrin (B) in largemouth bass fed a contaminated diet (n = 3 per treatment). All dosed fish had significantly higher whole-body concentrations of their respective chemical than the controls. One replicate fed sinking p,p-DDE contaminated feed was significantly higher than all other p,p-DDE fed fish. There were no significant differences among replicates fed dieldrin contaminated diets. Bars with different letter designations indicate a significant difference..................................84 3-3 Day 30, percentage of total dose accumulated SD within each replicate (n = 3 per replicate). There were no significant differences in the amount accumulated x

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between replicates of either feed type for dieldrin. There was a significant difference between one DDE sinking diet replicate and one DDE floating diet replicate. Bars with different letter designations indicate a significant difference.85 3-4 Day 50 whole-body concentration SDof p,p-DDE (A) or dieldrin (B) in largemouth bass fed a contaminated diet (n = 3 per replicate). All dosed fish had significantly higher whole-body concentrations of their respective chemical than the controls. There was no significant difference among p,p-DDE replicates. Whole-body concentration was significantly different between dieldrin replicates. Bars with different letter designations indicate a significant difference..................................................................................................................86 3-5 Day 50, percentage of total dose accumulated SD within each replicate (n = 3 per replicate). There were no significant differences in the amount accumulated between replicates of either for either p,p-DDE or dieldrin. Bars with different letter designations indicate a significant difference.................................................87 3-6 Gonadal concentrations of OCPs. Replicates of p,p-DDE were not significantly different, however, there was a significant difference in gonadal concentrations of dieldrin between replicates most likely due to the difference in total g of dieldrin fed...............................................................................................................88 3-7 Day 50, Female circulating hormones SD (n ranged from 3 to 5). Significant differences in E 2 were found between both replicates of p,p-DDE and dieldrin fed fish and controls. No significant differences were found in 11-KT levels in females. Bars with different letter designations indicate a significant difference..89 3-8 Day 50, male circulating hormones SD (n ranged from 6 to 8). Significant differences were found in both E2 and 11-KT levels in all treatment groups and replicates as compared to controls. Bars with different letter designations indicate a significant difference...............................................................................90 3-9 Whole carcass concentration of floating style feed replicates over time, A p,p-DDE, B Dieldrin....................................................................................................91 3-10 Female and male GSI SD for Days 30 (n = 16 and 20, respectively) and 50 (n = 21 and 44, respectively). There was no significant difference in GSI between Days 30 and 50 when GSI should have increased dramatically. Bars with different designation indicate significant difference................................................92 xi

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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 AN EVALUATION OF DOSING METHODS AND EFFECTS OF p,p-DDE AND DIELDRIN IN LARGEMOUTH BASS (Micropterus salmoides floridanus) By Jennifer Keene Muller August 2003 Chair: Timothy S. Gross Major Department: Physiological Sciences Previous work indicates that persistent organochlorine pesticides (OCPs) have a wide range of effects on steroidogenesis and reproduction in fish. The present study investigated consistency of dose administration, dose accumulation and potential reproductive effects of two OCPs, p,p-DDE and dieldrin, on Florida largemouth bass (Micropterus salmoides floridanus). The first study used 60-day slow release pellets inserted into the intraperitoneal cavity. Twenty-five fish comprised each of eight treatment groups: sham; placebo (matrix only pellet); 3 p,p-DDE; and 3 dieldrin doses. Exposures were initiated before onset of the reproductive season (January) so that the full dose would be released prior to spawning (March). Weight, length, condition factor, gonadosomatic (GSI) and hepatosomatic (HSI) indices, and circulating hormones were measured. Preliminary contaminant analysis of gonadal tissue indicated that pellets did not release a consistent dose in each treatment. Statistical analyses, therefore, were more appropriately based on xii

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actual pesticide concentrations in blood plasma. No change in GSI, HSI, or any health parameter was observed in fish treated with either chemical indicating a lack of acute toxicity. Results showed a decrease in estradiol production in female bass with high blood plasma concentrations of both pesticides (p,p-DDE 200 1,000 ppb; dieldrin 100 500 ppb). Also, male bass treated with dieldrin, but not p,p-DDE, demonstrated a dose-dependent decrease in circulating 11-ketotestosterone concentration at a blood plasma concentration range of 100 600 ppb. Altered steroid synthesis may result in asynchronous spawning and reduced reproductive success seen in areas of the Ocklawaha River Basin, Florida contaminated with OCPs. The second study evaluated an oral exposure method for largemouth bass to OCPs. Feed pellets were dosed with 5 ppm p,p-DDE or 1 ppm dieldrin. Groups of 14 fish were fed sinking or floating diets to determine which feed produced a more consistent whole-body dose. Variability within replicates, total accumulated dose, and day 50 hormone concentrations were measured. Floating pellets delivered a more consistent dose. A steady state whole body concentration was achieved at approximately Day 30 and maintained through the end of the study at Day 50. Treated bass had depressed sex steroid concentrations at Day 50 even though whole body concentrations were lower than those found in natural environments in the Ocklawaha River Basin with high OCP levels. The research presented provides evidence of contaminated feed as a more appropriate exposure method for consistent OCP dosing in largemouth bass. Also, p,p-DDE and dieldrin in single chemical exposures can modulate endocrine function in largemouth bass. The two OCPs studied may contribute to the effects of a more complex mixture of chemicals found in the environment. xiii

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CHAPTER 1 INTRODUCTION Background Upper Ocklawaha River Basin The Upper Ocklawaha River basin in Central Florida has received much attention in the last 20 years because of declining habitats and wildlife populations. The historically hypereutrophic and contaminated Lake Apopka (located 25 km northwest of Orlando (Pollman et al., 1988)) serves as the rivers headwaters. From Lake Apopka, water flows north through a chain of lakes, including Lake Beauclair, Lake Dora, Lake Harris, Lake Eustis and Lake Griffin, before channeling into the Ocklawaha River. Much of the marsh and wetland area surrounding this chain of lakes was diked and drained for agricultural use or muck farming beginning in the 1940s and continuing through the 1980s (Benton et al., 1991). The removal of thousands of hectares of shallow lake bottom resulted in the loss of spawning habitat for Florida largemouth bass (Micropterus salmoides floridanus) and other sportfish, resulting in a decline of a once-thriving sportfishery prior to the 1950s. Huffstutler et al. (1965) determined that the population of largemouth bass in Lake Apopka was primarily adults and reproduction was not sufficient to maintain a thriving population. To add to the loss of habitat, the contamination of Lake Apopka resulted from inadequate sewage treatment discharge from the city of Winter Garden beginning in 1922, a major pesticide spill by a nearby chemical manufacturer in 1980, and ongoing agricultural sources that release nutrient and 1

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2 pesticide rich irrigation water into many lakes in the Ocklawaha chain (Pollman et al., 1988). In an attempt to restore the Upper Ocklawaha River basin to a semi-natural habitat, the state of Florida authorized the St. Johns River Water Management District (SJRWMD) to acquire muck farmland to flood and restore the marsh/wetland habitat. By 1993, the SJRWMD had acquired 2,630 hectares of muck farms adjacent to Lake Griffin (the Emeralda Marsh) and reflooded portions during 1992 to 1995 (Marburger et al., 1999). The objectives of this management practice were to reduce nutrient loading and chemical runoff of applied pesticides and fertilizers and to return the marsh to its natural habitat. However, flooding the land led to leaching of pollutants from the soil and their subsequent biomagnification in the food chain. Sport and forage fish were stocked in many flooded areas in 1992, 1993, and 1994, but with limited success (Benton and Douglas, 1996). Contributing factors included low dissolved oxygen, low water levels, and overabundant vegetation. Although the surviving largemouth bass had a high growth rate, reproductive success was poor (assessed by the number of young of year counted); therefore, a viable fishery could not be reestablished (Benton and Douglas, 1996). Chemical Contamination in the Upper Ocklawaha Basin Recent studies have begun to determine the extent of contamination in Lakes Apopka and Griffin and in an area of reclaimed muck farms on the north shore of Lake Griffin known as the Emeralda Marsh Conservation Area (EMCA). Lake Apopka exhibits elevated levels of organochlorine pesticides (OCPs) in soil, fish (ATRA, 1997; Marburger et al., 1999), and alligators (Heinz et al., 1991; Guillette et al., 1999a and 1999b). In 1997, a risk assessment was conducted for a muck farm on the north shore of

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3 Lake Apopka that showed soil concentrations ranging from 136 ppb for endrin (760 ppb for dieldrin) to 20,000 ppb for DDT (6,000 ppb for p,p-DDE) and 36,000 ppb for toxaphene (ATRA, 1997). A study at the EMCA showed soil concentrations of prominent OCPs (p,p-DDE, dieldrin, and toxaphene) to be over 3,000 ppb, 500 ppb and 40, 000 ppb, respectively, with a total OCP load of over 68,000 ppb (Marburger et al., 2002). The EMCA study also determined concentrations of OCPs in various tissues of several fish species. Black crappie (Pomoxis nigromaculatus) generally showed less total OCP accumulation than largemouth bass or brown bullhead catfish (Ameiurs nebulosus). Concentrations in largemouth bass ovary and fat reached over 4,000 ppb and 17,000 ppb, respectively for total DDT derivatives; 100 ppb and 700 ppb, respectively for dieldrin; and 4,000 ppb and 20,000 ppb, respectively for toxaphene (Marburger et al., 2002). Selection of Animal Model The demand for a sports-fishing industry and environmental relevance as an indicator species make the Florida largemouth bass an excellent model for this and several historical studies. Marburger et al. (2002) indicated largemouth bass as bioaccumulating large amounts of OCPs from the environment. This can be attributed to top predator status of largemouth bass and persistence of OCPs in the environment due to their lipophilicity. In addition to high contaminant concentrations, depressed sex steroids (17-estradiol (E 2 ) and 11-ketotestosterone (11-KT)), reversed sex steroid ratios, and reduced survival of fry have been observed in fish from these areas (Benton and Douglas, 1996; Marburger et al., 1999). Because largemouth bass are an important sport-fish in Florida and nationally, much effort is being placed into the restoration of a viable fishery in the EMCA and Upper Ocklawaha River Basin (Benton et al., 1991; Benton and

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4 Douglas, 1994; Marburger et al., 1999). For these reasons, the Florida largemouth bass was chosen as an animal model for laboratory studies of OCP exposure. Largemouth Bass Reproductive Cycle In Florida, largemouth bass reach sexual maturity at 1 year of age and generally at a length of 250 mm (Chew, 1974). They are synchronous spawners with a spawning season that can extend from January to April (Clugston, 1966). It is generally accepted that the act of spawning is triggered by an increase in water temperature in the spring to a range of 20 24C. Carr (1942) found that peak spawning occurs in the Gainesville, Florida area in late March. Teleost fish, such as the largemouth bass, follow a well-established reproductive cycle that is influenced by environmental (temperature and photoperiod) and endogenous (hormone) factors. These factors stimulate the hypothalamus to secrete gonadotropin-releasing hormone (GnRH), norepinephrine (NE) and other neuropeptides that act to stimulate the pituitary to secrete the primary teleost gonadotropins GTH-I and GTH-II (Van Der Kraak et al., 1998). GTH-I and GTH-II have varying roles but will generally stimulate gonadal sex steroid hormone production in preparation for spermiation in males and oocyte maturation, vitellogenesis, and ovulation in females. Increasing E 2 concentrations in females stimulates the liver to produce vitellogenin, a phosphoglycolipoprotein that serves as a yolk precursor in oviparous vertebrates (Wahli et al., 1981). Vitellogenin produced in the liver is released into circulation to travel to the gonad where it is sequestered as a nutrient source in developing oocytes. The vitellogenin gene has an estrogen responsive element in the promoter region and is transcribed in response to an estrogen-estrogen receptor (ER) complex (Wahli et al., 1981). Both males and females have copies of the vitellogenin gene, however it takes a

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5 threshold concentration of E 2 (only normally achieved in females) to produce measurable levels of vitellogenin (Wallace, 1970; Ryffel, 1978; Hori et al., 1979; Wahli et al., 1981; Copeland et al., 1986). Because males do not normally produce vitellogenin, it has been used in many studies as an indicator of exposure of males to estrogenic compounds (Pelissero et al., 1993; Sumpter and Jobling, 1995; Shelby et al., 1996; Monosson et al., 1996; Celius and Walther, 1998; Smeets et al., 1999; Madsen et al., 2002; Okoumassoun et al., 2002a and 2002b). A study performed by Gross, et al. (2002) on hatchery-reared bass in Florida characterized the annual cycles of circulating sex steroid hormones, vitellogenin, and gonad development. Circulating sex steroid hormones begin to rise prior to the spawning season. For male bass, 11-KT was the only sex steroid observed to show a strong seasonal pattern and had a peak concentration of about 2,800 pg/ml in February. E 2 was detected in males but at concentrations of about one-third of females. Female bass showed three distinctly seasonal hormones: E 2 testosterone (T), and progesterone (P). E 2 showed the strongest pattern with circulating levels nearly twice those of T at a peak concentration of 4,000 pg/ml in February. Circulating vitellogenin levels closely mimicked those of E 2 rising in November and peaking in January at about 6 mg/ml. Progesterone reached a peak concentration in early April corresponding with peak spawning activity, indicating a possible role in ovulation. 11-KT was detected in females, but at levels nearly one-half that of males. Similar peak concentrations of sex steroids were found in wild-caught female bass from Lake Woodruff, Florida (Timothy S. Gross, unpublished data).

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6 Corresponding gonad weight as a percentage of body weight or gonadosomatic index (GSI) begins to rise in October for both females and males, peaking in February to March at approximately 6% and 2%, respectively (Gross et al., 2002). Increasing GSI correlates with an increase in gonad maturation as seen histologically for both sexes (Nelson, 2001). In females, much of the dramatic increase in gonadal weight is due to the sequestering of a large amount of vitellogenin in developing oocytes (Wallace and Selman, 1981; Van Der Kraak et al., 1998). A similar study of Florida largemouth bass performed in Texas showed a peak GSI between March and April (Rosenblum et al., 1994). OCP Contaminant Effects on Fish Reproduction Although some reproductive parameters such as circulating sex steroids, gonad development, and reproductive success in fish (Marburger et al., 1999 and 2002; Gallagher et al., 2001) and American alligators, Alligator mississippiensis, (Gross et al., 1994; Guillette et al., 1994; 1995; 1999a; 1999b) have been shown to be abnormal in the contaminated areas of the Ocklawaha basin, there are few studies linking OCPs as a causative agent. Many OCPs are thought to be endocrine disrupters and much research has been focused in this area. Exposure to some OCPs by painting them onto the surface of red-eared slider turtle, Trachemys scripta elegans, (Willingham and Crews, 1999) and alligator (Gross et al., 1994) eggs showed that p,p-DDE can cause sex reversal, producing females when incubated at male producing temperatures. p,p-DDE has also been shown to be a potent anti-androgen in rats (Kelce et al., 1995) by binding to the androgen receptor (AR); therefore, preventing the transcription of testosterone resulting in demasculinization. Similar results have been found in some fish species such as

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7 guppies, Poecilia reticulata, (Baatrup and Junge, 2001; Bayley et al., 2002), white sturgeon, Acipenser transmontanus, (Foster et al., 2001), and goldfish (testes only), Carassius auratus, (Wells and Van Der Kraak, 2000). Alternatively, p,p-DDE can also bind to the ER in Hartley guinea pigs (Danzo et al., 2002), channel catfish, Ictalurus punctatus, (Nimrod and Benson, 1997), rainbow trout, Onchorhynkus mykiss (Matthews et al., 2000), and largemouth bass, (Garcia et al., 1997; Larkin et al., 2002). Other prominent OCPs such as methoxychlor, DDT, dieldrin, and endosulfan, are also able to bind to the ER of human, mouse (Mus musculus), chicken (Gallus gallus), green anole (Anolis carolinesis) and rainbow trout (Matthews et al., 2000; Tollefsen et al., 2002a). Matthews et al. (2000) demonstrated that methoxychlor, p,p-DDT, o,p-DDE, p,p-DDE, -endosulfan, and dieldrin are able to bind weakly to GST-ER fusion proteins from human, mouse, chicken and green anole, as well as, to displace all E 2 from GST-ER fusion proteins from a rainbow trout demonstrating a high binding affinity. The ability of any exogenous compound to bind a sex steroid hormone receptor and agonize and/or antagonize the action of an endogenous hormone can severely affect normal endocrine function and ultimately lead to a decrease in reproductive success. Normal androgen concentration and action are critical for development of the male fish gonad and spermatogenesis. Chemicals able to bind the ER can increase or decrease estrogen function. In males, even weak binding of a substance to the ER could induce an upregulation of estrogen function because males generally show only a low-level endogenous estrogen expression. This could lead to the production of vitellogenin in males, possibly coupled with blocked AR functions, potentially resulting in demasculinization (lowered testosterone level and testosterone-induced functions) and

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8 feminization (increased estrogen-like function). Vitellogenesis in cultured hepatocytes from male fish has been shown to occur with exposure to o,p-DDT ( Sumpter and Jobling, 1995; Smeets et al., 1999; and Leaos-Castaeda et al., 2002), tamoxifen, nonylphenol, bisphenol A (Sumpter and Jobling, 1995), and phytoestrogens (Pelissero et al., 1993). Some in vivo studies have shown vitellogenin production induced in male and female fish exposed to OCPs (Okoumassoun et al., 2002b; Madsen et al., 2002; Leaos-Castaeda et al., 2002). In females, competitive binding of the ER could reduce estrogen function, however, the exogenous compound will likely have a much lower potency than the endogenous estrogen. Inhibitory binding to the ER could lead to impaired gonadal development and decreased vitellogenesis, ultimately ending in poor egg quality and decreased reproductive success. Fish and alligators living in areas with high OCP levels demonstrate depressed circulating sex steroid levels (Guillette et al., 1995; Crain et al., 1997; Guillette et al., 1999a, 1999b; Marburger et al., 1999; Foster et al., 2001; Seplveda et al., 2002). This could be a result of a disruption in the hypothalamus-pituitary-gonad axis. No data was found on gonadotropin level changes in response to OCP exposure, however, studies on fish exposed to bleached kraft mill effluent have shown concomitant decreases in gonadotropin and sex steroid levels (Van Der Kraak et al., 1992; McMaster et al., 1995). Another possibility for OCP action includes inducing or inhibiting the activity of liver biotransformation enzymes that could increase/decrease the metabolism and excretion of any xenobiotics, as well as, endogenous hormones. An intensely used marker of induced cytochrome P450, a major class of phase I biotransformation enzymes and specifically CYP1A, is the activity of ethoxyresorufin-O-deethylase (EROD). High

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9 OCP levels can induce EROD in the white sturgeon (Foster et al., 2001), Nile tilapia (Oreochromis niloticus) (Zapata-Perez et al., 2000), rat (Wade et al., 2002), and deer mouse (Peromyscus maniculatus) (Dickerson et al., 1999). Dickerson et al. (1999) and Foster et al. (2001) also noted a decrease in circulating sex steroids in the deer mouse and white sturgeon, respectively, correlating with the increase in EROD activity. Also, a study on English sole (Parophrys vetulus) from Puget Sound, an area contaminated with PCBs and PAHs, showed significant negative correlations between mixed function oxygenase (MFO) induction and estradiol concentrations (Johnson et al., 1988). Because CYP1A is not the major isoform for metabolizing sex steroid hormones in fish, increased EROD activity is not likely related to the decreases in sex steroid concentrations (Snowberger and Stegeman, 1987). On the other hand, CYP3A is known to hydroxylate sex steroids in fish (Kullman et al., 2000). While no studies were found that had tested CYP3A induction/inhibition in fish exposed to pollutants, studies have shown it is induced in human hepatocytes and rats exposed to OCPs (You et al., 1999; Coumoul et al., 2002). More research is needed to elucidate the mechanism behind the correlation of biotransformation enzyme induction and decreased sex steroid levels. Increased metabolism and clearance of sex steroids may be due to the heightened activity of major metabolizing enzymes. Research Significance This research aims to provide information on dosing efficiency and the reproductive response of largemouth bass to p,p-DDE and dieldrin exposure, two predominant OCPs found in the Upper Ocklawaha River Basin. Single chemical exposures were performed to assess the potential contribution of each pesticide to overall reproductive function and steroidogenic declines known to occur when fish are

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10 environmentally or experimentally exposed to contaminant mixtures. Endpoints were measured on multiple levels of biological organization (biochemical, organ, and whole body) to gain insight into the ecological relevance of the response measured.

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CHAPTER 2 EXPOSURE OF LARGEMOUTH BASS TO p,p-DDE AND DIELDRIN BY SLOW-RELEASE MATRIX PELLET Introduction The most prominent organochlorine pesticides (OCPs) found in soils of the Ocklawaha River Basin are DDT derivatives, toxaphene, and dieldrin (Marburger et al., 2002). These chemicals bioaccumulate in Florida largemouth bass (Micropterus salmoides floridanus) and other top predators, such as the alligator (Alligator mississippiensis), and are suspected causes for decreased reproductive success of these animals due to impaired endocrine function (Gross et al., 1994; Guillette et al., 1994; 1995; 1999a; 1999b; Gallagher et al., 2001). p,p-DDE and dieldrin were selected to characterize single chemical dose-response relationships for reproductive function in Florida largemouth bass. Characterization of the reproductive effects of two major OCP contributors to this system will aid in determining if it is the actions of either of these chemicals alone that is a causative agent for depressed sex steroid concentrations and will also assist in future studies determining their contribution to the larger complex mixture that exists in the area. The objectives of this study were to evaluate the slow-release pellet dosing method and to determine potential dose effects of p,p-DDE and dieldrin on reproductive function in largemouth bass. Health and reproductive parameters were measured on multiple levels of biological organization: organism (weight, length, and condition factor (K)), organ (gonadosomatic index (GSI) and hepatosomatic index (HSI)) and biochemical 11

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12 (circulating hormones). This diversity of endpoints should provide insight into the biological significance of the doses tested. Specific Aims 1. Create internal doses of either p,p-DDE or dieldrin in largemouth bass prior to the onset of spawning season and evaluate the actual dose achieved. 2. Evaluate the effects of p,p-DDE or dieldrin dose on health (HSI and K) and reproductive (GSI and circulating sex steroid hormones) biomarkers. 3. Determine the biological relevance of the dose response. Null Hypothesis Reproductive and health parameters will not vary across treatments regardless of pesticide or dose administered. Alternative Hypothesis Because these chemicals have been shown to be endocrine active and wildlife exposed to them exhibit decreased circulating hormone concentrations, it is hypothesized that exposure of largemouth bass to p,p-DDE or dieldrin in a laboratory setting will also cause a decrease in circulating hormones which, in turn, will alter gonadal development, resulting in a decreased GSI. The doses administered were within the range of what is found in large, seemingly healthy adult bass in the wild. Therefore, no effects on weight, length, or K were predicted. Materials and Methods Experimental Animals Florida largemouth bass of one to two years of age were obtained from a fish hatchery (American Sports Fish, AL) in December 2001. The fish were transferred to the United States Geological Survey Biological Resources Division Center for Aquatic Resource Studies (USGS-BRD-CARS) facility where they were housed in 6,116 liter

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13 concrete runs (366 cm x 183 cm x 91 cm) equipped with a flow-through system supplied by on-site pond water and aeration. On Day 0 of the experiment (January 11, 2002) bass had a mean SD weight, length, and K of 141.9 22.6 g, 213.1 10 mm, and 1.46 0.11, respectively, indicating fish were healthy and of reproductive size. Water quality was measured as dissolved oxygen, temperature, and pH twice a week. Water quality parameters were all within acceptable ranges for the duration of the experiment: dissolved oxygen ranged from 4.28 to 10.20 mg/L (measurements were taken at 8 a.m. when oxygen levels would be expected to be at the lowest), temperature ranged from 11.7 to 24.0 C, and pH ranged from 7.5 to 8.6. Fish were fed Floating Finfish Silver feed (Zeigler Bros., Inc. Gardners, PA) ad libitum twice a week. Experimental Design Sixty-day release pellets were inserted intraperitoneally with a steel trochar into each of 25 fish per treatment on the eleventh of January, 2002. Treatments included 3 doses of p,p-DDE (2.5 mg, 5.0 mg, and 10.0 mg), 3 doses of dieldrin (0.25 mg, 0.5 mg, and 1.0mg), placebo (matrix only) and sham (no pellet inserted), total of 200 fish. Each fish was also implanted with an intraperitoneal pit tag (Trovan Corp., Bel Air, MD) to identify individuals for repeat measurements. Fish from all treatments were randomly divided into two concrete runs holding 100 fish each. Each run contained fish from all treatment groups. Sex ratios within a tank were unknown because the sex of a bass is undistinguishable externally during this time of year. Chemicals and Dosing The organochlorine pesticides 2,2-bis(4-chlorophenyl)-1,1-dichloroethylene (p,p-DDE, Lot # 09020KU, 99.4% purity) and 1,2,3,4,10,10-hexachloro-6,7-epoxy-1,4,4a,5,6,7,8,8a-octahydro-1,4,5,8-dimethanonaphthalene (dieldrin, Lot # 077H3578,

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14 91.2% purity) were obtained from Aldrich Chemical Company (Milwaukee, WI). The pesticides were shipped to Innovative Research of America, Inc. (Sarasota, FL) where they were incorporated into 60-day slow-release matrix pellets at doses of 2.5 mg, 5.0 mg and 10.0 mg p,p-DDE and 0.25 mg, 0.5 mg and 1.0 mg dieldrin, per pellet. Cholesterol, lactose, celluloses, phosphates and, stearates where used as carriers and chemical binders in the pellets. Concentrations of chemical in the pellets were calculated, based on average weight of the fish used in the study and lipid content of the gonad, to cover the range of internal doses found in the Ocklawaha system. Blood and Tissue Collection Weight, to the nearest gram (g), and length in millimeters (mm), were measured and blood was collected from each fish at Days 0 (1/11/02), 30 (2/10/02), and 60 (3/12/02). Condition factor was calculated for each time point (K = weight/length 3 x 100). At each collection, approximately 1 ml of blood was obtained from the caudal vein using a heparinized 20-gauge needle, dispensed into a 3 ml heparinized vacutainer, labeled and stored on ice until centrifuged. Blood samples were centrifuged at 1,000 x g, 4C for 20 minutes to separate red blood cells from plasma. Plasma was removed with a transfer pipette, placed in a cryovial and stored at C. At the end of exposure (Day 60) all fish were sacrificed with a blow to the head. Liver and gonads were removed and weighed to the nearest 0.01g for determination of GSI (gonad weight/body weight x 100) and HSI (liver weight/body weight x 100). Liver, gonad, muscle, and blood samples were collected for contaminant analysis from 2 fish per treatment, and gonads only from a subset of five females per treatment. Female gonads were also processed for histological analysis of reproductive stage. A cross-section of one ray of each collected gonad was placed in a histological cassette and fixed

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15 in 10% buffered formalin. Gonad, liver, and muscle samples were wrapped in tin foil and blood remained in the glass vacutainer; all were frozen at C until OCP analysis by gas chromatography-mass spectrometry (GC-MS). Gonad Histology Tissue staining with hematoxylin and eosin (H&E), sectioning, and slide mounting was performed by Histology Tech Services (Gainesville, FL). Ovaries were observed under a light microscope at 40X and stage of sexual maturation was assigned (Seplveda, 2000). Briefly, stage 1 ovaries are undeveloped with mostly primary phase follicles. Stage 2 ovaries are previtellogenic with primary and secondary phase follicles, but have no vitellogenic follicles. Stage 3 ovaries are early vitellogenic with some vitelline granules in follicles of varying size and no fully developed eggs. Stage 4 ovaries are late vitellogenic with a majority of follicles containing numerous vitelline granules and fully developed eggs are present. Analysis of Circulating Sex Steroid Hormones Blood plasma from largemouth bass was analyzed for 11-KT and E 2 using a previously validated 3 H radioimmunoassay (RIA) method (Gross et al., 2000). All samples were assayed in duplicate and values were reported as pg/ml of plasma. Standard curves were prepared in phosphate buffered saline plus gelatin and sodium azide (PBSGA) with known amounts (15, 30, 60, 125, 250, 500, 1000, and 2000 pg) of radioinert E 2 (ICN Biomedicals, Costa Mesa, CA) or 11-KT (Sigma Chemicals, St. Louis, MO) and 15,000 cpm of 3 H-E 2 or 3 H-11-KT. Each plasma sample (50 l) was extracted twice with diethyl ether prior to RIA analysis. Diethyl ether (4 ml) was added to each sample, tubes were vortexed and the ether was evaporated off, leaving behind the solid phase precipitate. The procedure was repeated in the same test tube to insure

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16 evaporation of all liquid plasma. PBSGA buffer and antibodies against sex steroid hormones were then added to the sample tube and incubated overnight at 4C. Antibodies were purchased from ICN Biomedicals (E 2 ) or Helix Biotech, Richmond, BC, Canada (11-KT). Following the antibody incubation, unbound antibody was removed by addition of dextran-coated charcoal and centrifugated for 10 min at 1,000 x g. Excess free antibody was stripped out of solution by binding to the charcoal and pelleting at the bottom of the test tube. Four hundred l of sample supernatant was removed and added to a scintillation vial with 4 ml of Scintiverse scintillation cocktail (Fisher Scientific, Pittsburgh, PA). Samples vials were then placed in a liquid scintillation counter (Pachard Tricarb, Model 1600) and counted for two minutes each. The minimum concentration distinguishable from zero (mean SE) was 94 14 pg/ml for E 2 and 29 13 pg/ml for 11-KT. Cross reactivities of the E 2 antiserum (produced and characterized by T.S. Gross, University of Florida) with other steroids were: 11.2% for estrone, 1.7% for estriol, and < 1.0% for 17-estradiol and androstenedione. Cross reactivity of the 11-KT antiserum with other steroids was: 9.65% for testosterone, 3.7% for -dihydrotestosterone, and < 1.0% for androstenedione. Analysis of Largemouth Bass Tissues for OCPs Chemical analysis was carried out at the Center for Environmental and Human Toxicology, University of Florida. Briefly, largemouth bass tissues (whole liver, half of the gonad, and left and right side muscle fillets) were homogenized to eliminate any concentration variability within the tissue due to differences in lipid content. A portion of the sample (2-5 g) was extracted into ethyl acetate. The sample was purified using C18 and NH2 SPE (solid phase extraction) cartridges. Total OCP content was determined by GC-MS, according to EPA method 8270 (EPA, 1983). Readings were not

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17 lipid normalized. Samples were analyzed multiple times in full scan mode for analyte identification and in selected ion mode (SIM) for quantitation to improve sensitivity. Percent recovery ranged between 75 and 100% with a limit of detection of 0.75-1.5 ng/g. Analysis of Largemouth Bass Blood Plasma for OCPs Blood plasma concentrations of p,p-DDE and dieldrin were determined by a competitive enzyme linked immunosorbent assay (ELISA) (Abraxis, LLC. Warmister, PA; DDE kit Lot#3C06437, Cyclodiene kit Lot #3C0643). In a competitive ELISA, a known amount of substrate (in this case p,p-DDE or dieldrin) is immobilized to the bottom of a 96-well plate. The sample to be tested is added, along with the specific antibody for the substrate. The substrate in the sample competes with the substrate bound to the plate for conjugation to a known amount of primary antibody. All sample and excess primary antibody are washed from the plate leaving the immobilized substrate with a percentage bound to the primary antibody. A secondary antibody is then added that is conjugated to a horseradish-peroxidase enzyme. This antibody only binds to the primary antibody and not to substrate that is not bound. Excess secondary antibody is then washed from the plate and a substrate for the enzyme linked to the secondary antibody, hydrogen peroxide, is added. When the hydrogen peroxide solution is cleaved by the horseradish-peroxidase, a blue color results. This reaction is stopped after a certain period of developing time by addition of a dilute acid solution. The plate is then read at 450 nm for quantification of the amount of color in each well, which is inversely proportional to the amount of substrate in the sample. First, all kit components and samples were first brought to room temperature. Each sample was diluted 1:50 in 10% methanol for DDE and 1:100 for dieldrin in 25% methanol. Concentrated wash buffer (5X) provided in the kit was diluted 1:5 with

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18 distilled water. Standards for DDE were made by adding p,p-DDE stock solution (10,000 ppb) to a 1:50 dilution of plasma from untreated bass with 10% methanol as per instructions. Standards for dieldrin were included in the kit and required only 1:100 dilution with 25% methanol in a glass test tube. Parafilm was placed over each test tube and vortexed to mix. On a 96 well plate, 25 l of each standard and sample was added to separate wells. Standards and samples were run in duplicate. Next, 100 l of rabbit anti-DDE/anit-cyclodiene antibody was added to each well. Wells were covered with tape and contents of the plate were mixed by moving the plate in a circular motion on the bench top for approximately 30 s. The plate was then incubated at room temperature for 60 min. Following incubation, the covering was removed and the contents dumped by vigorous shaking. Wells were then washed three times with 250 l 1X wash solution, followed by blotting off excess wash with a paper towel. Next, 100 l of anti-rabbit-horseradish peroxidase secondary antibody was added to each well, covered, and allowed to incubate at room temperature for 30 min. Following incubation, the covering was removed, contents dumped, and wells washed 3 times in 1X was solution as before. One hundred l of hydrogen peroxide color solution was then added to each well and incubated at room temperature for 20 min. Fifty l of acidic stopping solution was then added to each well. The plate was read at 450 nm on a MRX Microplate Reader (Dynex Technologies) within 15 min of application of the stopping solution. Results are given as ng/ml. The limit of detection for p,p-DDE was 62.5 ppb. Cross reactivity of the p,p-DDE antibody were 46% p,p-DDD, 16% o,p-DDD, 10% p,p-DDT, and 3.2% o,p-DDE. The limit of detection for dieldrin was 25 ppb. Cross reactivity of the cyclodiene

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19 antibody were 150% endosulfan, 58% heptachlor, 26% aldrin, 26% chlordane, and 8.2% toxaphene. High cross reactivity to other chemicals did not influence the analysis for this experiment as all fish were from an uncontaminated environment and had low to no background exposure to other chemicals. Statistical Analysis All statistical analyses were performed using Statistical Analysis System (SAS) software, version 9. The original experimental design called for statistical analysis comparing each treatment group by ANOVA and a multiple comparison procedure. The Univariate procedure was run on all data sets to determine if the data was normally distributed. ANOVAs were then performed and significance was set at = 0.05. Duncans Multiple Range test followed as a multiple comparison procedure to determine which treatment groups differed. Results are presented as means SD. As detailed in the results section, dosing was inconsistent within treatment groups. Therefore, further statistical analysis was done by linear regression of each parameter against the blood plasma chemical concentration of each individual fish (SAS, = 0.05). The independent variable was blood plasma concentration of either chemical and the dependent variables were weight, length, K, HSI, GSI, blood plasma E 2 and 11-KT concentrations. Percent change in hormone concentrations from Days 0 to 30 and from Days 30 to 60 were also regressed against Day 30 and Day 60 blood plasma p,p-DDE and dieldrin concentrations, respectively. Individual fish that exhibited increased concentrations of both p,p-DDE and dieldrin were considered outliers and removed from the data set because possible interactions of these pesticides remain unknown.

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20 Results Largemouth Bass Survivorship over the course of exposure was 95.5% excluding one incident that occurred on January 22 where water flow was lost in both tanks for approximately 10 hours overnight and 15 fish died due to dissolved oxygen levels falling below 3 mg/L. This incident brought survivorship down to 88%. No long term adverse effects on surviving fish were observed. Weight, length and K steadily increased over the course of 60 days for males and females (Figures 2-1 and 2-2) and did not vary between treatments. Health Parameters: K and HSI Condition factor did not vary among treatments in females or males over the course of the study. Female Day 60 HSI did not vary between treatment groups or from controls; however, males in the dieldrin 1.0 mg treatment group had a significantly higher HSI compared to all other treatments (Figure 2-3). Gonad Histology Histological analysis of female gonads taken at Day 60 indicated that 89.2% of females were in late vitellogenesis (stage 4) with the remaining in early vitellogenesis (stage 3) (Figure 2-4). Reproductive Parameters: GSI and Circulating Sex Steroids GSI GSI, determined on Day 60, did not differ between treatment groups in neither females nor males (Figure 2-5). Average GSI was 4.0 1.7% (n = 65) and 0.66 0.16% (n = 90) for females and males, respectively.

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21 17-Estradiol In female bass, the placebo and DDE 2.5 mg treatment groups consistently showed higher, though not always significant, circulating E 2 concentrations. At Day 0 the females in the placebo group had a significantly higher E 2 concentration compared to DDE 5.0, DDE 10.0, and all dieldrin treatment groups (Figure 2-6a). The E 2 concentrations in these treatments, however, did not differ from the Sham treatment. At Day 30, females in the placebo and DDE 2.5 treatment groups had a significantly higher E 2 concentration as compared to all other treatments (Figure 2-6b). By Day 60, variability within treatment groups had increased and only females from the DDE 2.5 group demonstrated a significantly higher E 2 concentration than DDE 10.0 and dieldrin 0.5 females (Figure 2-6c). In male bass, the DDE 5.0 mg treatment group consistently demonstrated higher circulating E 2 concentrations as compared to all other treatments. At Days 0 and 30 all treatments except DDE 5.0 exhibited no significant difference from each other (Figure 2-7a,b). At Day 60 variability within treatment groups had increased. Males in the DDE 5.0 treatment group had a significantly higher circulating E 2 concentration as compared to males in the sham and dieldrin 0.5 treatments (Figure 2-7c). 11-Ketotestosterone In female bass, most treatment groups demonstrated much higher variability in circulating 11-KT concentrations as compared to controls. At Day 0, females in the DDE 5.0, dieldrin 0.5, and dieldrin 1.0 had a significantly higher 11-KT concentration than sham, placebo, DDE 2.5, and DDE 10.0 treatments (Figure 2-8a). By Day 30, only females in the DDE 5.0 treatment group had significantly higher 11-KT concentration

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22 (Figure 2-8b). At Day 60 all treatment groups had similar 11-KT concentrations (Figure 2-8c). In male bass, placebo and DDE 2.5 treatment groups demonstrated a consistently, though not always significant, higher circulating 11-KT concentration as compared to sham and other treatment groups. At Day 0, males in the DDE 2.5 treatment group had significantly higher 11-KT concentrations than males in the sham group (Figure 2-9a). At Day 30, 11-KT concentrations were significantly higher in the DDE 2.5 treatment group as compared to all treatments except placebo (Figure 2-9b). Day 60 11-KT concentrations for males in the DDE 2.5 treatment group remained significantly higher than sham, DDE 10.0 and all dieldrin treatments (Figure 2-9c). In Vivo Treatment Dosing Consistency Contaminant analyses of gonad (n = 7 per treatment), liver, muscle, and blood (n = 2 per treatment), revealed that either the pellets within each treatment group differed or the pellets did not release a consistent dose. Some fish had exposure to both p,p-DDE and dieldrin indicating some pellets may have been cross-contaminated (Figure 2-10; Table 2-1). Also, examination of the peritoneal cavity from each fish at the end of the study demonstrated that pellets had varying degrees of dissolution (i.e. some pellets remained whole while others completely dissolved). Consequently, gonadal concentrations reached in each treatment group were highly variable. Despite the high variability in gonadal dose, mean p,p-DDE gonadal concentrations demonstrated an apparent dose-response (Figure 2-10a). The range of gonadal p,p-DDE concentrations achieved in this study included the range of gonadal p,p-DDE levels found in 1996 of 3 year old bass in the Emeralda Marsh Conservation Area (EMCA), 1300 to 4200 ppb (total DDT derivatives) (Marburger, 2002). On the other hand, gonadal dieldrin doses

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23 were not specific for their target dose (i.e. the low dieldrin treatment group received a mean dose higher than that of the high dieldrin treatment group) and the range achieved greatly exceed that which is found in the wild for gonadal dieldrin, 70 to 130 ppb (Marburger, 2002; Figure 2-10b). Because tissue (liver, gonad, muscle) was analyzed from only two fish per treatment, it was impossible to determine individual exposures for every animal. However, a blood plasma sample was obtained from every fish at three time points. Therefore, a linear regression analysis (SAS) was performed to determine how well blood concentrations correlated with specific tissue concentrations for each chemical. In females and males, p,p-DDE blood concentrations were significantly correlated to liver (female liver = -8 + 1.5(blood), r 2 = 0.8345; male liver = -46 + 1.4(blood), r 2 = 0.8807) and gonad (female gonad = 120 + 10(blood), r 2 = 0.9520; male gonad = 0.4 + 10.5(blood), r 2 = 0.7520). However, muscle only had a significant relationship with blood p,p-DDE concentration in females (female muscle = 36 + 0.25(blood), r 2 = 0.5858; male muscle = 9 + 0.5(blood), r 2 = 0.4041) (Figure 2-11). For dieldrin, female and male blood concentrations were significantly correlated to liver (female liver = -5 + 0.6(blood), r 2 = 0.9986; male liver = 52.5 + 0.3(blood), r 2 = 0.9650) and gonad (female gonad = -51 + 5.2(blood), r 2 = 0.9262; male gonad = 412 + 1.5(blood), r 2 = 0.6824), but muscle dieldrin was only correlated with blood dieldrin in males (female muscle = -9 + 0.3(blood), r 2 = 0.8527; male muscle = 22.7 + 0.14(blood), r 2 = 0.9120) concentrations (Figure 2-12). Because a significant linear relationship between blood and tissue concentration existed, a method for analyzing blood plasma concentrations was established. However, caution should be taken in extrapolating blood plasma concentrations to gonadal concentrations,

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24 especially dieldrin, as some linear regression models showed a weak goodness of fit, explaining less than half of the variability of the data set. New ELISA detection kits for p,p-DDE and dieldrin were developed by Abraxis, LLC (Warminster, PA) for detection in largemouth bass blood plasma. This method was validated in our lab by analyzing a pool of blood plasma samples from treated bass in parallel by GC-MS and the new ELISA method. A linear regression was used to determine the correlation of results obtained by the two methods for both pesticides (Figure 2-13). It can be seen in the graph of this data that the ability of the ELISA to detect concentrations below 100 ppb is limited. All control fish, most treated fish at Day 30, and some treated fish at Day 60 register OCP concentrations below 100 ppb indicating they are at, near, or below the limit of detection of the ELISA kits. However, correlation between GC-MS and ELISA methods was highly significant and linear for both p,p-DDE and dieldrin. Therefore, ELISA was used to analyze blood samples from every treated fish, and 2 males and 2 females from control groups (mean concentrations of these fish were substituted in as the exposure level for all other control fish), at Days 30 and 60 in order to assess the achieved dose over time. Health Parameters: K and HSI (Regression) At Day 0, all fish were healthy with an average condition factor of 1.46 0.11. Condition factor increased steadily over the course of the exposure period, as body weight increased mostly due to the developing gonads (Figure 2-1c and Figure 2-2c). HSI was determined at Day 60 and did not show a significant correlation with blood plasma concentration of either p,p-DDE (Figure 2-14) or dieldrin (Figure 2-15). Average HSI SD for all fish at Day 60 was 3.02 0.16% (n = 156).

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25 Reproductive Parameters: GSI and Circulating Sex Steroids (Regression) GSI GSI was determined on Day 60 and neither males nor females showed a significant relationship with blood plasma concentration of either p,p-DDE (Figure 2-16) or dieldrin (Figure 2-17). Average GSI SD was 4.0 1.7% (n = 65) and 0.66 0.16% (n = 90) for females and males, respectively. 17-Estradiol On Day 0 (January 11) female and male bass had mean SD E 2 concentrations of 527 278.6 and 264 106.4 pg/ml, respectively (Figure 2-18). E 2 /11-KT ratio at Day 0 was 2.5 1.9 for females and 0.5 .2 for males. In fish treated with p,p-DDE, both Day 30 circulating E 2 and percent change in E 2 from Days 0 to 30 showed a significant negative relationship with increasing blood plasma concentration (r 2 = 0.2815 and 0.1401, respectively) in females (Figure 2-19) while no relationship between Day 30 E 2 and blood plasma p,p-DDE existed in males (Figure 2-20). Circulating E 2 also showed a significant negative relationship with blood plasma p,p-DDE (r 2 = 0.1588) at Day 60 in females (Figure 2-21a). However, the GSI and percent change in E 2 from Days 30 to 60 showed no relationship with blood plasma concentration (Figure 2-21b,c). Again, males exhibited no relationship between blood plasma p,p-DDE and E 2 (Figure 2-22). In fish treated with dieldrin neither females nor males showed a significant relationship with blood plasma concentration at Day 30 (Figures 2-23 and 2-24). At Day 60, female circulating E 2 (r 2 = 0.0933), but not percent change in E 2 from Days 30 to 60, showed a significant negative relationship with blood plasma dieldrin (Figure 2-25a,b). This decrease in E 2 was not correlated with GSI (Figure 2-25c). Males showed no

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26 relationship between E 2 and blood plasma dieldrin concentrations at Day 60 (Figure 2-26). 11-Ketotestosterone Day 0 circulating 11-KT concentrations were 301 211.6 and 731 345.9 pg/ml, respectively for females and males (Figure 2-18). In fish treated with p,p-DDE, Day 30 circulating 11-KT concentrations showed no relationship with blood plasma concentration in females (Figure 2-19) or males (Figure 2-20). Also, at Day 60 neither females nor males showed a relationship between 11-KT and blood plasma p,p-DDE or GSI (Figures 2-21 and 2-22). At Day 30, neither sex of fish treated with dieldrin showed a significant relationship between blood plasma concentration and circulating 11-KT (Figures 2-23a and 2-24a). However, males did show a trend of decreasing 11-KT and a significant negative relationship between blood plasma dieldrin and percent change in 11-KT from Days 0 to 30 (r 2 = 0.1627) (Figure 2-24). At Day 60, females again demonstrated no relationship between blood plasma dieldrin concentration and 11-KT (Figure 2-25). Male circulating 11-KT had a significant negative relationship with blood plasma dieldrin at Day 60 (r 2 = 0.1610) and no relationship with GSI. Also, no relationship existed between blood plasma dieldrin concentration and percentage change in 11-KT from Days 30-60 (Figure 2-26). Discussion The data collected from this study indicates that time-release pellets are not an ideal dosing route for largemouth bass and represents a common problem in ecotoxicology research: applied vs. attained dose. Many studies simply apply a dose and do not perform the necessary analysis to determine 1) how much of the applied dose was attained, 2) if

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27 the applied dose reached the target tissue, and 3) the variability of the attained dose within treatment groups or replicates. This study demonstrated the extreme variability inherent to the pellet dosing method (Figure 2-10). Tissue samples were taken at two timepoints during the study to examine tissue difference in bioaccumulation/chemical partitioning, but also served as an indicator of variability of dose within treatment groups. Originally, the study design called for statistical analysis (ANOVA) to be run by treatment group. ANOVA and regression analyses were in agreement and showed no biologically significant differences between treatment groups for weight, length, K, HSI or GSI. However, the two analysis methods resulted in different conclusions of dose effects on circulating sex steroids. ANOVA analysis indicated high doses of p,p-DDE and all doses of dieldrin significantly reduced E 2 levels in females at Day 30, as compared to placebo, but not to sham, however, this relationship is similar to the differences in hormone levels between treatments at Day 0 (Figure 2-6). Regression analysis demonstrated a relationship of depressed E 2 concentrations in females at Day 30 only with increasing p,p-DDE dose (Figure 2-19). Day 60 data analyzed by ANOVA only shows a reduction in E 2 at the high p,p-DDE and medium dieldrin doses as compared to the low p,p-DDE dose and not controls (Figure 2-6c). On the other hand, regression analysis showed a relationship with decreased E 2 concentrations only at the highest doses of both pesticides (Figure 2-21). ANOVA analysis of male sex steroid levels also was hindered by the high variability of sex steroid concentrations within a treatment group. No clear relationships between dose and 11-KT concentration could be drawn. Means were clearly lower in the dieldrin treatments, but were not significantly different from controls (Figure 2-9).

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28 However, regression analysis demonstrates a significant negative relationship between dieldrin dose and 11-KT concentrations. Also apparent from the ANOVA analysis was that fish in some treatment groups had consistently higher/lower sex steroid concentrations as compared to all other treatment groups. This may have been due to differences in relative sexual maturity or size of the fish, however, there were no differences among treatments in size (ANOVA and regression), regression of GSI against sex steroid concentrations showed no relationship (Figure 2-21, 2-22, 2-24, 2-25), and histologic analysis of the gonad at Day 60 showed that almost 90% of females were at the same reproductive stage (Figure 2-4). The cause of this variability is unknown. Analysis based only on the applied dose (implant concentration) could lead to erroneous conclusions. The problem lies in that the high doses were not consistently higher than the low or medium doses, especially in the case of dieldrin where gonadal dose indicated the high treatment group received a lower dose than the low treatment group. If a relationship between OCP dose and circulating E 2 in females and 11-KT in males truly does exist, it would be impossible to determine from a by treatment statistical analysis due to the variability in dose within a treatment group. Not to be ignored, however, is the variability in hormone data within the control groups, especially as hormone concentrations are at peak levels. Like many biological endpoints, circulating hormone concentrations show a wide inter-individual variability making statistical modeling difficult. The data sets of hormone concentrations generated from this study do not appear to fit well to a linear model. r 2 values of statistically significant relationships ranged from 0.15 to 0.28 indicating the biological significance of these relationships is weak. Further statistical

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29 analysis with a non-linear model is warranted. However, the results of this study suggest that p,p-DDE and dieldrin may be able to depress sex steroid synthesis in largemouth bass exposed to high doses. In general, this study provides evidence that p,p-DDE and dieldrin are endocrine modulators in the Florida largemouth bass at concentrations above what was found in three year old female bass gonadal tissues collected from contaminated areas of the Ocklawaha River Basin in 1996 (Marburger et al.1999; 2002). Another study of OCP concentrations in female largemouth bass ovaries was performed in 2002 and showed mean ovarian concentration of total OCPs to be approximately 40,000 ppb, 10-fold higher than previously reported, indicating concentrations achieved in the present study were considerably low (Seplveda et al., 2003). However, the two studies show approximately the same level of dieldrin in bass ovaries. This conflicting data can most likely be resolved by evaluating the time periods the studies were conducted in. Marburger et al. (1999, 2002) collected samples in 1996 of fish that had been stocked in 1993, while Seplveda et al. (2003) collected samples in 2002 from fish with a mean age of six years. The difference in exposure time would allow for the bioaccumulation of larger concentrations of OCPs over time. In this study, neither chemical induced a change in any health parameter (weight, length, K, and HSI), indicating that acutely toxic doses were not achieved. Of note is that circulating levels of E 2 and 11-KT were approximately 2-fold less than reported for the same age, pond-reared fish in the Gross et al. (2000) study and was most likely due to stress induced by captive holding conditions (i.e. unnatural housing and high stocking density). Fish were stocked at a density of 1 bass per 61 liters of water. At this density,

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30 largemouth bass remain healthy but will not spawn. For reference, spawning enclosures built on the edge of a pond at the USGS-CARS facility are stocked at an approximate density of 1 fish per 2500 liters. The ratio of E 2 to 11-KT, however, was within normal ranges, greater than 1 for females and less than 1 for males, indicating that these fish were sexually mature (Marburger et al., 1999). The average size of the fish used in this study indicates they were approximately 1-2 years of age and most likely going through their first reproductive cycle. This may have increased the variability in sex steroid concentrations. Overall, increasing p,p-DDE concentration over time caused a decrease in circulating E 2 levels in females, with no other health or reproductive effects and was not correlated with relative gonad size. p,p-DDE concentrations reached by Day 30 of exposure were enough to stifle the seasonal increase in E 2 the magnitude of which increased with increasing dose as shown by the relationship of p,p-DDE dose to the percent change in E 2 concentration from Days 0 to 30. Still demonstrating a dose-dependant depression at Day 60, E 2 levels in females was not correlated with GSI. This is most likely because the gonad begins to mature in October and reaches a considerable size (approximately 5% body weight) by January when this study was initiated (Gross et al., 2000). Peak GSI of about 6% occurs is in February. A relevant dose of p,p-DDE was most likely not achieved by the time peak GSI was reached and therefore could have little to no effect on this parameter. A study of male guppies (Poecilia reticulata) demonstrated a decrease in GSI when fed a total of 15 and 150 g of p,p-DDE over 30 days; the guppies fed 1.5 g showed no change in GSI (Baatrup and Junge, 2001), however, no contaminant analysis was done in this study so attained dose in unknown.

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31 p,p-DDE also had no effect on circulating 11-KT levels in males or females in this study despite its anti-androgenic properties in mammals (Kelce et al., 1995) and its ability to bind the AR of some fish species (Wells and Van Der Kraak, 2000). It is possible that p,p-DDE may not bind the AR in largemouth bass. If it was a competitive inhibitor, testosterone levels would be expected to decrease providing less substrate for 11-KT synthesis and therefore a dose dependent decrease in circulating 11-KT. In comparison, low p,p-DDE doses (liver concentration 400 ppb, 2 to 4 times less than what was achieved in this study) in male summer flounder (Paralichthys dentatus) injected IP showed no change in GSI or hormone concentrations (Mills et al., 2001). Conversely, white sturgeon (Acipenser transmontanus) from the Columbia River with liver p,p-DDE concentrations over 700 ppb had severely depressed testosterone and 11-KT concentrations (Foster et al., 2001). Similarily, largemouth bass collected from the Emeralda Marsh and the St. Johns River had depressed sex steroid concentrations (E 2 in females and 11-KT in males), as well as, decreased GSI (Marburger et al., 1999; Seplveda et al., 2002). These fish however, were exposed to multiple pesticides and other chemicals and any mixture effects are unknown. Difference between species most likely contributes to apparently varying effects of p,p-DDE. Dieldrin had no statistically significant effects on circulating sex steroid levels in either sex at Day 30 of exposure. However, the percent increase in 11-KT concentration from Day 0 to Day 30 was significantly less in dieldrin-treated versus untreated males, and although not statistically significant, 11-KT levels tended to decrease with increasing doses of dieldrin. By Day 60, females and males treated with dieldrin had lower circulating E 2 and 11-KT, respectively, than controls. GSI was unaffected, probably due

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32 to the timing of the exposure, and was not correlated with 11-KT concentrations. No studies of in vivo dieldrin-only exposure in fish were found to compare the results to. However the studies mentioned above from the Emeralda Marsh and the St. Johns River also demonstrated elevated levels of dieldrin in largemouth bass tissue and depression in sex steroid hormones (Marburger et al., 1999; Seplveda et al., 2002). This may be due, in part, to the effects of dieldrin. The mechanisms of action for p,p-DDE and dieldrin as endocrine modulators in fish are unknown and many possibilities exist. First, either of these chemicals could bind to the plasma sex steroid binding protein (SSBP) and disturb endogenous hormone binding making it available in the blood for biotransformation and elimination. This is a more likely mechanism for dieldrin because its effects include decreases in both hormones assayed. However, two studies on rainbow trout blood plasma have shown little to no ability for dieldrin or DDT derivatives to bind the SSBP effectively (Milligan et al., 1998; Tollefson, 2002b). Second, either pesticide may be able to inhibit aromatase activity. Aromatase in fish converts testosterone to E 2 and 11-KT. Inhibition of this enzyme would lead to a decrease in E 2 in females and 11-KT in males, as is the case for dieldrin. Evidence of this mechanism has been demonstrated in alligators in OCP contaminated areas of Lake Apopka (Crain et al., 1997). However, p,p-DDE exposure in a male rat study demonstrated increased aromatase activity (You et al., 2001). Third, these OCPs may be able to induce enzymes responsible for biotransformation of sex steroids thereby eliminating them at a faster rate than they are produced. This mechanism stands alone in that it is not directly involved in the endocrine

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33 system and demonstrates a classic toxic action of inducing catabolism. The catabolic pathway for sex steroids has not been fully characterized in fish but is believed to include the cytochrome P450 isoform 3A (Stegman and Hahn, 1994). Studies on dieldrin in human hepatoma cells (Coumoul et al., 2002) and rats (You et al., 1999) have shown that increasing OCP load can induce CYP3A. However, species differences make extrapolation of this data to fish uncertain. A marker of liver function and CYP induction in general, EROD, has shown increased activity with increasing OCP load and related decreases in circulating sex steroids in several studies (Dickerson et al., 1999; Zapata-Perez et al., 2000; Foster et al., 2001; Wade et al., 2002), however, a relation of EROD to sex steroid metabolism has not been established. Lastly, p,p-DDE and dieldrin could disrupt normal endocrine function in fish by interfering with the feedback pathways of the hypothalamus-pituitary-gonad axis. p,p-DDE and dieldrin have both demonstrated the ability to bind estrogen and androgen receptors (Kelce et al., 1995; Danzo, 1997; Nimrod and Benson, 1997; Matthews et al., 2000; Wells and Van Der Kraak, 2000; Anderson et al., 2002) and could therefore bind to these receptors anywhere in the feedback pathway. Binding of high concentrations at the hypothalamus and pituitary would inhibit further production of GnRH and GTH, respectively, effectively shutting down the signal to the gonad for sex steroid synthesis with some temporality between exposure and effect. This is the mechanism by which the normal annual cycling of hormones occurs (Janz and Weber, 2000). Low sex steroid levels serve as a stimulus for further sex steroid production (positive feedback) and is sex specific for which hormone is upregulated. However, when plasma hormone concentrations reach a threshold concentration, the signal for steroid synthesis is shut off

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34 (negative feedback) and hormone levels slowly return to baseline. If these pesticides are able to mimic E 2 and/or 11-KT action, then it is possible they could bind to receptors in the brain and induce the hypothalamus or pituitary to shut down sex steroid synthesis. A human hypothalamic cell line study examined the effects of OCPs (methoxychlor and chlorpyrifos) on this pathway and found low levels of exposure stimulates GnRH release, theoretically by stimulating sex steroid receptors initiating positive feedback (Gore, 2002). However, Van Der Kraak et al. (1992) demonstrated decreased GTH in white sucker fish (Catostomus commersoni) exposed to bleached kraft pulp mill effluent (BKME). Although BKME does not contain OCPs it does contain some chlorinated chemicals that are thought to be endocrine active. It is clear that more research is needed to elucidate the mechanism by which p,p-DDE and dieldrin exposure leads to decreased plasma sex steroid concentrations. This study did not include any endpoints that would indicate the functionality of p,p-DDE or dieldrin as enzyme inducers/inhibitors or estrogen/androgen mimics. Aromatase, CYP activity, circulating vitellogenin, GnRH, and GTH could serve as indicators of the biochemical actions of these pesticides. In order to draw accurate conclusions about how OCPs cause a depression in sex steroid levels, it is imperative to understand by which mechanisms they are acting. However, this study did not aim to elucidate the mechanism of action for these chemicals. It serves only as evidence that p,p-DDE and dieldrin exposure at environmentally relevant levels cause a decrease in circulating sex steroids in largemouth bass. This study also addressed the accuracy of dosing fish by time-release pellets and characterized relevant endpoints of exposure and their dose response to p,p-DDE and dieldrin. The results show that two of the

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35 predominant chemicals found in the contaminated areas of the Ocklawaha River Basin do contribute to the depressed sex steroid levels and altered reproductive success of largemouth bass exposed in the wild. The data presented here will be used to design future studies aimed at determining the effects of these pesticides in binary, ternary, and complex mixtures, similar to what is found in the environment.

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36 Day 0Day 30Day 60 0 100 200 300Weight (g)a Day 0Day 30Day 60 0 100 200 300Length (mm)b Day 0Day 30Day 60 0 1 2 3Condition Factorc Figure 2-1. Female body condition over time. Box plots of weight (a), length (b), and condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 63). Box plot contains the 25 th to the 75 th quartile, line in box indicates the median, whiskers extend to the minimum and maximum value.

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37 Day 0Day 30Day 60 0 100 200 300Weight (g)a Day 0Day 30Day 60 0 100 200 300Length (mm)b Day 0Day 30Day 60 0 1 2 3cCondition Factor Figure 2-2. Male body condition over time. Box plots of weight (a), length (b), and condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 93). Box plot contains the 25 th to the 75 th quartile, line in box indicates the median, whiskers extend to the minimum and maximum value.

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38 ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 1 2 3 4 5aaaaaaaaaFemale HSI (%) ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 1 2 3 4 5abaa,ba,ba,ba,ba,bbMale HSI (%) Figure 2-3. Day 60 HSI. There was no significant difference in HSI among treatment group for females (a). Males in the Dieldrin 1.0 mg treatment group showed a significant increase in HSI over Sham and DDE 2.5 mg treatments (b). Sample size per treatment ranged from 7 to 11 for females and 7 to 16 for males. Treatments with the same lower case letter were not significantly different.

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39 CA GV GV YV Figure 2-4. Histological section of stage 4 vitellogenic female gonad viewed at 40X. CA = cortical alveoli, GV = germinal vesicle, YV = yolk vesicle.

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40 ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0.0 2.5 5.0 7.5 10.0aaaaaaaaaFemale GSI (%) ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0.0 0.5 1.0 1.5aaaaaaaabMale GSI (%) Figure 2-5. Day 60 GSI. There was no significant difference in GSI among treatment group for females (a) or males (b). Sample size per treatment ranged from 7 to 11 for females and 7 to 16 for males. Treatments with the same lower case letter were not significantly different.

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41 ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 500 1000 1500aabbbbbEstradiol (pg/ml)a,ba,b ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 500 1000 1500bEstradiol (pg/ml)aabbbbbb ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 500 1000 1500cbaa,ba,ba,bba,ba,bEstradiol (pg/ml) Figure 2-6. Female circulating estradiol at Days 0 (a), 30 (b), and 60 (c). Placebo and DDE 2.5 treatments showed consistently higher plasma E 2 concentrations than all other treatments. Sample size per treatment ranged from 7 to 11. Treatments with the same lower case letter were not significantly different.

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42 ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 250 500 750 1000aEstradiol (pg/ml)aaaabaaa,b ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 250 500 750 1000bEstradiol (pg/ml)aaabaaaa ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 250 500 750 1000Estradiol (pg/ml)caaa,ba,ba,ba,ba,bb Figure 2-7. Male circulating estradiol at Days 0 (a), 30 (b), and 60 (c). DDE 5.0 mg treatment showed a consistently higher mean than all other treatments. Sample size per treatment ranged from 7 to 16. Treatments with the same lower case letter were not significantly different.

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43 ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 250 500 750 100011-Ketotestosterone(pg/ml)aa,ba,bbbaaaa ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 250 500 750 100011-Ketotestosterone(pg/ml)baaaaaa,bba,b ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 250 500 750 100011-Ketotestosterone(pg/ml)caaaaaaaa Figure 2-8. Female circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Treatment groups demonstrated high variability compared to controls. By Day 60, no difference existed between treatment groups and controls. Sample size per treatment ranged from 7 to 11. Treatments with the same lower case letter were not significantly different.

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44 ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 500 1000 150011-Ketotestosterone(pg/ml)aa,ba,ba,ba,ba,ba,bab ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 500 1000 1500baa,baaaaab11-Ketotestosterone(pg/ml) ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Dieldrin 0.25Dieldrin 0.5Dieldrin 1.0 0 500 1000 150011-Ketotestosterone(pg/ml)a,baaaaaa,bbc Figure 2-9. Male circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Placebo and DDE 2.5 mg treatment groups demonstrated consistently higher 11-KT. At Day 30 and 60 all dieldrin treatments and higher DDE treatments demonstrated decreased 11-KT. Sample size per treatment ranged from 7 to 16. Treatments with the same lower case letter were not significantly different.

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45 ShamPlaceboDDE 2.5DDE 5.0DDE 10.0Eustis 1996 0 5000 10000DDE Targetdose Treatmentp,p'-DDE (ppb) a ShamPlaceboDieldrin 0.25Dieldrin 0.5Dieldrin 1.0Eustis 1996 0 1000 2000 3000 4000Dieldrin Targetdose TreatmentDieldrin (ppb) b Figure 2-10. Day 60 mean female gonadal dose. p,p-DDE (a) and dieldrin (b) concentration SD for each treatment group (n = 6) compared to the target dose for each treatment. Also included is the mean gonadal concentration of either OCP found in the Eustis property of EMCA.

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46 Table 2-1. GC-MS results of tissue samples for bioaccumulation measurements at Day 30 of exposure. Blue and green numbers highlight fish treated with p,p-DDE and dieldrin, respectively. a DDE DDE DDE DDE Avg. Std. Dev. Treatment Sex Liver gonad blood muscle Gonad DDE Sham M 5.51 10.00 1.32 14.30 6.08 F 4.41 18.60 14.49 Placebo M 4.27 23.59 5.77 1.31 22.105 2.10 F 4.48 20.62 2.5 1.47 DDE 2.5 F 1733.85 7808.63 907.83 113.34 6384.47 2014.07 F 349.01 4960.30 481.58 204.42 DDE 5.0 F 198.41 2475.52 116.67 92.48 6644.74 737.67 F 1241.63 10813.95 1006.45 348.37 DDE 10.0 M 363.37 5108.70 316.07 318.06 2920.96 2204.06 M 642.11 2953.26 390.70 111.61 M 280.84 700.93 278.57 48.47 Dieldrin 0.25 M 95.40 365.85 101.56 19.88 F 121.60 1521.21 171.95 26.28 Dieldrin 0.5 F 7.31 36.20 7.15 M 558.37 1024.94 270.59 Dieldrin 1.0 M 6.85 182.08 6.38 1.40 M 5.87 49.56 13.24 5.23 b Dieldrin Dieldrin Dieldrin Dieldrin Avg. Gonad Std. Dev. Treatment Sex Liver gonad blood muscle Dieldrin Sham M 11.03 20.00 3.95 27.17 10.13 F 13.23 34.33 2.90 Placebo M 11.4 31.45 2.62 34.135 3.80 F 14.95 36.82 2.94 DDE 2.5 F 13.11 32.72 1.37 F 8.73 38.02 2.21 DDE 5.0 F 55.20 567.83 55.56 33.29 F 129.19 1453.49 193.55 53.60 DDE 10.0 M 11.63 50.72 5.56 M 13.44 29.75 3.35 M 8.64 24.03 1.35 Dieldrin 0.25 M 143.81 515.12 196.88 37.28 1478.92 1363.01 F 223.92 2442.71 384.15 57.81 Dieldrin 0.5 F 153.51 1245.17 230.04 1247.28 2.98 M 748.35 1249.38 519.10 Dieldrin 1.0 M 127.34 1443.64 172.34 54.72 1095.39 492.50 M 114.48 747.14 282.35 92.11

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47 0 250 500 750 1000 1250 0 5000 10000 15000r2=0.9520DDE blood (ppb)DDE gonad (ppb)a 0 100 200 300 400 500 0 2500 5000 7500r2=0.7520DDE blood (ppb)DDE gonad (ppb)d 0 250 500 750 1000 1250 0 1000 2000r2=0.8345DDE blood (ppb)DDE liver (ppb)b 0 100 200 300 400 500 0 250 500 750r2=0.8807DDE blood (ppb)DDE liver (ppb)e 0 250 500 750 1000 1250 0 100 200 300 400r2=0.5858DDE blood (ppb)DDE muscle (ppb)c 0 100 200 300 400 500 0 100 200 300 400r2=0.4041DDE blood (ppb)DDE muscle (ppb)f Figure 2-11. Regression analysis of gonad (a, d), liver (b, e) and muscle (c, f) DDE concentrations against blood DDE concentrations in females (a c; n = 9) and males (d f; n = 6). Significant linear relationships exist between blood plasma and tissue DDE concentrations, except in male muscle.

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48 0 100 200 300 400 500 600 0 1000 2000 3000Dieldrin blood (ppb)Dieldrin gonad (ppb)r2=0.9262a 0 250 500 750 1000 1250 0 1000 2000 3000Dieldrin blood (ppb)Dieldrin gonald (ppb)r2=0.6824d 0 100 200 300 400 500 600 0 100 200 300 400Dieldrin blood (ppb)Dieldrin liver (ppb)r2=0.9986b 0 250 500 750 1000 1250 0 100 200 300 400 500Dieldrin blood (ppb)Dieldrin liver (ppb)r2=0.9650e 0 100 200 300 400 500 600 0 100 200Dieldrin blood (ppb)Dieldrin muscle (ppb)r2=0.8527c 0 250 500 750 1000 1250 0 100 200Dieldrin blood (ppb)Dieldrin muscle (ppb)r2=0.9120f Figure 2-12. Regression analysis of gonad (a, d), liver (b, e), and muscle (c, f) dieldrin concentrations against blood dieldrin concentrations in females (a c; n = 4) and males (d f; n = 6). Significant linear relationships exist between blood plasma and tissue dieldrin concentrations, except in female muscle.

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49 0 250 500 750 1000 1250 1500 0 500 1000 1500r2=0.7790aDDE GC-MSDDE ELISA 0 100 200 300 400 500 600 0 250 500 750r2=0.8678bDieldrin GC-MSDieldrin ELISA Figure 2-13. Linear regression of ELISA DDE (a) or dieldrin (b) results of pooled blood samples against GC-MS results (n = 12). A highly significant and linear relationship exists between the two methods.

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50 0 250 500 750 1000 0 1 2 3 4 5 6r2=0.0001aPlasma DDE (ppb)Female HSI (%) 0 500 1000 1500 2000 0 1 2 3 4 5 6r2=0.0026bPlasma DDE (ppb)Male HSI (%) Figure 2-14. Linear regression of HSI against blood plasma DDE concentrations. No significant correlation was found for females (a, n = 37) or males (b, n = 53).

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51 0 100 200 300 400 0 1 2 3 4 5 6r2=0.0005aPlasma Dieldrin (ppb)Female HSI (%) 0 100 200 300 400 500 0 1 2 3 4 5 6r2=0.0815bPlasma Dieldrin (ppb)Male HSI (%) Figure 2-15. Linear regression of HSI against blood plasma dieldrin concentration. No significant correlation was found for females (a, n = 34) or males (b, n = 51).

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52 0 250 500 750 1000 0.0 2.5 5.0 7.5 10.0r2=0.0071aPlasma DDE (ppb)Female GSI (%) 0 500 1000 1500 2000 0.0 0.5 1.0 1.5r2=0.0273bPlasma DDE (ppb)Male GSI (%) Figure 2-16. Linear regression of GSI against blood plasma DDE concentration. No significant correlation was found for females (a, n = 37) or males (b, n = 53).

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53 0 100 200 300 400 0.0 2.5 5.0 7.5 10.0r2=0.0028aPlasma Dieldrin (ppb)Female GSI (%) 0 100 200 300 400 500 0.0 0.5 1.0 1.5r2=0.0024bPlasma Dieldrin (ppb)Male GSI (%) Figure 2-17. Linear regression of GSI against blood plasma dieldrin concentration. No significant correlation was found for females (a, n = 34) or males (b, n = 51).

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54 Estradiol11-KT 0 250 500 750 1000Hormone Concentration(pg/ml)a Estradiol11-KT 0 500 1000 1500Hormone Concentration(pg/ml)b Figure 2-18. Day 0 plasma hormone concentrations. Mean SD (a) female, n = 59 (b) male, n = 88.

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55 0 100 200 300 400 0 500 1000 150011-KT E2 ar2=0.2815r2=0.0351Blood Plasma DDE (ppb)Hormone Concentration(pg/ml) 0 100 200 300 400 -100 0 100 200 300E2 11-KT r2=0.1401r2=0.0000bBlood Plasma DDE (ppb)% Change Figure 2-19. Day 30 Hormones Female p,p-DDE treated (n = 30). Linear regression against DDE dose (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30. Significant negative relationships exist between DDE dose and circulating E 2 and percent change in E 2

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56 0 100 200 300 400 0 1000 2000E2 11-KT ar2=0.0468r2=0.0008Blood Plasma DDE (ppb)Hormone Concentration(pg/ml) 0 100 200 300 400 -100 0 100 200 300 400E2 11-KT r2=0.0114r2=0.0329bBlood Plasma DDE (ppb)% Change Figure 2-20. Day 30 Hormones Male p,p-DDE treated (n = 47). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to day 30 against DDE dose. No significant relationships were found.

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57 0 250 500 750 1000 0 500 1000 1500E2 11-KT r2=0.1588r2=0.0240aBlood Plasma DDE (ppb)Hormone Concentration(pg/ml) 0 250 500 750 1000 -100 0 100 200E2 11-KT r2=0.0482r2=0.0128Blood Plasma DDE (ppb)% Changeb 0.0 2.5 5.0 7.5 10.0 0 500 1000 1500r2=0.0791cFemale GSIFemale Estradiol (pg/ml) Figure 2-21. Day 60 Hormones Female p,p-DDE treated (n = 35). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 30 to Day 60 against p,p-DDE dose and (c) ciculating E 2 to GSI. A significant negative relationship was found between DDE dose and circulating E 2

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58 0 250 500 750 1000 1250 1500 0 1000 2000E2 11-KT r2=0.0216r2=0.0116aBlood Plasma DDE (ppb)Hormone Concentration(pg/ml) 0 250 500 750 1000 1250 1500 -500 0 500 1000E2 11-KT r2=0.0526r2=0.0145bBlood Plasma DDE (ppb)% Change 0.00 0.25 0.50 0.75 1.00 1.25 0 1000 2000r2=0.0043cMale GSIMale 11-KT (pg/ml) Figure 2-22. Day 60 Hormones Male p,p-DDE treated (n = 51). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 30 to Day 60 against dieldrin dose and (c) circulating 11-KT against GSI. No significant relationships were found.

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59 0 50 100 150 200 0 500 1000 1500E2 11-KT r2=0.0213r2=0.0367aBlood Plasma Dieldrin (ppb)Hormone Concentration(pg/ml) 0 50 100 150 200 -100 0 100 200 300 400E2 11-KT r2=0.0097r2=0.0148bBlood Plasma Dieldrin (ppb)% Change Figure 2-23. Day 30 Hormones Female dieldrin treated (n = 34). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30 against dieldrin dose. No significant relationships were found.

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60 0 50 100 150 200 0 500 1000 1500E2 11-KT r2=0.0898r2=0.0032aBlood Plasma Dieldrin (ppb)Hormone Concentration(pg/ml) 0 50 100 150 200 -100 0 100 200E2 11-KT r2=0.0137r2=0.1627bBlood Plasma Dieldrin (ppb)% Change Figure 2-24. Day 30 Hormones Male dieldrin treated (n = 42). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30 against dieldrin dose. A significant negative relationship exists between dieldrin dose and percent change in 11-KT from Day 0 to Day 30.

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61 0 100 200 300 400 500 600 0 500 1000 1500E2 11-KT Blood Plasma Dieldrin (ppb)Hormone Concentration(pg/ml)ar2=0.1597r2=0.0003 0 100 200 300 400 500 600 -200 -100 0 100 200 300E2 11-KT r2=0.0001br2=0.0007Blood Plasma Dieldrin (ppb)% Change 0.0 2.5 5.0 7.5 10.0 0 500 1000 1500r2=0.1110cFemale GSIFemale Estradiol (pg/ml) Figure 2-25. Day 60 Hormones Female dieldrin treated (n= 34). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 30 to Day 60 against blood plasma dieldrin and (c) circulating E 2 against GSI. A significant negative relationship between blood plasma dieldrin and circulating E 2 was found.

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62 0 100 200 300 400 500 600 700 0 500 1000 1500E2 11-KT aBlood Plasma Dieldrin (ppb)Hormone Concentration(pg/ml)r2=0.2212r2=0.0818 0 100 200 300 400 500 600 700 -100 0 100 200E2 11-KT r2=0.0096r2=0.0675bBlood Plasma Dieldrin (ppb)% Change 0.00 0.25 0.50 0.75 1.00 1.25 0 500 1000 1500r2=0.0030cMale GSIMale 11-KT (pg/ml) Figure 2-26. Day 60 Hormones Male dieldrin treated (n = 51). Linear regression of (a) circulating hormones (b) percent change in hormone concentration from Day 0 to Day 30 against dieldrin dose and (c) circulating 11-KT against GSI. Significant negative relationships between circulating 11-KT and dieldrin dose, as well as, between percent change in E 2 from Day 30 to Day 60 were found.

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CHAPTER 3 ACCUMULATION OF DIETARY p,p-DDE AND DIELDRIN BY LARGEMOUTH BASS: A PILOT STUDY Introduction Large variability in the amount of chemical released from pellets placed intraperitoneally (see Chapter 2) into largemouth bass led to the need for developing and validating a new exposure method. The primary exposure route for fish to persistent pesticides is accumulation through the food web (Woodwell et al., 1967; Macek and Korn, 1970; Ruus et al., 2002). This research describes a pilot study aimed to determine the accumulation rate of p,p-DDE and dieldrin in largemouth bass fed a contaminated diet. A review of the literature provided varying degrees of dietary accumulation of either pesticide based on species and dose range tested. Previous studies on pesticide uptake were performed by mixing a pesticide with an oil carrier and coating the feed pellets with the oil solution. Fish were fed at a range of 0.5 4.0% body weight per day, depending on species. No studies were found that addressed p,p-DDE specifically, but several evaluated DDT accumulation. When fed a DDT contaminated diet, rainbow trout accumulated 20 24% of the DDT dose available over 140 days (Macek et al., 1970); brook trout (Salvelinus fontinalis) accumulated 35% over 120 days (Macek and Korn, 1970); and Atlantic menhaden (Brevoortia tyrannus) accumulated 17 27% over 48 days (Warlen et al., 1977). Dietary dieldrin accumulation was around 10% for rainbow trout over 140 days (Macek et al., 1970) and striped bass, Morone saxatilis, over 84 days (Santerre et al., 1997). 63

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64 Another point of investigation in this study was the variation in accumulation rates between two types of feed: sinking and floating. Application of pesticide to floating feed is done by mixing the chemical with an oil carrier and using the oil as a top dressing on the feed pellets. Sinking feed is manufactured by grinding up the feed pellets and mixing the chemicals directly in with the pellet powder, then reconstituting the powder back into pellets. The sinking pellets, presumably, have less pellet-to-pellet variation in achieved chemical dose, however largemouth bass are usually top feeders and it was unclear if they would consistently feed on sinking pellets. Validation of a dietary exposure method was intended to provide useful information on the accumulation rates of p,p-DDE and dieldrin, variability inherent to the method, and difference between pellet type. This study was also intended to determine uptake and elimination rates of these OCPs over time. Specific Aims To determine the bioaccumulation rate and final whole body concentration of p,p-DDE and dieldrin in largemouth bass when exposed orally for 30 and 50 days by contaminated feed. To determine if an internal steady state dose is achieved after 30 and 50 days. To determine the variation in achieved whole body concentration for largemouth bass, both within a tank and among replicates, fed different feed types: sinking or floating. To determine if either chemical induced a change in circulating sex steroid hormones at the achieved doses. Null Hypothesis Accumulation rates and circulating sex steroid levels will not vary between individual fish or feed type.

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65 Alternative Hypothesis Because largemouth bass within each replicate are of varying size and have varying degrees of aggressive food seeking behavior, accumulation rates will vary between individual fish, treatment replicate, and feed type. Feeding aggression may be an important factor for sinking feed as largemouth bass typically will not feed off the bottom, introducing a greater possibility for variability in accumulation rates among individual fish fed sinking feed. In addition, exposure of bass to p,p-DDE and dieldrin, potential endocrine modulators, may decrease the concentration of circulating sex steroids in treated versus control fish. Materials and Methods Experimental Animals Largemouth bass, one to two years of age, were obtained from a fish hatchery (American Sports Fish, AL) in November 2002. The fish were transferred to the United States Geological Survey Biological Resource Division Center for Aquatic Resource Studies (USGS-BRD-CARS) facility where they were weighed and measured, then housed in groups of 14 fish in 700 L round tanks equipped with a flow-through system supplied by on-site well water and aeration. On Day 0 of the experiment (December 5, 2002) bass had an average weight, length, and condition factor SD of 181.6 34.2 g, 233.2 13.3 mm, and 1.42 0.10, respectively, indicating fish were healthy and of reproductive size. Water quality was measured as dissolved oxygen, ammonia, temperature and pH twice a week at a depth of approximately 15 cm, early in the morning (approximately 8 9 a.m.) when the dissolved oxygen is at its lowest. Water quality parameters were all within acceptable ranges for the duration of the experiment: dissolved oxygen ranged

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66 from 7.49 to 9.35 mg/L, ammonia content stayed below 1 ppm, temperature ranged from 17.4 to 20.8 C, and pH ranged from 7.8 to 8.1. Chemicals and Feed The organochlorine pesticides p,p-DDE (2,2-bis(4-chlorophenyl)-1,1-dichloroethylene, Lot # 09020KU, 99.4% purity) and dieldrin (1,2,3,4,10,10-hexachloro-6,7-epoxy-1,4,4a,5,6,7,8,8a-octahydro-1,4,5,8-dimethanonaphthalene, Lot # 077H3578, 91.2% purity) were obtained from Aldrich Chemical Company (Milwaukee, WI). The pesticides were shipped to Zeigler Brothers, Inc. (Gardners, PA) where they were incorporated into feed pellets, sinking and floating for both chemicals, at concentrations of 5 ppm p,p-DDE and 1 ppm dieldrin. Contaminated sinking pellets were manufactured by incorporating a measured amount of chemical into sinking pellets that had been ground to a powder, then mixed thoroughly. The feed powder was then reconstituted into pellets. Contaminated floating feed was manufactured by incorporating a measured amount of chemical into a fish oil mixture that was used to top dress the pellets. Oil mixture and pellets were mixed thoroughly in a mixer to achieve consistent coating of all pellets. A control was also manufactured for each type of feed. The oil control received a top dressing of pure fish oil and the sinking feed was ground and reconstituted. Each chemical and feed type was tested in duplicate (Figure 3-1). One sample of each control type and duplicate samples of all treated feed types were sent to New Jersey Feed Laboratory, Inc. (Trenton, NJ) by Zeigler Feed, Inc. for a chlorinated pesticide and PCB (polychlorinated biphenyl) screen analysis. All feed had a background level of aldrin at 0.03 ppm. Sinking and floating controls had no other detectable OCPs or PCBs. Sinking p,p-DDE (target 5 ppm) and dieldrin (target 1 ppm) feeds had actual mean SD concentrations of 3.6 0.7 and 0.89 0.25, respectively.

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67 Floating p,p-DDE (target 5 ppm) and dieldrin (target 1 ppm) feeds had actual concentrations of 3.9 0.3 and 1.07 0.05, respectively. Although sinking feed was designed to deliver a more accurate and less variable dose, it actually was farther from the target dose and more variable per pellet than floating feed. Feeding Rate The average weight of largemouth bass in this study was used to determine the feeding rate. Contaminated or control feed was administered to each tank at 5% mean body weight for fish (14) within a tank per day (recommended rate from Rick Stout, FFWCC-Fish Hatchery, Richloam, FL). Day 30 and Day 50 Endpoints Weight, to the nearest gram, and length in millimeters, were measured; blood was collected; and gonads were removed and weighed from a subset of 3 fish per tank on Day 30 and all remaining fish (n = 11) on Day 50. Condition factor was calculated (K = weight/length 3 x 100,000). Approximately 1 ml of blood was obtained from the caudal vein using a heparinized 20-gauge needle, dispensed into a heparinized vacutainer, labeled and stored on ice until centrifuged. Blood samples were centrifuged at approximately 1,000x g, 4C for 20 min to separate red blood cells from plasma. Plasma was removed with a transfer pipette, placed in a cryovial and stored at 0C until assayed for circulating sex steroids. Each fish carcass (and gonad for Day 50 fish) was wrapped in aluminum foil, placed in a plastic bag and labeled. Carcass (n = 3 for both Day 30 and 50) and gonads (n = 3, Day 50 only) were sent for GC-MS analysis of p,p-DDE and dieldrin.

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68 Analysis of Largemouth Bass Tissue for OCPs Analysis was conducted by the Center for Environmental and Human Toxicology, University of Florida. Briefly, the largemouth bass carcass/tissue was homogenized and a portion of the sample (2-5 grams) was extracted into ethyl acetate. The sample was cleaned-up using C18 and NH2 SPE (solid phase extraction) cartridges. Total OCP content was determined by GC-MS, following EPA method 8270 (EPA, 1983). Samples were analyzed multiple times in full scan mode for analyte identification and in selected ion mode (SIM) for quantitation to improve sensitivity. Percent recovery ranged between 65 and 100% with a limit of detection of 0.75-1.5 ng/g. Analysis of Circulating Sex Steroid Hormones Blood plasma from largemouth bass was analyzed for 11-KT and E 2 using a previously validated 3 H radioimmunoassay (RIA) method (Gross et al., 2000). All samples were assayed in duplicate and values reported as pg/ml of plasma. Standard curves were prepared in phosphate buffered saline plus gelatin and sodium azide (PBSGA) buffer with known amounts (15, 30, 60, 125, 250, 500, 1000, and 2000 pg) of radioinert E 2 (ICN Biomedicals, Costa Mesa, CA) or 11-KT (Sigma Chemicals, St. Louis, MO) and 15,000 cpm of 3 H-E 2 or 3 H-11-KT. Each plasma sample (50 l) was extracted twice with diethyl ether prior to RIA analysis. PBSGA buffer and antibodies against sex steroid hormones were added to the sample tube and incubated overnight at 4C. Antibodies were purchased from ICN Biomedicals (E 2 ) or Helix Biotech, Richmond, BC, Canada (11-KT). Following the antibody incubation unbound antibody was separated out by addition of dextran-coated charcoal and centrifugation for 10 minutes at 1,000 x g. Binding to the charcoal in the pellet stripped excess free antibody out of solution. Four hundred l of sample supernatant was removed and added to a

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69 scintillation vial with 4 ml of Scintiverse scintillation cocktail (Fisher Scientific, Pittsburgh, PA). Samples vials were then placed in a liquid scintillation counter (Pachard Tricarb, Model 1600) and counted for two minutes each. The minimum concentration distinguishable from zero (mean SE) 148 18 pg/ml for E 2 and 29 18 pg/ml for 11-KT. Cross-reactivities (produced and characterized by T.S. Gross, University of Florida) of the E 2 antiserum with other steroids were: 11.2% for estrone, 1.7% for estriol, and < 1.0% for 17-estradiol and androstenedione. Cross reactivity of the 11-KT antiserum with other steroids was: 9.7% for testosterone, 3.7% for -dihydrotestosterone, and < 1.0% for androstenedione. Statistical Analysis All parameters were analyzed using the Statistical Analysis System (SAS), version 9. All data were run first by the Means Procedure to determine averages and standard deviations for all treatment groups and replicates. All data presented in the results are means SD. To determine any statistical significance between replicates, feed type, and chemical, an analysis of variance (ANOVA) procedure, followed by the multiple comparisons procedure, Duncans Multiple Range Test, for all parameters (SAS) to determine differences between replicates. Significance was declared at = 0.05. Results Largemouth Bass There was no mortality of largemouth bass throughout the duration of the study. The condition factor at Days 30 and 50 were 1.45 0.09 and 1.42 0.10, respectively, and did not differ between feed types (sinking or floating) or pesticide administered (p,p-DDE or dieldrin). This indicates a healthy population was maintained throughout the

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70 exposure period. Largemouth bass grew an average of 3 0.4 mm and gained an average of 2 0.8 g over the course of the study. Feeding Rate Largemouth bass used in this study did not eat at the recommended rate of 5% body weight per day. Instead, pellets in the amount of 5% of tanks total body weight were offered and the pellets remaining after approximately 10 min were counted. From the number of remaining pellets, an estimate of total consumption was made for each tank on each day (Table 3-1 and 3-2). Over time it became apparent that the fish were eating at an approximate rate of 1% body weight per day, and so the feeding rate was adjusted to 1%. Total weight of pellets consumed over 30 or 50 days was used to determine total exposure. 30-Day Exposure p,p-DDE accumulation After 30 days exposure, bass in the control and dieldrin groups had a background level of 7 1.7 ppb of p,p-DDE which did not vary among tanks, but did significantly differ from fish fed p,p-DDE contaminated feed types; sinking and floating (Figure 3-2A). Replicate 1 (Tank 3) for sinking p,p-DDE contaminated feed was fed approximately 115 g of p,p-DDE per fish over the course of 30 days. The whole-body concentration was 495 144 ppb, equal to 81 15% accumulation. Replicate 2 (Tank 4) was fed an average of 114 g of p,p-DDE per fish over the course of 30 days. However, the whole-body concentration was 350 77 ppb, equal to only 54 3.7% accumulation. The whole-body concentrations between replicates for sinking p,p-DDE contaminated

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71 feed were significantly different from each other (Figure 3-2,A) but the percent of the total dose accumulated was not significantly different between replicates (Figure 3-3). Replicates 1 and 2 (Tanks 5 and 12) for floating p,p-DDE contaminated feed consumed an average of 109 and 106 g of p,p-DDE per fish, respectively, over the course of 30 days. The whole-body concentrations were 341 51 and 347 41 ppb, respectively, equating to 65 24.5% and 49 19.1% accumulation of p,p-DDE (Figures 3-2 and 3-3). The whole-body concentration and percentage of total dose accumulated between replicates for floating p,p-DDE contaminated feed were not significantly different from each other. Overall, one replicate of the sinking p,p-DDE treatment was higher in whole-body concentration and percent accumulation relative to all other replicates, sinking and floating. The fish fed floating p,p-DDE diets had a lower standard deviation from the mean, indicating less variability among fish within a replicate. Dieldrin accumulation After 30 days exposure, bass in the control and p,p-DDE groups had a background level of 1.7 0.34 ppb of dieldrin, which did not vary among tanks, but did significantly differ from all fish fed dieldrin contaminated feed; sinking and floating (Figure 3-2B). Replicates 1 and 2 (Tanks 6 and 10) for sinking dieldrin contaminated feed were fed an average of 23 g of dieldrin per fish, respectively, over 30 days. Whole-body concentration determined by GC-MS were 148 96 and 88 30 ppb, respectively, equating to 74 24% accumulation for replicate 1 and 82 34% accumulation for replicate 2. The whole-body concentration and percentage of dose accumulated were not

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72 significantly different between replicates for sinking dieldrin contaminated feed (Figures 3-2, B and 3-3). Replicates 1 and 2 (Tanks 2 and 8) for floating dieldrin contaminated feed were fed an average of 19 and 21 g of dieldrin per fish, respectively, over 30 days. Whole-body concentrations, as determined by GC-MS, were 114 7 and 120 6 ppb, respectively, equal to accumulation rates of 93 16% for replicate 1 and 89 17% for replicate 2. The whole-body concentration and percentage of dose accumulated did not differ between replicates (Figures 3-2, B and 3-3). Overall, the whole-body concentration and percent accumulation did not differ between replicates and feed types, sinking or floating. However, the standard deviation from the mean was smaller for replicates of the floating dieldrin diet. GSI and sex steroids GSI for males (n = 20, 0.42 0.07) and females (n = 16, 1.5 0.28) in all treated groups was not significantly different from controls after 30 days exposure (analysis not shown). Blood plasma concentrations of sex steroids in both females and males did not vary significantly by treatment after 30 days (data not shown). Because only 3 fish from each treatment replicate were sampled, the sample size for either males or females ranged from 0-3 depending on what was sampled for that replicate. No valid conclusions can be drawn from the statistical analysis performed due to small sample size. 50-Day Exposure The data from 30 days exposure indicated that there was less variability (lower standard deviation from the mean) among fish within a tank, as well as among replicates,

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73 for the floating pellet feed type. Use of sinking pellets was discontinued after 30 days and fish were sacrificed. Contaminant analysis was run only for a subset of three fish from each replicate of either a floating p,p-DDE diet or a floating dieldrin diet at Day 50. It was assumed that the background levels of either chemical in the control fish did not change from Day 30 values and control bass were not analyzed for contaminants. p,p-DDE accumulation After 50 days exposure, bass in the dieldrin group accumulated a background of 6.5 0.5 ppb of p,p-DDE, which did not vary among tanks, was not significantly different from Day 30 controls, but did significantly differ from all fish fed floating p,p-DDE contaminated feed (Figure 3-4, A). Fish from replicate 1 (Tank 5) were fed approximately 150 g p,p-DDE over the course of 50 days. The whole-body concentration after 50 days was 314 30.5 ppb, equal to an accumulation of 36 9% of the total dose administered. Replicate 2 (Tank 12) was fed 150 g p,p-DDE per fish over 50 days and had a final whole-body concentration of 337 41 ppb, equal to an accumulation of 48 15% of the total dose. Analysis of whole-body concentration and percentage of total dose accumulated between replicates revealed no significant differences (Figures 3-4, A and 3-5). Gonadal concentration of p,p-DDE was determined to be 420 103.8 ppb for replicate 1 and 482 74.3 ppb for replicate 2 (Figure 3-6). Although gonad concentrations demonstrated greater variability, means of the replicates were not significantly different. The gonadal dose contributed approximately 0.5% and 1.0% of the whole-body concentration for males and females, respectively.

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74 Overall, a diet consisting of floating pellets contaminated with p,p-DDE produced a fairly consistent dosing method over the 50 day exposure time. Dieldrin accumulation After 50 days exposure, bass in the p,p-DDE group accumulated a background of 1.1 0.4 ppb of dieldrin which did not vary among tanks, was not significantly different from Day 30 controls, but did significantly differ from all fish fed a diet of floating dieldrin contaminated feed (Figure 3-4, B). Each fish from replicates 1 and 2 (Tanks 2 and 8) were fed approximately 26 and 31 g dieldrin, respectively, over the course of 50 days. The average whole-body concentration after 50 days was 92 13 ppb for replicate 1 and 116 20 ppb for replicate 2, equating to accumulations of 61 15% and 64 14% of the total dose administered, respectively. The whole-body concentration of dieldrin for fish in replicate 2 were significantly higher than the fish in replicate 1, which is to be expected as replicate 2 was fed approximately 20% more dieldrin (Figures 3-4A). The percentage of the total dose accumulated did not differ significantly between replicates (Figure 3-5). Gonadal dieldrin concentrations of fish in replicates 1 and 2 were 196 64.9 and 547 144 ppb, respectively (3-6). Corresponding to the whole-body concentration, gonadal dose of dieldrin for fish in replicate 2 was significantly higher, nearly double that of replicate 1. The gonadal dose contributed approximately 1.0% to the whole-body concentration of replicate 1 and 3.5% to the total internal dose of replicate 2 with no apparent difference between males and females. GSI and sex steroids GSI for males (n = 44, 0.40 0.11) and females (n = 22, 1.31 0.46) did not vary among replicates or chemical administered.

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75 Female sex steroids, both E 2 and 11-KT, did not significantly differ between replicates for any treatment. E 2 concentrations were significantly lower for both p,p-DDE and dieldrin treatment as compared to the controls, however, there was no difference in 11-KT for either treatment in females as compared to controls (Figure 3-7). Floating control replicate 1 had three females with means of 386 82.4 pg/ml for E 2 and 78 21.4 pg/ml for 11-KT. Floating control replicate 2 had four females with means of 404 61.9 pg/ml for E 2 and 115 13.6 pg/ml for 11-KT. p,p-DDE replicates 1 and 2 had five and three females with means of 215 86.6 and 153 38.1 pg/ml for E 2 and 25 13.3 and 38 46.8 pg/ml for 11-KT, respectively. Dieldrin replicates 1 and 2 had four and three females with means of 125 100.1 and 158 86.3 pg/ml for E 2 and 20 24.4 and 10 5 pg/ml for 11-KT, respectively. Male sex steroids did not significantly differ between replicates of any treatment and for both p,p-DDE and dieldrin treated males, E 2 and 11-KT were significantly lower than controls (Figure 3-8). Floating control replicate 1 had eight males with means of 289 40.8 pg/ml for E 2 and 270 191.9 pg/ml for 11-KT. Floating control replicate 2 had seven males with means of 238 48.3 pg/ml for E 2 and 251 102.6 pg/ml for 11-KT. p,p-DDE replicates 1 and 2 had six and seven females with means of 137 136.8 and 102 60.7 pg/ml for E 2 and 79 92.2 and 66 15.7 pg/ml for 11-KT, respectively. Dieldrin replicates 1 and 2 had seven and eight males with means of 73 30 and 78 48 pg/ml for E 2 and 117 106 and 32 9.2 pg/ml for 11-KT, respectively. Discussion The results of this study suggest that diets containing OCPs provide an accurate route of administration for largemouth bass with only moderate variability within replicates. Sinking style feed produced greater variability both within and among

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76 replicates for both pesticides and was most likely due to competition among fish for the pellets before they sank to the bottom of the tank. This style of pellet is not conducive to creating consistent doses in all fish within a tank. The floating style feed pellets showed greater consistency in achieved whole body dose. Standard deviations from the mean, as well as replicate differences, were lower in tanks fed floating style feed. The statistically significant difference in whole body concentration seen at Day 50 in the dieldrin treated groups occurred because of the difference in the total amount of dieldrin fed between the tanks; the percentage of total dose accumulated did not differ. The consistency observed within the floating feed type replicates indicates the superiority of floating feed over sinking feed as an exposure route despite the suggestion that the method of preparing oil dressed pellets is inherently less accurate. The remainder of the discussion will focus only on the results obtained from fish fed a floating style diet. The rapid uptake of both p,p-DDE and dieldrin from the diet of largemouth bass is consistent with studies by Grzenda et al. (1971) that demonstrated rapid uptake for both pesticides in goldfish fed contaminated diets. In these studies, even though exposure continued past the initial uptake phase, accumulation rate showed a tendency to level off after 30 days of exposure and tissues maintained a consistent dose, indicating the rate of elimination was equal to that of the dosing (i.e. a steady state was reached). In Atlantic menhaden, Brevoortia tyrannus, (Warlen et al., 1977), no equilibrium point was found within a 60 day oral exposure to DDT. Studies on rainbow trout show that equilibrium for both DDT and dieldrin did not occur until 140 days after initial dietary exposure (Macek et al., 1970). Variability in absorption and elimination strategies may explain the differences seen between fish species.

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77 While this study measured whole body concentrations after just two time points, it is clear that internal dose of either pesticide did not increase after Day 30 despite continued daily dosing (Figure 3-9). Therefore, results suggest that largemouth bass may undergo a rapid uptake phase followed by increased elimination. Absorption efficiency from the intestine of fish can be greater than 50% for lipophilic compounds such as OCPs (James and Kleinow, 1994). The percentage of the total dose accumulated at Day 30 in this study was approximately 50% for p,p-DDE and 90% for dieldrin. The accumulation rates dropped by approximately 10% and 30%, respectively, by Day 50. This could be due to a decrease in absorption efficiency or an increase in biotransformation and elimination of the absorbed dose. Most likely, elevated elimination rates are the cause for the decrease in overall accumulation; OCPs are known to induce liver biotransformation enzymes at high doses in fish (Machala et al., 1998; Zapata-Perez et al., 2000; Foster et al., 2001) and mammals (Dickerson et al., 1999; Sanderson et al., 2002; Leavens et al., 2002). Inconsistent with all other studies on accumulation of DDE/DDT and dieldrin in various fish species was the total percentage of the accumulated dose. For DDT, uptake was stated to be between 20 35% in several studies on rainbow trout, brook trout and Atlantic menhaden over periods of 140, 120, and 48 days, respectively (Macek et al., 1970; Macek and Korn, 1970; Warlen et al., 1977). The largemouth bass in this study accumulated a total of 35 48% of the p,p-DDE dose at Day 50, approximately 10% more than previously reported. For dieldrin, studies with rainbow trout (Macek et al., 1970) and striped bass (Santerre et al., 1997) showed only a 10% accumulation over 140 and 84 days, respectively, while the largemouth bass in this study accumulated 60 64%

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78 of the total dieldrin dose after 50 days. Of mention is that the DDE/DDT and dieldrin accumulation studies cited were generally of much longer durations and only reported final percent accumulation, which may account for the smaller total accumulated dose than reported here. The higher rate of accumulation and/or retention of lipophilic compounds such as OCPs may also be due to the high fat content of largemouth bass, especially at the time period in which this experiment was conducted. The lipid content of largemouth bass is between 6 10%, increasing during the spawning season due to the development of the gonad (Barziza and Gatlin, 2000; Brecka et al., 1996). Reported lipid contents for rainbow trout and herring (Atlantic menhaden is in the herring family) are most similar to the bass at approximately 6 and 9%, respectively while carp is 1 4% (Ackman, 1995). Species variation in absorption efficiency may also play a large role. Gonad concentrations of both OCPs had large variability within replicates, but generally related to a consistent percentage of the whole-body concentration. p,p-DDE showed a greater partitioning into the female gonad than the male most likely due to the high lipid content of the developing oocytes within the female gonad. Dieldrin did not show such a difference in partitioning between the sexes, however, there was a replicate difference due to different total doses of dieldrin fed over 50 days. It seems that as the available dieldrin dose increases, the greater the percentage is that is stored in the fatty tissue of the gonad. The sample size for the estimations of OCP concentrations within a tank was small (n = 3) and there were not enough animals to determine differences between sexes in one replicate. A larger sample size would aid in determining if a dieldrin partitioning difference exists between sexes.

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79 Nonetheless, the achieved whole-body concentrations of both p,p-DDE and dieldrin demonstrated an effect on the endocrine system of treated fish. This manifested as a decrease in circulating E 2 in female and decreases in both E 2 and 11-KT in males after 50 days of exposure. From the data collected at Day 30, it is difficult to determine if there were any trends in sex steroid levels due to treatment because of the uneven sex ratio and small sample size collected. The GSI and circulating level of sex steroids from Day 50 indicate that the fish used in this study were not sexually mature, as GSI levels were 5 10 times less than normal and hormone levels were 10-fold less than what has been shown for both male and female adult hatchery-raised bass (Gross et al., 2002). In addition, GSI in males and females did not increase from Day 30 to Day 50. In fact, GSIs from those time points did not significantly differ for males or females (Figure 3-10) as normally occurs during this time period. The endocrine disruption seen in these fish may not be of the same type or magnitude found in sexually mature fish containing the same dose. Depressed sexual development may have been due to stress induced by placement of fish in captive conditions. Whole-body concentrations of treated fish, achieved by Day 30 and maintained until Day 50, were approximately 375 ppb for p,p-DDE and 100 ppb for dieldrin. Gonadal concentrations determined for three fish per replicate were highly variable and inconsistent between replicates for dieldrin (Figure 3-6). However, when the average gonadal doses of pesticide achieved in this study are compared to gonadal doses of fish found in affected areas of the Ocklawaha River basin, p,p-DDE levels were lower by 10-fold and dieldrin levels were approximately equal (Marburger et al., 1999). (See Chapter 2 discussion for theories of endocrine disruption by OCPs.)

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80 In summary, exposing largemouth bass to OCPs by incorporation into an oil carrier and subsequent coating of feed pellets is an effective, accurate, and reproducible dosing method. Variability in achieved whole-body concentration and percentage of total dose accumulated was minimal for treatment with both p,p-DDE and dieldrin. A rapid uptake followed by a sustained whole-body concentration after 30 days indicates that study length should be a minimum of one month. In addition, effects of these OCPs on endocrine function during the onset of reproductive season were not manifest until after 50 days of continuous exposure, despite maintenance of approximately the same whole-body concentration. This indicates that the study length should be around two months and begin before January to obtain appropriate exposure duration to mimic chronic exposure in the wild before the onset of spawning. Of future interest may be the study of induction of biotransformation enzymes over the course of the exposure period. Further studies examining OCP accumulation (absorption and elimination) over a longer period of time should also be done to optimize the oral exposure method for largemouth bass.

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81 8 7 Dieldrin Control Floating Sinking 5 6 DDE Dieldrin Floating Sinking 12 11 4 3 DDE Control DDE DDE Floating Floating Sinking Sinking 9 10 1 2 Control Dieldrin Control Dieldrin Sinking Sinking Floating Floating Figure 3-1. Tank setup for preliminary feeding study. Each set of four tanks had a divided well water and air supply line and separate drainage lines.

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82 Table 3-1. Ten-day increment totals of feed eaten per tank. The total feed eaten by each tank was totaled at Day 30 and used to estimate the average amount of pesticide ingested by each fish. Treatment Tank Feed Days 1-10 (g) Feed Days 11-20 (g) Feed Days 21-30 (g) Day 30 Total Feed (g) Grams per Fish Tank DDE (g) Per Fish DDE (g) Tank Dieldrin (g) Per Fish Dieldrin (g) Control Floating 1 196.3 174.3 175.3 916.6 36.7 0 0 0 0 Dieldrin Floating 2 182.4 150.8 138.9 805.2 32.2 0 0 805.15 32.2 DDE Sinking 3 201.1 197.8 175.0 972.8 38.9 4863.9 194.6 0 0 DDE Sinking 4 192.4 201.4 175.3 962.8 38.5 4813.8 192.6 0 0 DDE Floating 5 198.0 173.3 175.3 917.9 36.7 4589.7 193.6 0 0 Dieldrin Sinking 6 199.9 193.4 174.0 960.7 38.4 0 0 960.7 38.4 Control Sinking 7 200.8 187.7 170.3 947.2 37.9 0 0 0 0 Dieldrin Floating 8 189.2 163.8 175.3 881.4 35.3 0 0 881.4 35.3 Control Sinking 9 194.6 200.5 175.3 961.6 38.5 0 0 0 0 Dieldrin Sinking 10 191.0 195.6 172.8 946.1 37.8 0 0 946.1 37.8 Control Floating 11 200.8 172.2 175.2 921.1 36.8 0 0 0 0 DDE Floating 12 192.9 161.2 175.3 883.5 35.3 4417.4 176.7 0 0

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83 Table 3-2. Ten-day increment totals of feed eaten per tank (floating only after Day 30). The total feed eaten by each tank was totaled at Day 50 and used to estimate the average amount of pesticide ingested by each fish. Treatment Tank Feed Days 31-40 (g) Feed Days 41-50 (g) Day 50 Total Feed (g) Grams per Fish Tank DDE (g) Per Fish DDE (g) Tank Dieldrin (g) Per Fish Dieldrin (g) Control Floating 1 106.7 103.3 1123.2 51.1 0 0 0 0 Dieldrin Floating 2 73.6 74.7 930.5 42.3 0 0 536.7 25.6 DDE Floating 5 90.4 87.8 1076.4 48.9 5381.8 244.6 0 0 Dieldrin Floating 8 108.2 104.1 1096.2 49.8 0 0 677.3 30.8 Control Floating 11 96.6 105.4 1109.3 50.4 0 0 0 0 DDE Floating 12 102.1 90.2 1071.8 48.7 5358.9 243.6 0 0

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84 Sinking Control-1Sinking Control-2Floating Control-1Floating Control-2Sinking DDE-1Sinking DDE-2Floating DDE-1Floating DDE-2 0 250 500 750Sinking Floating abbbccccAp,p'-DDE (ppb) Sinking Control-1Sinking Control-2Floating Control-1Floating Control-2Sinking Dieldrin-1Sinking Dieldrin-2Floating Dieldrin-1Floating Dieldrin-2 0 50 100 150Sinking Floating aaaabbbbbbBDieldrin (ppb) Figure 3-2. Day 30 whole-body concentration SD of p,p-DDE (A) and dieldrin (B) in largemouth bass fed a contaminated diet (n = 3 per treatment). All dosed fish had significantly higher whole-body concentrations of their respective chemical than the controls. One replicate fed sinking p,p-DDE contaminated feed was significantly higher than all other p,p-DDE fed fish. There were no significant differences among replicates fed dieldrin contaminated diets. Bars with different letter designations indicate a significant difference.

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85 DDEDieldrin 0 50 100 150Sinking Floating aba,ba,baaaa% Accumulation Figure 3-3. Day 30, percentage of total dose accumulated SD within each replicate (n = 3 per replicate). There were no significant differences in the amount accumulated between replicates of either feed type for dieldrin. There was a significant difference between one DDE sinking diet replicate and one DDE floating diet replicate. Bars with different letter designations indicate a significant difference.

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86 Control Floating-1Control Floating-2DDE Floating-1DDE Floating -2 0 100 200 300 400aabbAp,p'-DDE (ppb) Control Floating-1Control Floating-2Dieldrin Floating-1Dieldrin Floating-2 0 50 100 150abccBDieldrin (ppb) Figure 3-4. Day 50 whole-body concentration SDof p,p-DDE (A) or dieldrin (B) in largemouth bass fed a contaminated diet (n = 3 per replicate). All dosed fish had significantly higher whole-body concentrations of their respective chemical than the controls. There was no significant difference among p,p-DDE replicates. Whole-body concentration was significantly different between dieldrin replicates. Bars with different letter designations indicate a significant difference.

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87 DieldrinDDE 0 25 50 75 100Floating-1 Floating-2 aaaa% Accumulation Figure 3-5. Day 50, percentage of total dose accumulated SD within each replicate (n = 3 per replicate). There were no significant differences in the amount accumulated between replicates of either for either p,p-DDE or dieldrin. Bars with different letter designations indicate a significant difference.

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88 DDEDieldrin 0 250 500 750Floating-1 Floating-2 Gonadal Concentration(ppb)aaab Figure 3-6. Gonadal concentrations of OCPs. Replicates of p,p-DDE were not significantly different, however, there was a significant difference in gonadal concentrations of dieldrin between replicates most likely due to the difference in total g of dieldrin fed.

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89 Control Floating-1Control Floating-2Dieldrin Floating-1Dieldrin Floating-2DDE Floating-1DDE Floating -2 0 100 200 300 400 500Control Dieldrin DDE AaabbbbEstradiol (pg/ml)Female Control Floating-1Control Floating-2Dieldrin Floating-1Dieldrin Floating-2DDE Floating-1DDE Floating -2 0 100 200 300 400 500Control Dieldrin DDE aaaaaaB11-Ketotestosterone(pg/ml)Female Figure 3-7. Day 50, Female circulating hormones SD (n ranged from 3 to 5). Significant differences in E 2 were found between both replicates of p,p-DDE and dieldrin fed fish and controls. No significant differences were found in 11-KT levels in females. Bars with different letter designations indicate a significant difference.

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90 Control Floating-1Control Floating-2Dieldrin Floating-1Dieldrin Floating-2DDE Floating-1DDE Floating -2 0 100 200 300 400 500Control Dieldrin DDE aabbbbAEstradiol (pg/ml)Male Control Floating-1Control Floating-2Dieldrin Floating-1Dieldrin Floating-2DDE Floating-1DDE Floating -2 0 100 200 300 400 500Control Dieldrin DDE aabbbbB11-Ketotestosterone(pg/ml)Male Figure 3-8. Day 50, male circulating hormones SD (n ranged from 6 to 8). Significant differences were found in both E2 and 11-KT levels in all treatment groups and replicates as compared to controls. Bars with different letter designations indicate a significant difference.

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91 0.030.050.0 0 100 200 300 400Replicate 1 Replicate 2 ADayWhole carcassp,p'-DDE(ppb) 0.030.050.0 0 50 100 150Replicate 1 Replicate 2 BDayWhole carcass dieldrin(ppb) Figure 3-9. Whole carcass concentration of floating style feed replicates over time, A p,p-DDE, B Dieldrin.

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92 Day 30Day 50 0 1 2AaaFemale GSI Day 30Day 50 0.00 0.25 0.50 0.75BaaMale GSI Figure 3-10. Female and male GSI SD for Days 30 (n = 16 and 20, respectively) and 50 (n = 21 and 44, respectively). There was no significant difference in GSI between Days 30 and 50 when GSI should have increased dramatically. Bars with different designation indicate significant difference.

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CHAPTER 4 GENERAL CONCLUSIONS / FUTURE RESEARCH General Conclusions The data from the pellet study demonstrated that slow-release matrix pellets are an ineffective dosing method for largemouth bass. It was clear from visual observation of the intraperitoneal cavity that in some fish, the pellet did not dissolve completely, if at all, while other fish had only a small piece or none of the original pellet left after 60 days. It is unclear why the p,p-DDE pellets demonstrated a consistent dose-response in gonadal concentration while the dieldrin pellets did not. In any case, slow-release pellets are not recommended as an exposure route in fish. The preliminary oral dosing study was commissioned to develop a new and more consistent dosing method aiming to avoid extreme variability within a treatment group. Top-dressed floating feed was determined to deliver the most consistent dose probably because it allowed bass to follow their instinctive feeding behavior patterns. However, validation of this feed method in a preliminary study does not provide justification for assuming low variability in dose achieved in future studies. Subsets of fish, or any animal involved in a toxicology study, should be analyzed for attained dose whenever endpoints are to be measured in order to be aware of variations in dose within treatment groups if they exist. The research from the studies presented here provides evidence that p,p-DDE and dieldrin are effective endocrine modulators. Data from the oral dosing study indicates that low level exposure at levels similar to what is found in the environment causes a decrease in circulating sex steroids. However, results from the pellet study indicate only 93

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94 high levels of exposure can depress sex steroid synthesis. The two studies are difficult to compare as OCP measurements were made on blood plasma in the pellet study and whole carcass in the second. The studies differed in that the pellet study, only E 2 in females demonstrated a negative relationship with p,p-DDE dose, while the oral dosing study demonstrated a decrease in female E 2 and both E 2 and 11-KT in males. Also inconsistent, both E 2 and 11-KT concentrations were depressed in females and males treated with dieldrin in the oral dosing study, while in the pellet study only E 2 in females and 11-KT in males were depressed. This may be due a difference in the season in which the studies were performed. The feeding study was started in early December when sex steroid levels are just beginning to rise enabling the pesticides to act over a sensitive time-point in the reproductive season. If the mechanism of action is in fact estrogen/androgen mimicking, threshold levels of pesticide at an earlier time-point could have a negative feedback on the sex steroid synthesis pathway before it has a chance to increase steroidogenesis to produce significant quantities of hormone. The threshold OCP effect level may also be lower at earlier stages of the reproductive cycle. Both studies were similar in that neither chemical had an effect on health parameters and that circulating sex steroid levels were 2 4 fold lower than those reported for pond-reared fish, most likely due to captive holding conditions. Finally, apparent threshold doses for endocrine modulation can be estimated from the oral dosing study. The apparent threshold whole body concentration of p,p-DDE and dieldrin delivered by oral dosing early in the reproductive season was 300 ppb and 100 ppb, respectively, correlating to environmentally relevant concentrations.

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95 Future Research Future research on the reproductive effects of p,p-DDE and dieldrin in largemouth bass should focus on elucidating a mechanism of action. Analysis of endpoints involved in the sex steroid hormone synthesis pathway (GnRH and GTH), vitellogenesis, and sex steroid biotransformation and elimination (CYP3A) would provide such biochemical information. Another facet for examination is whether the decrease in circulating sex steroids caused by these pesticides induces a change in 1) the ability of bass to spawn successfully and 2) the quality/hatchability/survival of the offspring. Further development of oral dosing by contaminated feed method is still needed. Longer studies with multiple dose levels should be done to determine when a steady-state of concentration is reached, if it is maintained, and the percent accumulation at higher doses. Lastly, comparison between pesticide exposure studies is difficult because of the variety of exposure measurements that are employed, if at all. A study examining the partitioning of OCPs in the various tissue compartments of largemouth bass is needed in order for correct comparisons between studies can be made.

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LIST OF REFERENCES Ackman, R. G. 1995. Composition and nutritive value of fish and shellfish. pg. 117-156. in A. Ruiter ed. Fish and Fishery Products. Cab International, Wallingford, England. Anderson, H. R., Vinggaard, A. M., Rasmussen, T. H., Gjermandsen, I. M., and Bonefeld-Jorgensen, E. C. 2002. Effects of currently used pesticides in assays for estrogenicity, androgenicity, and aromatase activity in vitro. Toxicology and Applied Pharmacology 179: 1-12. ATRA, Inc., 1997. Environmental risk assessment of a Lake Apopka muck farm wetlands restoration. Published by St. Johns River Water Management District, Palatka, FL. Barziza, D. E. and Gatlin III, D. M. 2000. An evaluation of total body electrical conductivity to estimate body composition of largemouth bass, Micropterus salmoides. Aquatic Living Resources 13: 439-447. Battrup, E. and Junge, M. 2001. Antiandrogenic pesticides disrupt sexual characteristics in the adult male guppy (Poecilia reticulata). Environmental Health Perspectives 109: 1063-1068. Bayley, M., Junge, M., and Baatrup, E. 2002. Exposure of juvenile guppies to three antiandrogens causes demasculinization and a reduced sperm count in adult males. Aquatic Toxicology 56: 227-239. Benton, J. and Douglas, D. 1994. Ocklawaha fisheries investigations completion report, 1991-1994: Study XIII, Assessment of fisheries restoration potential for reclaimed agricultural lands in the upper Ocklawaha Basin; Study XIV, Ocklawaha chain of lakes largemouth bass population studies; Study XV, Black crappie production in Lake Griffin. State of Florida Game and Fresh Water Fish Commission. Benton, J. and Douglas, D. 1996. Ocklawaha Fisheries Investigations: 1 July 1994 through 30 June 1995: Study XIII Assessment of fisheries restoration potential for reclaimed agricultural lands in the Upper Ocklawaha Basin. State of Florida Game and Fresh Water Commission. Benton, John, Douglas, David, and Prevatt, Louis. 1991. Completion report as required by federal aid in fish restoration: Wallop-Breaux project F-30-18, Ocklawaha Basin Fisheries Investigations Study XII. Lake Apopka fisheries studies. 96

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97 Brecka, B. J., Wahl, D. H., and Hooe, M. L. 1996. Growth, survival, and body composition of largemouth bass fed various commercial diets and protein concentrations. The Progressive Fish-Culturist 58: 104-110. Carr, M. H. 1942. The breeding habits, embryology and larval development of the large-mouthed black bass in Florida. Proceedings of the New England Zoological Club 20: 43-77. Celius, T. and Walther, B. T. 1998. Differential sensitivity of zonagenesis and vitellogenesis in Atlantic salmon (Salmo salar L) to DDT pesticides. The Journal of Experimental Zoology 281: 346-353. Chew, Robert L., 1974. Early life history of the Florida largemouth bass. Florida Game and Freshwater Commission. Fish Bulletin No. 7, Tallahassee, FL. Clugston, J. P. 1966. Centrarchid spawning in the Florida Everglades. Quarterly Journal of the Florida Academy of Sciences 29: 137-143. Copeland, P. A., Sumpter J. P., Walker, T. K., and Croft, M. 1986. Vitellogenin levels in male and female rainbow trout (Salmo gairdneri richarson) at various stages of the reproductive cycle. Comparative Biochemistry and Physiology, Part B 83: 487-493. Coumoul, X., Diry, M., and Barouki, R. 2002. PXR-dependent induction of human CYP3A4 gene expression by organochlorine pesticides. Biochemical Pharmacology 64: 1513-1519. Crain, D. A., Guillette, L. J., Rooney, A. A., and Pickford, D. B. 1997. Alterations in steroidogenesis in alligators (Alligator mississippiensis) exposed naturally and experimentally to environmental contaminants. Environmental Health Perspectives 105: 528-533. Danzo, B. J. 1997. Environmental xenobiotics may disrupt normal endocrine function by interfering with the binding of physiological ligands to steroid receptors and binding proteins. Environmental Health Perspectives 105: 294-301. Danzo, B. J., Shappell, H. W., Banerjee, A., and Hachey, D. L. 2002. Effects of nonylphenol, 1,1-dichloro-2,2-bis(p-chlorophenyl)ethylene (p,p'-DDE), and pentachlorophenol on the adult female guinea pig reproductive tract. Reproductive Toxicology 16: 29-43. Dickerson, R. L., McMurry, C. S., Smith, E. E., Taylor, M. D., Nowell, S. A., and Frame, L. T. 1999. Modulation of endocrine pathways by 4,4'-DDE in the deer mouse Peromyscus maniculatus. The Science of The Total Environment 223: 97-108.

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98 Foster, E. P., Fitzpatrick, M. S., Feist, G. W., Schreck, C. B., Yates, J., Spitsbergen, J. M., and Heidel, J. R. 2001. Plasma Androgen Correlation, EROD Induction, Reduced Condition Factor, and the Occurrence of Organochlorine Pollutants in Reproductively Immature White Sturgeon (Acipenser transmontanus) from the Colomia River, USA. Archives of Environmental Contamination and Toxicology 41: 182-191. Gallagher, E. P., Gross, T. S., and Sheehy, K. M. 2001. Decreased glutathione S-transferase expression and activity and altered sex steroids in Lake Apopka brown bullheads (Ameriurus nebulosus). Aquatic Toxicology 55: 223-237. Garcia, E. F., McPherson, R. J., Martin, T. H., Poth, R. A., and Greeley, M. S. 1997. Liver cell estrogen receptor binding in prespawning female largemouth bass, Micropterus salmoides, environmentally exposed to polychorinated biphenyls. Archives of Environmental Contamination and Toxicology 32: 309-315. Gore, A. C. Organochlorine pesticides directly regulate gonadotropin-releasing hormone gene expression and biosynthesis in the GT1-7 hypothalamic cell line. Molecular and Cellular Endocrinology 192: 157-170. Gross, T. S., Guillette, L. J., Percival, H. F., Masson, G. R., Matter, J. M., and Woodward, A. R. 1994. Contaminant-induced reproductive anomalies in Florida. Comparative Pathology Bulletin 26: 2-8. Gross, T. S., Seplveda, M. S., Wieser, C. M., Wiebe, J. J., Schoeb, T. R., Denslow, N. D., and Johnson, W. E. 2000. Characterization of annual reproductive cycles for pond-reared Florida largemouth bass (Micropterus salmoides floridanus). Proceedings of the Black Bass 2000 Symposium, American Fisheries Society, St. Louis, MO, USA 205-212. Grzenda, A. R., Taylor, W. J., and Paris, D. F. 1971. The uptake and distribution of chlorinated residues by goldfish (Carassius auratus) fed a 14 C-Dieldrin contaminated diet. Transactions of the American Fisheries Society 2: 215-221. Guillette, L. J., Brock, J. W., Rooney, A. A., and Woodward, A. R. 1999a. Serum Concentrations of Various Environmental Contaminants and Their Relationship to Sex Steroid Concentration and Phallus Size in Juvenile American Alligators. Archives of Environmental Contamination and Toxicology 36: 447-455. Guillette, L. J., Crain, D. A., Rooney, A. A., and Pickford, D. B. 1995. Organization versus activation: The role of endocrine-disrupting contaminants (EDCs) during embryonic development in wildlife. Environmental Health Perspectives, Supplement 103: 157-164. Guillette, L. J., Gross, T. S., Masson, G. R., Matter, J. M., Percival, H. F., and Woodward, A. R. 1994. Developmental abnormalities of the gonad and abnormal sex hormone concentrations in juvenile alligators from contaminated and control lakes in Florida. Environmental Health Perspectives 102: 680-688.

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99 Guillette, L. J., Woodward, A. R., Crain, D. A., Pickford, D. B., Rooney, A. A., and Percival, H. F. 1999b. Plasma steroid concentrations and male phallus size in juvenile alligators from seven Florida lakes. General and Comparative Endocrinology 116: 356-372. Heinz, G. H., Percival, H. F. and Jennings, M. L. 1991. Contaminants in American alligator egges from Lake Apopka, Lake Griffin, and Lake Okeechobee, Florida. Environmental Monitoring and Assessment 16: 277-285. Huffstutler, K. K., Burgess, J. E., and Glenn, B. B. 1965. Biological, physical and chemical study of Lake Apopka, 1962-1964. Florida State Board of Health. Jacksonville, FL. Hori, S. H., Kodama, T., and Tanahashi, K. 1979. Induction of vitellogenin synthesis in goldfish by massive doses of androgens. General and Comparative Endocrinology 37: 306-320. James, M. O. and Kleinow, K. M. 1994. Trophic Transfer of Chemicals in the Aquatic Environment. Pages 1-36 in D. C. Malins and G. K. Ostrander, editors. Aquatic Toxicology: Molecular, Biochemical, and Cellular Perspectives. Lewis Publishers, Boca Raton, FL. Janz, D. M. and Weber, L. P. 2000. Microscopic Functional Anatomy: Endocrine System. Pages 415-440 in The Laboratory Fish. Academic Press, San Diego, CA. Johnson, L. L., Casillas, E., Collier, T. K., McCain, B. B., and Varanasi, U., 1988. Contaminatn effects on ovarian development in English sole Parophrys vetulus from Puget Sound, Washington. Canadian Journal of Fisheries and Aquatic Sciences 45: 2133-2146. Kelce, W. R., Stone, C. R., Laws, S. C., Gray, L. E., Kemppainen, J. A., and Wilson, E. M. 1995. Persistent DDT metabolite p,p-DDE is a potent androgen receptor antagonist. Nature 375(6532): 581-585. Kullman, S. W., Hamm, J. T., and Hinton, D. E. 2000. Identification and characterization of a cDNA encoding cytochrome P450 3A from the fresh water teleost medaka (Oryzias latipes). Archives of Biochemistry and Biophysicsq 380: 29-38. Larkin, P., Sabo-Attwood, T., Kelso, J., and Denslow, N. D. 2002. Gene expression analysis of largemouth bass exposed to estradiol, nonylphenol, and p,p-DDE. Comparative Biochemistry and Physiology, Part B 133: 543-557. Leaos-Castaeda, O., Van Der Kraak, G., Lister, A., Sim-Alvarez, R., and Gold-Bouchot, G. 2002. o,p'-DDT induction of vitellogenesis and its inhibition by tamoxifen in Nile tilapia (Oreochromis niloticus). Marine Environmental Research 54: 703-707.

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100 Leavens, T. L., Sparrow, B. R., and Devito, M. J. 2002. Lack of antiandrogenic effects in adult male rats following acute exposure to 2,2-bis(4-chlorophenyl)-1,1-dicloroethylene (p,p'-DDE). Toxicology 174: 69-78. Macek, K. J. and Korn, S. 1970. Significance of the food chain in DDT accumulation in fish. Journal Fisheries Research Board of Canada 27: 1496-1498. Macek, K. J., Rodgers, C. R., Stalling, D. L., and Korn, S. 1970. The uptake, distribution and elimination of dietary 14 C-DDT and 14 C-Dieldrin in rainbow trout. Transactions of the American Fisheries Society 4: 689-695. Machala, M., Drbek, Neca, J., Kolrov, J., and Svobodov, Z. 1998. Biochemical markers for differentiation of exposures to noplanar polychlorinated biphenyls, organochlorine pesticides, or 2,3,7,8-tetrachlorodibenzo-p-dioxin in trout liver. Ecotoxicology and Environmental Safety 41: 107-111. Madsen, L. L., Korsgaard, B., and Bjerregaard, P. 2002. 4-tert-octylphenol and 17-estradiol applied by feeding to flounder Platichthys flesus: induction of vitellogenin and accumulation in tissues. Marine Environmental Research 54: 729-733. Marburger, Joy E. Johnson, William E., Douglas, David R., and Gross, Timothy S. 1999. Pesticide Contamination of fish and Sediments in the Emeralda Marsh Conservation Area: Relevance to Fisheries Establishment in Flooded Muck Farms. Technical Memorandum No. 31 to the St. Johns Water Management District. Marburger, Joy E., Johnson, William E., Gross, Timothy S., Douglas, David R., and Di, Jian., 2002. Residual organochlorine pesticides in soils and fish from wetland restoration areas in central Florida, USA. Wetlands 22:705-711. Matthews, J., Celius, T., and Halgren, R. Z. T. 2000. Differential estrogen receptor binding of estrogenic substances: a species comparison. The Journal of Steroid Biochemistry and Molecular Biology 74: 223-234. McMaster, M. E., Van Der Kraak, G. J., and Munkittrick, K. R. 1995. Exposure to bleached kraft pulp mill effluent reduces the steroid biosynthetic capacity of white sucker ovarian follicles. Comparative Biochemistry and Physiology, Part C 112: 169-178. Milligan, S. R., Khan, O., and Nash, M. 1998. Competitive binding of xenobiotic oestrogens to rat alpha-fetoprotein and to sex steroid binding proteins in human and rainbow trout (Oncorhynchus mykiss) plasma. General and Comparative Endocrinology 112: 89-95.

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101 Monosson, E., Hogson, F. G., Fleming, W. J., and Sullivan, C. V. 1996. Blood plasma levels of sex steroid hormones and vitellogenin in striped bass (Monrone saxatilis) exposed to 3,3',4,4'-Tetrachlorobiphenyl (TCB). Bulletin of Environmental Contamination and Toxicology 56: 782-787. Nelson, M. S. 2001. Thesis. The annual reproductive cycle of largemouth bass (Micropterus salmoides). University of Florida. Nimrod, A. C. and Benson, W. H. 1997. Xenobiotic interaction with and alteration of channel catfish estrogen receptor. Toxicology and Applied Pharmacology 147: 381-390. Okoumassoun, L.-E., Averill-Bates, D., Gagne, F., Marion, M., and Denizeau, F. 2002a. Assessing the estrogenic potential of organochlorine pesticides in primary cultures of male rainbow trout (Oncorhynchus mykiss) hepatocytes using vitellogenin as a biomarker. Toxicology 178: 193-207. Okoumassoun, L.-E., Brochu, C., Deblois, C., Akponan, S., Marion, M., Averill-Bates, D., and Denizeau, F. 2002b. Vitellogenin in tilapia male fishes exposed to organochlorine pesticides in Oueme River in Republic of Benin. The Science of The Total Environment 299: 163-172. Pelissero, C., Flouriot, G., Foucher, J. L., Bennetau, B., Dunogues, J., Le Gac, F., and Sumpter, J. P. 1993. Vitellogenin synthesis in cultured hepatocytes; an in vitro test for the estrogenic potency of chemicals. The Journal of Steroid Biochemistry and Molecular Biology 44: 263-272. Pollman, Curtis D., Graetz, D. A., Ramsey, F. V., Reddy, K. R., and Sullivan, T. J. 1988. Feasibility of sediment removal and reuse for the restoration of Lake Apopka: Final Report: Prepared for St. Johns River Water Management District. Rosenblum, P. M., Brandt, T. M., Mayes, K. B., and Hutson, P. 1994. Annual cycles of growth and reproduction in hatchery-reared Florida largemouth bass, Micropterus salmoides floridanus, raised on forage or pellet diets. Journal of Fish Biology 44: 1045-1059. Ruus, A., Sanvik, M., Ugland, K. I., and Skaare, J. U. 2002. Factors influencing activities of biotransformation enzymes, concentrations and compositional patterns of organochlorine contaminants in members of a marine food web. Aquatic Toxicology 61: 73-87. Ryffel, G. U. 1978. Synthesis of vitellogenin, an attractive model for investigating hormone-induced gene activation. Molecular and Cellular Endocrinology 12: 237-246.

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102 Sanderson, J. T., Boerma, J., Lansbergen, G. W. A., and van den Berg, M. 2002. Induction and inhibition of aromatase (CYP19) activity by various classes of pesticides in H295R human adrenocortical carcinoma cells. Toxicology and Applied Pharmacology 182: 44-54. Santerre, C. R., Blazer, V. S., Khanna, N., Reinert, R. E., and Barrows, F. T. 1997. Absorption of dietary dieldrin by striped bass. Bulletin of Environmental Contamination and Toxicology 58: 334-340. Seplveda, M. S. 2000. Dissertation. Effects of paper mill effluent on the health and reproductive status of largemouth bass (Micropterus salmoides): field and laboratory studies. University of Florida. Seplveda, M. S., Johnson, W. E., Higman, J. C., Denslow, N. D., Schoeb, T. R., and Gross, T. S. 2002. An evaluation of biomarkers of reproductive function and potential contaminant effects in Florida largemouth bass (Micropterus salmoides floridanus) sampled from the St. Johns River. The Science of the Total Environment 289: 133-144. Seplveda, M. S., Wiebe, J. J., Honeyfield, D. C., Hinterkopf, J. P., Johnson, W. E., Gross, T. S. 2003. Organochlorine pesticides and thiamine in eggs of largemouth bass and the American alligator, and their relationship with early life-stage mortality. Submitted to Marine Environmental Research. Shelby, M. D., Newbold, R. R., Tully, D. B., Chae, K., and Davis, V. L. 1996. Assessing environmental chemicals for estrogenicity using a combination of in vitro and in vivo assays. Environmental Health Perspectives 104: 1296-1300. Smeets. Jean M. W., van Holsteign, I., Giesy, J. P., Seinen, W., and van den Berg, M. 1999. Estrogenic potencies of several environmental pollutants, as determined by vitellogenin induction in a carp hepatocyte assay. Toxicological Sciences 50: 206-213. Snowberger, E. A. and Stegeman, J. J. 1987. Patterns and regulation of estradiol metabolism by hepatic microsomes from two species of marine teleosts. General and Comparative Endocrinology 66: 256-265. Stegman, J. J. and Hahn, M. E. 1994. Biochemistry and Molecular Biology of Monooxygenases: Current Perspectives on Forms, Functions, and Regulation ofCytochrom P450 in Aquatic Species. Pages 87-206 in D. C. Malins and G. K. Ostrander, editors. Aquatic Toxicology: Molecular, Biochemical, and Cellular Perspectives. Lewis Publishers, Boca Raton, FL. Sumpter, J. P. and Jobling, S. 1995. Vitellogenesis as a biomarker for estrogenic contamination of the aquatic environment. Environmental Health Perspectives, Supplement 103: 173-178.

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103 Tollefsen, K. E., Mathisen, R., Stenersen, J. 2002a. Estrogen mimics bind with similar affinity and specificity to the hepatic estrogen receptor in Atlantic salmon (Salmo salar) and rainbow trout (Oncorhynchus mykiss). General and Comparative Endocrinology 126: 14-22. Tollefsen, K. E. 2002b. Interaction of estrogen mimics, singly and in combination, with plasma sex steroid-binding proteins in rainbow trout (Oncorhynchus mykiss). Aquatic Toxicology 56: 215-225. U.S. Environmental Protection Agency, 1990. Method 8270B, Rev. 2. of Test methods for evaluating solid waste (SW-846), 1983. Office of Solid Waste, Washington, D.C. Van der Kraak, G., Chang, J. P., and Janz, D. M. 1998. Reproduction. Pages 465-488 in D. H. Evans, editor, The Physiology of Fishes. CRC Press, Boca Raton, FL. Van der Kraak, G. J., Munkittrick, K. R., McMaster, M. E., Portt, C. B., and Chang, J. P. 1992. Exposure to bleached kraft pulp mill effluent disrupts the pituitary-gonadal axis of white sucker at multiple sites. Toxicology and Applied Pharmacology 115: 224-233. Wade, M. G., Foster, W. G., Younglai, E. V., McMahon, A., Leingartner, K., Yagminas, A., Blakey, D., Fournier, M., Desaulniers, D., and Hughes, C. L. 2002. Effects of subchronic exposure to a complex mixture of persistent contaminants in male rats: systemic, immune, and reproductive effects. Toxicological Sciences 67: 131-143. Wahli, W., Dawid, I. B., Ryffel, G. U., and Weber, R. 1981. Vitellogenesis and the vitellogenin gene family. Science 212: 298-304. Wallace, R. A. 1970. Studies on amphibian yolk IX. Xenopus vitellogenin. Biochimica et Biophysica Acta 215: 176-183. Wallace, R. A. and Selman, K. 1981. Cellular and ynammic aspects of oocyte growth in teleosts. American Zoologist 21: 325-343. Warlen, S. M., Wolfe, D. A., Lewis, C. W., and Colby, D. R. 1977. Accumulation and retention of dietary 14 C-DDT by Atlantic menhaden. Transactions of the American Fisheries Society 106. Wells, K. and Van Der Kraak, G. 2000. Differential binding of endogenous steroids and chemicals to androgen receptors in rainbow trout and goldfish. Environmental Toxicology and Chemistry 19: 2059-2065. Willingham, E. and Crews, D. 1999. Sex reversal effects of environmentally relevant xenobiotic concentrations on the red-eared slider turtle, a species with temperature-dependent sex determination. General and Comparative Endocrinology 113: 429-435.

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104 Woodwell, G. M., Wurster, C. F., and Isaacson, P. A. 1967. DDT residues in an east coast estuary: a case of biological concentration of a persistent insecticide. Science 156: 821-824. You, L., Chan Sanny K., Bruce, J. M., Archibeque-Engle, S., Casanova, M., Corton, J. C., and Heck, H. d. 1999. Modulation of testosterone-metabolizing hepatic cytochrome P-450 enzymes in developing Sprague-Dawley rats following in utero exposure to p,p'-DDE. Toxicology and Applied Pharmacology 158: 197-205. You, L., Sar, M., Bartolucci, E., Ploch, S., and Whitt, M. 2001. Induction of hepatic aromatase by p,p-DDE in adult male rats. Molecular and Cellular Endocrinology 178: 207-214. Zapata-Prez, O., Sim-Alvarez, R., Norea-Barroso, E., Gemes, J., Gold-Bouchot, G., Ortega, A., and Albores-Medina, A. 2000. Toxicity of sediments from Baha de Chetumal, Mexico, as assessed by hepatic EROD induction and histology in nile tilapia Oreochromis niloticus. Marine Environmental Research 50: 385-391.

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BIOGRAPHICAL SKETCH Jennifer Keene Muller was born August 17, 1977 and grew up in Atlanta, GA and Winter Park, FL. She attended Winter Park High School and graduated in May 1995. She then attended Valencia Community College through the summer of 1997 and received her Associate in Arts. In fall of 1997 she transferred to Eckerd College in St. Petersburg, FL where she completed a Bachelor of Science degree with honors, majoring in marine biology, in January 2000. After undergraduate study, she was a biological technician under Dr. Nannette Nascone-Yoder at Eckerd College. Her work there focused on examining differential left-right gene expression in the development of the heart of the African claw-toed frog, Xenopus laevis. In the summer of 2001, she began her work in the Ecotoxicology laboratory of Dr. Timothy S. Gross at the United States Geological Survey; and enrolled for her graduate program in the Fall, with Dr. Gross as her advisor. At the time of completion of her thesis, Jennifer had procured a position as a staff toxicologist at Applied Pharmacology and Toxicology, Inc. in Alachua, FL. 105


Permanent Link: http://ufdc.ufl.edu/UFE0001107/00001

Material Information

Title: An Evaluation of dosing methods and effects of p,p'-DDE and dieldrin in Florida largemouth bass (Micropterus salmiodes floridanus)
Physical Description: Mixed Material
Language: English
Creator: Muller, Jennifer Keene ( Dissertant )
Gross, Timothy S. ( Thesis advisor )
Sepùlveda, Maria ( Reviewer )
Barber, David ( Reviewer )
Publisher: University of Florida
Place of Publication: Gainesville, Fla.
Publication Date: 2003
Copyright Date: 2003

Subjects

Subjects / Keywords: Physiological Sciences thesis, M.S.
Dissertations, Academic -- UF -- Physiological Sciences
Spatial Coverage: United States--Florida

Notes

Abstract: Previous work indicates that persistent organochlorine pesticides (OCPs) have a wide range of effects on steroidogenesis and reproduction in fish. The present study investigated consistency of dose administration, dose accumulation and potential reproductive effects of two OCPs, p,p'-DDE and dieldrin, on Florida largemouth bass (Micropterus salmoides floridanus). The first study used 60-day slow release pellets inserted into the intraperitoneal cavity. Twenty-five fish comprised each of eight treatment groups: sham; placebo (matrix only pellet); 3 p,p'-DDE; and 3 dieldrin doses. Exposures were initiated before onset of the reproductive season (January) so that the full dose would be released prior to spawning (March). Weight, length, condition factor, gonadosomatic (GSI) and hepatosomatic (HSI) indices, and circulating hormones were measured. Preliminary contaminant analysis of gonadal tissue indicated that pellets did not release a consistent dose in each treatment. Statistical analyses, therefore, were more appropriately based on actual pesticide concentrations in blood plasma. No change in GSI, HSI, or any health parameter was observed in fish treated with either chemical indicating a lack of acute toxicity. Results showed a decrease in estradiol production in female bass with high blood plasma concentrations of both pesticides (p,p'-DDE 200 - 1,000 ppb; dieldrin 100 - 500 ppb). Also, male bass treated with dieldrin, but not p,p'-DDE, demonstrated a dose-dependent decrease in circulating 11-ketotestosterone concentration at a blood plasma concentration range of 100 - 600 ppb. Altered steroid synthesis may result in asynchronous spawning and reduced reproductive success seen in areas of the Ocklawaha River Basin, Florida contaminated with OCPs. The second study evaluated an oral exposure method for largemouth bass to OCPs. Feed pellets were dosed with 5 ppm p,p'-DDE or 1 ppm dieldrin. Groups of 14 fish were fed sinking or floating diets to determine which feed produced a more consistent whole-body dose. Variability within replicates, total accumulated dose, and day 50 hormone concentrations were measured. Floating pellets delivered a more consistent dose. A steady state whole body concentration was achieved at approximately Day 30 and maintained through the end of the study at Day 50. Treated bass had depressed sex steroid concentrations at Day 50 even though whole body concentrations were lower than those found in natural environments in the Ocklawaha River Basin with high OCP levels. The research presented provides evidence of contaminated feed as a more appropriate exposure method for consistent OCP dosing in largemouth bass. Also, p,p'-DDE and dieldrin in single chemical exposures can modulate endocrine function in largemouth bass. The two OCPs studied may contribute to the effects of a more complex mixture of chemicals found in the environment.
Subject: bass, DDE, dieldrin, dosing, largemouth, pesticide
General Note: Title from title page of source document.
General Note: Includes vita.
Thesis: Thesis (M.S.)--University of Florida, 2003.
Bibliography: Includes bibliographical references.
General Note: Text (Electronic thesis) in PDF format.

Record Information

Source Institution: University of Florida
Holding Location: University of Florida
Rights Management: All rights reserved by the source institution and holding location.
System ID: UFE0001107:00001

Permanent Link: http://ufdc.ufl.edu/UFE0001107/00001

Material Information

Title: An Evaluation of dosing methods and effects of p,p'-DDE and dieldrin in Florida largemouth bass (Micropterus salmiodes floridanus)
Physical Description: Mixed Material
Language: English
Creator: Muller, Jennifer Keene ( Dissertant )
Gross, Timothy S. ( Thesis advisor )
Sepùlveda, Maria ( Reviewer )
Barber, David ( Reviewer )
Publisher: University of Florida
Place of Publication: Gainesville, Fla.
Publication Date: 2003
Copyright Date: 2003

Subjects

Subjects / Keywords: Physiological Sciences thesis, M.S.
Dissertations, Academic -- UF -- Physiological Sciences
Spatial Coverage: United States--Florida

Notes

Abstract: Previous work indicates that persistent organochlorine pesticides (OCPs) have a wide range of effects on steroidogenesis and reproduction in fish. The present study investigated consistency of dose administration, dose accumulation and potential reproductive effects of two OCPs, p,p'-DDE and dieldrin, on Florida largemouth bass (Micropterus salmoides floridanus). The first study used 60-day slow release pellets inserted into the intraperitoneal cavity. Twenty-five fish comprised each of eight treatment groups: sham; placebo (matrix only pellet); 3 p,p'-DDE; and 3 dieldrin doses. Exposures were initiated before onset of the reproductive season (January) so that the full dose would be released prior to spawning (March). Weight, length, condition factor, gonadosomatic (GSI) and hepatosomatic (HSI) indices, and circulating hormones were measured. Preliminary contaminant analysis of gonadal tissue indicated that pellets did not release a consistent dose in each treatment. Statistical analyses, therefore, were more appropriately based on actual pesticide concentrations in blood plasma. No change in GSI, HSI, or any health parameter was observed in fish treated with either chemical indicating a lack of acute toxicity. Results showed a decrease in estradiol production in female bass with high blood plasma concentrations of both pesticides (p,p'-DDE 200 - 1,000 ppb; dieldrin 100 - 500 ppb). Also, male bass treated with dieldrin, but not p,p'-DDE, demonstrated a dose-dependent decrease in circulating 11-ketotestosterone concentration at a blood plasma concentration range of 100 - 600 ppb. Altered steroid synthesis may result in asynchronous spawning and reduced reproductive success seen in areas of the Ocklawaha River Basin, Florida contaminated with OCPs. The second study evaluated an oral exposure method for largemouth bass to OCPs. Feed pellets were dosed with 5 ppm p,p'-DDE or 1 ppm dieldrin. Groups of 14 fish were fed sinking or floating diets to determine which feed produced a more consistent whole-body dose. Variability within replicates, total accumulated dose, and day 50 hormone concentrations were measured. Floating pellets delivered a more consistent dose. A steady state whole body concentration was achieved at approximately Day 30 and maintained through the end of the study at Day 50. Treated bass had depressed sex steroid concentrations at Day 50 even though whole body concentrations were lower than those found in natural environments in the Ocklawaha River Basin with high OCP levels. The research presented provides evidence of contaminated feed as a more appropriate exposure method for consistent OCP dosing in largemouth bass. Also, p,p'-DDE and dieldrin in single chemical exposures can modulate endocrine function in largemouth bass. The two OCPs studied may contribute to the effects of a more complex mixture of chemicals found in the environment.
Subject: bass, DDE, dieldrin, dosing, largemouth, pesticide
General Note: Title from title page of source document.
General Note: Includes vita.
Thesis: Thesis (M.S.)--University of Florida, 2003.
Bibliography: Includes bibliographical references.
General Note: Text (Electronic thesis) in PDF format.

Record Information

Source Institution: University of Florida
Holding Location: University of Florida
Rights Management: All rights reserved by the source institution and holding location.
System ID: UFE0001107:00001


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AN EVALUATION OF DOSING METHODS AND EFFECTS OF p,p '-DDE AND
DIELDRIN IN FLORIDA LARGEMOUTH BASS
(Micropterus salmoides floridanus)















By

JENNIFER KEENE MULLER


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


2003

































Copyright 2003

by

Jennifer K. Muller















ACKNOWLEDGMENTS

I would like to extend my gratitude to Dr. Timothy S. Gross for providing

resources, funding, staff, and guidance over the course of my studies. This research

could not have been accomplished without the aid of the entire Ecotoxicology Lab at

USGS-BRD-CARS (Gainesville, FL). Special thanks go to Carla Weiser, Nikki

Kernaghan, Shane Ruessler, Jon Weibe, Janet Buckland, Travis Smith, Jessica Grosso

and my fellow graduate students Jessica Noggle, Eileen Monck, Kevin Johnson, Heath

Rauschenberger, and Sekeenia Haynes.

I would like to thank my other committee members, Dr. Maria Sepulveda and Dr.

David Barber, for their invaluable assistance in revising this thesis. Their comments

greatly improved my writing. Also, I would like to acknowledge Dr. Christopher J.

Borgert for study design and editorial assistance.

This research was made possible through a grant to Dr. Gross and Dr. Borgert from

the American Chemistry Council.

Lastly, I would like to thank my husband, Josh, and my family for all of the love,

support, and most of all, patience they have given me over the past two years. My life

and my ability as a scientist benefits from their presence.
















TABLE OF CONTENTS

page

A C K N O W L E D G M E N T S ................................................................................................. iii

LIST OF TABLES ....................................................... ............ .............. .. vii

LIST OF FIGURES .............. ................................. ............. ........... viii

ABSTRACT ........ .............. ............. .. ...... .......... .......... xii

CHAPTER

1 IN TR OD U CTION ............................................... .. ......................... ..

B background ........................ ........... ...... ...................... ...............
U pper O cklaw aha River B asin ...................................... ............. ...................
Chemical Contamination in the Upper Ocklawaha Basin............... ..................
Selection of A nim al M odel............................................................................ ...... 3
Largem outh Bass Reproductive Cycle ........................................ ...... ............... 4
OCP Contaminant Effects on Fish Reproduction.............................. .....................6
R research Significance......... .............................................................. .......... .......

2 EXPOSURE OF LARGEMOUTH BASS TO p,p '-DDE AND DIELDRIN BY
SLOW-RELEASE MATRIX PELLET ............................................ .................11

In tro d u ctio n .............................................................................................. 1 1
S p ecific A im s ...............................................................12
N u ll H y p o th e sis ............................................................................................. 12
A alternative H ypothesis ................................................................................. 12
M materials and M methods ....................................................................... .................. 12
E xperim ental A nim als .................................................................... ............... 12
E x p erim ental D esig n ............................................. ........................................ 13
Chemicals and Dosing ........................................... ............... .............13
B lood and Tissue C collection ........................................ .......... ............... 14
Gonad Histology ................................................................. ...... ........ 15
Analysis of Circulating Sex Steroid Hormones....................................... 15
Analysis of Largemouth Bass Tissues for OCPs ................................................16
Analysis of Largemouth Bass Blood Plasma for OCPs ......................... .....17
Statistical A naly sis .......................................... ............ ...... ........19
R results ................. ... ......... ............... ...................................20









Largem south Bass ....................... .............. ......... ......................... 20
H health Param eters: K and H SI ........................................ ........................ 20
G onad H istology.................. .............. ...... ................. ............ .. .............. 20
Reproductive Parameters: GSI and Circulating Sex Steroids ............................20
G S I ............................................................................................................... 2 0
17P -E strad io l ............................................................2 1
11-Ketotestosterone ......... ...... ....................... .. ......................21
In Vivo Treatment Dosing Consistency ......... ......... .. ........ ............ 22
Health Parameters: K and H SI (Regression).....................................................24
Reproductive Parameters: GSI and Circulating Sex Steroids (Regression)........25
G S I ............................................................................................................... 2 5
1 7P -E stra d io l .......................................................................................... 2 5
11-K etotestosterone............. .. .... .... ...... ............ ................ ......26
Discussion ..................................................26

3 ACCUMULATION OF DIETARY p,p '-DDE AND DIELDRIN BY
LARGEMOUTH BASS: A PILOT STUDY..........................................................63

In tro d u ctio n .......................................................................................6 3
S p ecific A im s ...............................................................64
N u ll H y p o th e sis ............................................................................................. 6 4
Alternative Hypothesis .........................................................65
M materials and M methods .................................. ..........................................65
Experim mental A nim als ............................................................ ............... 65
C hem icals and F eed ........................................... ................... ............... 66
Feeding R ate.................. .................. ................. ......... ......... ...67
D ay 30 and D ay 50 Endpoints...................................... ...................... ........... 67
Analysis of Largemouth Bass Tissue for OCPs ...............................................68
Analysis of Circulating Sex Steroid Hormones.......................................68
Statistical A analysis .......................... ............ ...........................69
R e su lts .................. ........................................................... ................6 9
L argem south B ass .................................................... .. ........................ .. 69
F feeding R ate ................................................................... ............ 70
30-D ay Exposure ..................................................... ... .. ............ 70
p,p '-D D E accum ulation ........................................ ......................... 70
D ieldrin accum ulation ...................................................... .................. 71
G SI and sex steroids ........................................................ .. .......... .. ..72
50-D ay E exposure ................................................ ..... ... .. ............ 72
p,p '-D D E accum ulation ........................................ ......................... 73
D ieldrin accum ulation ............................................ ......................... 74
G SI and sex steroids ........................................................ .. .......... .. ..74
D iscu ssio n ...................................... ................................................. 7 5

4 GENERAL CONCLUSIONS / FUTURE RESEARCH ................. .... ............93

G en eral C o n clu sio n s .............................................. .......................... ..................... 9 3
Future R research ............................................. 95


v









L IST O F R E F E R E N C E S ...................................... .................................... ....................96

BIOGRAPHICAL SKETCH ............................................................. ..................105















LIST OF TABLES


Table page

2-1 GC-MS results of tissue samples for bioaccumulation measurements at Day 30
of exposure. Blue and green numbers highlight fish treated with p,p '-DDE and
dieldrin, respective ely ..................................................................... ................ .. 46

3-1 Ten-day increment totals of feed eaten per tank. The total feed eaten by each
tank was totaled at Day 30 and used to estimate the average amount of pesticide
ingested by each fish. .......................................... ........................ 82

3-2 Ten-day increment totals of feed eaten per tank (floating only after Day 30).
The total feed eaten by each tank was totaled at Day 50 and used to estimate the
average amount of pesticide ingested by each fish. ............................................83















LIST OF FIGURES


Figure p

2-1 Female body condition over time. Box plots of weight (a), length (b), and
condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 63). Box
plot contains the 25th to the 75th quartile, line in box indicates the median,
whiskers extend to the minimum and maximum value ........................................36

2-2 Male body condition over time. Box plots of weight (a), length (b), and
condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 93). Box
plot contains the 25th to the 75th quartile, line in box indicates the median,
whiskers extend to the minimum and maximum value ........................................37

2-3 Day 60 HSI. There was no significant difference in HSI among treatment
group for females (a). Males in the Dieldrin 1.0 mg treatment group showed a
significant increase in HSI over Sham and DDE 2.5 mg treatments (b). Sample
size per treatment ranged from 7 to 11 for females and 7 to 16 for males.
Treatments with the same lower case letter were not significantly different .........38

2-4 Histological section of stage 4 vitellogenic female gonad viewed at 40X. CA =
cortical alveoli, GV = germinal vesicle, YV = yolk vesicle. ................................39

2-5 Day 60 GSI. There was no significant difference in GSI among treatment
group for females (a) or males (b). Sample size per treatment ranged from 7 to
11 for females and 7 to 16 for males. Treatments with the same lower case letter
w ere not significantly different. ........................................ ......................... 40

2-6 Female circulating estradiol at Days 0 (a), 30 (b), and 60 (c). Placebo and DDE
2.5 treatments showed consistently higher plasma E2 concentrations than all
other treatments. Sample size per treatment ranged from 7 to 11. Treatments
with the same lower case letter were not significantly different..............................41

2-7 Male circulating estradiol at Days 0 (a), 30 (b), and 60 (c). DDE 5.0 mg
treatment showed a consistently higher mean than all other treatments. Sample
size per treatment ranged from 7 to 16. Treatments with the same lower case
letter were not significantly different. ............................... .. ....................... 42

2-8 Female circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Treatment groups
demonstrated high variability compared to controls. By Day 60, no difference
existed between treatment groups and controls. Sample size per treatment









ranged from 7 to 11. Treatments with the same lower case letter were not
significantly different. ........................... ..................................... .... ...... ...... 43

2-9 Male circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Placebo and DDE 2.5
mg treatment groups demonstrated consistently higher 11-KT. At Day 30 and 60
all dieldrin treatments and higher DDE treatments demonstrated decreased
11-KT. Sample size per treatment ranged from 7 to 16. Treatments with the
same lower case letter were not significantly different................ ..................44

2-10 Day 60 mean female gonadal dose. p,p '-DDE (a) and dieldrin (b)
concentration SD for each treatment group (n = 6) compared to the target
dose for each treatment. Also included is the mean gonadal concentration of
either OCP found in the Eustis property of EMCA. ........................... ...............45

2-11 Regression analysis of gonad (a, d), liver (b, e) and muscle (c, f) DDE
concentrations against blood DDE concentrations in females (a c; n = 9) and
males (d f; n = 6). Significant linear relationships exist between blood plasma
and tissue DDE concentrations, except in male muscle .........................................47

2-12 Regression analysis of gonad (a, d), liver (b, e), and muscle (c, f) dieldrin
concentrations against blood dieldrin concentrations in females (a c; n = 4) and
males (d f; n = 6). Significant linear relationships exist between blood plasma
and tissue dieldrin concentrations, except in female muscle. ................................48

2-13 Linear regression of ELISA DDE (a) or dieldrin (b) results of pooled blood
samples against GC-MS results (n = 12). A highly significant and linear
relationship exists between the two methods. .................................. .................49

2-14 Linear regression of HSI against blood plasma DDE concentrations. No
significant correlation was found for females (a, n = 37) or males (b, n = 53)........50

2-15 Linear regression of HSI against blood plasma dieldrin concentration. No
significant correlation was found for females (a, n = 34) or males (b, n = 51)........51

2-16 Linear regression of GSI against blood plasma DDE concentration. No
significant correlation was found for females (a, n = 37) or males (b, n = 53)........52

2-17 Linear regression of GSI against blood plasma dieldrin concentration. No
significant correlation was found for females (a, n = 34) or males (b, n = 51)........53

2-18 Day 0 plasma hormone concentrations. Mean SD (a) female, n = 59 (b) male,
n = 8 8 ........................................................................... 5 4

2-19 Day 30 Hormones Femalep,p '-DDE treated (n = 30). Linear regression
against DDE dose (a) circulating hormones (b) percent change in hormone
concentration from Day 0 to Day 30. Significant negative relationships exist
between DDE dose and circulating E2 and percent change in E2...........................55









2-20 Day 30 Hormones Malep,p '-DDE treated (n = 47). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day 0
to day 30 against DDE dose. No significant relationships were found .................56

2-21 Day 60 Hormones Femalep,p '-DDE treated (n = 35). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day 30
to Day 60 againstp,p '-DDE dose and (c) ciculating E2 to GSI. A significant
negative relationship was found between DDE dose and circulating E2. ...............57

2-22 Day 60 Hormones Malep,p '-DDE treated (n = 51). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day 30
to Day 60 against dieldrin dose and (c) circulating 11-KT against GSI. No
significant relationships were found. ............................................ ............... 58

2-23 Day 30 Hormones Female dieldrin treated (n = 34). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day 0
to Day 30 against dieldrin dose. No significant relationships were found..............59

2-24 Day 30 Hormones Male dieldrin treated (n = 42). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day 0
to Day 30 against dieldrin dose. A significant negative relationship exists
between dieldrin dose and percent change in 11-KT from Day 0 to Day 30..........60

2-25 Day 60 Hormones Female dieldrin treated (n= 34). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day 30
to Day 60 against blood plasma dieldrin and (c) circulating E2 against GSI. A
significant negative relationship between blood plasma dieldrin and circulating
E 2 w as fou n d ........................................................................................ 6 1

2-26 Day 60 Hormones Male dieldrin treated (n = 51). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day 0 to
Day 30 against dieldrin dose and (c) circulating 11-KT against GSI. Significant
negative relationships between circulating 11-KT and dieldrin dose, as well as,
between percent change in E2 from Day 30 to Day 60 were found .......................62

3-1 Tank setup for preliminary feeding study. Each set of four tanks had a divided
well water and air supply line and separate drainage lines. .................................81

3-2 Day 30 whole-body concentration SD ofp,p '-DDE (A) and dieldrin (B) in
largemouth bass fed a contaminated diet (n = 3 per treatment). All dosed fish
had significantly higher whole-body concentrations of their respective chemical
than the controls. One replicate fed sinkingp,p '-DDE contaminated feed was
significantly higher than all otherp,p '-DDE fed fish. There were no significant
differences among replicates fed dieldrin contaminated diets. Bars with
different letter designations indicate a significant difference. ...............................84

3-3 Day 30, percentage of total dose accumulated SD within each replicate (n = 3
per replicate). There were no significant differences in the amount accumulated









between replicates of either feed type for dieldrin. There was a significant
difference between one DDE sinking diet replicate and one DDE floating diet
replicate. Bars with different letter designations indicate a significant difference. 85

3-4 Day 50 whole-body concentration SDofp,p '-DDE (A) or dieldrin (B) in
largemouth bass fed a contaminated diet (n = 3 per replicate). All dosed fish
had significantly higher whole-body concentrations of their respective chemical
than the controls. There was no significant difference amongp,p '-DDE
replicates. Whole-body concentration was significantly different between
dieldrin replicates. Bars with different letter designations indicate a significant
difference ................. ....... ............................................................ 86

3-5 Day 50, percentage of total dose accumulated SD within each replicate (n = 3
per replicate). There were no significant differences in the amount accumulated
between replicates of either for eitherp,p '-DDE or dieldrin. Bars with different
letter designations indicate a significant difference. .............................................. 87

3-6 Gonadal concentrations of OCPs. Replicates ofp,p '-DDE were not significantly
different, however, there was a significant difference in gonadal concentrations
of dieldrin between replicates most likely due to the difference in total g of
dieldrin fed. .......................................... ............................ 88

3-7 Day 50, Female circulating hormones SD (n ranged from 3 to 5). Significant
differences in E2 were found between both replicates ofp,p '-DDE and dieldrin
fed fish and controls. No significant differences were found in 11-KT levels in
females. Bars with different letter designations indicate a significant difference. .89

3-8 Day 50, male circulating hormones SD (n ranged from 6 to 8). Significant
differences were found in both E2 and 11-KT levels in all treatment groups and
replicates as compared to controls. Bars with different letter designations
indicate a significant difference. ........................................ .......................... 90

3-9 Whole carcass concentration of floating style feed replicates over time, A -p,p '-
D D E, B D ieldrin. .......................................... ...... .. .... ............91

3-10 Female and male GSI SD for Days 30 (n = 16 and 20, respectively) and 50
(n = 21 and 44, respectively). There was no significant difference in GSI between
Days 30 and 50 when GSI should have increased dramatically. Bars with
different designation indicate significant difference........ ................ ............... 92














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

AN EVALUATION OF DOSING METHODS AND EFFECTS OF p,p '-DDE AND
DIELDRIN IN LARGEMOUTH BASS (Micropterus salmoidesfloridanus)

By

Jennifer Keene Muller

August 2003

Chair: Timothy S. Gross
Major Department: Physiological Sciences

Previous work indicates that persistent organochlorine pesticides (OCPs) have a

wide range of effects on steroidogenesis and reproduction in fish. The present study

investigated consistency of dose administration, dose accumulation and potential

reproductive effects of two OCPs, p,p '-DDE and dieldrin, on Florida largemouth bass

(Micropterus salmoidesfloridanus).

The first study used 60-day slow release pellets inserted into the intraperitoneal

cavity. Twenty-five fish comprised each of eight treatment groups: sham; placebo

(matrix only pellet); 3 p,p '-DDE; and 3 dieldrin doses. Exposures were initiated before

onset of the reproductive season (January) so that the full dose would be released prior to

spawning (March). Weight, length, condition factor, gonadosomatic (GSI) and

hepatosomatic (HSI) indices, and circulating hormones were measured. Preliminary

contaminant analysis of gonadal tissue indicated that pellets did not release a consistent

dose in each treatment. Statistical analyses, therefore, were more appropriately based on









actual pesticide concentrations in blood plasma. No change in GSI, HSI, or any health

parameter was observed in fish treated with either chemical indicating a lack of acute

toxicity. Results showed a decrease in estradiol production in female bass with high

blood plasma concentrations of both pesticides (p,p '-DDE 200 1,000 ppb; dieldrin 100

- 500 ppb). Also, male bass treated with dieldrin, but notp,p '-DDE, demonstrated a

dose-dependent decrease in circulating 11-ketotestosterone concentration at a blood

plasma concentration range of 100 600 ppb. Altered steroid synthesis may result in

asynchronous spawning and reduced reproductive success seen in areas of the Ocklawaha

River Basin, Florida contaminated with OCPs.

The second study evaluated an oral exposure method for largemouth bass to OCPs.

Feed pellets were dosed with 5 ppmp,p '-DDE or 1 ppm dieldrin. Groups of 14 fish were

fed sinking or floating diets to determine which feed produced a more consistent whole-

body dose. Variability within replicates, total accumulated dose, and day 50 hormone

concentrations were measured. Floating pellets delivered a more consistent dose. A

steady state whole body concentration was achieved at approximately Day 30 and

maintained through the end of the study at Day 50. Treated bass had depressed sex

steroid concentrations at Day 50 even though whole body concentrations were lower than

those found in natural environments in the Ocklawaha River Basin with high OCP levels.

The research presented provides evidence of contaminated feed as a more

appropriate exposure method for consistent OCP dosing in largemouth bass. Also, p,p '-

DDE and dieldrin in single chemical exposures can modulate endocrine function in

largemouth bass. The two OCPs studied may contribute to the effects of a more complex

mixture of chemicals found in the environment.














CHAPTER 1
INTRODUCTION

Background

Upper Ocklawaha River Basin

The Upper Ocklawaha River basin in Central Florida has received much attention

in the last 20 years because of declining habitats and wildlife populations. The

historically hypereutrophic and contaminated Lake Apopka (located 25 km northwest of

Orlando (Pollman et al., 1988)) serves as the river's headwaters. From Lake Apopka,

water flows north through a chain of lakes, including Lake Beauclair, Lake Dora, Lake

Harris, Lake Eustis and Lake Griffin, before channeling into the Ocklawaha River. Much

of the marsh and wetland area surrounding this chain of lakes was diked and drained for

agricultural use or "muck farming" beginning in the 1940's and continuing through the

1980's (Benton et al., 1991). The removal of thousands of hectares of shallow lake

bottom resulted in the loss of spawning habitat for Florida largemouth bass (Micropterus

salmoidesfloridanus) and other sportfish, resulting in a decline of a once-thriving

sportfishery prior to the 1950's. Huffstutler et al. (1965) determined that the population

of largemouth bass in Lake Apopka was primarily adults and reproduction was not

sufficient to maintain a thriving population. To add to the loss of habitat, the

contamination of Lake Apopka resulted from inadequate sewage treatment discharge

from the city of Winter Garden beginning in 1922, a major pesticide spill by a nearby

chemical manufacturer in 1980, and ongoing agricultural sources that release nutrient and









pesticide rich irrigation water into many lakes in the Ocklawaha chain (Pollman et al.,

1988).

In an attempt to restore the Upper Ocklawaha River basin to a semi-natural

habitat, the state of Florida authorized the St. John's River Water Management District

(SJRWMD) to acquire muck farmland to flood and restore the marsh/wetland habitat. By

1993, the SJRWMD had acquired 2,630 hectares of muck farms adjacent to Lake Griffin

(the Emeralda Marsh) and reflooded portions during 1992 to 1995 (Marburger et al.,

1999). The objectives of this management practice were to reduce nutrient loading and

chemical runoff of applied pesticides and fertilizers and to return the marsh to its natural

habitat. However, flooding the land led to leaching of pollutants from the soil and their

subsequent biomagnification in the food chain.

Sport and forage fish were stocked in many flooded areas in 1992, 1993, and

1994, but with limited success (Benton and Douglas, 1996). Contributing factors

included low dissolved oxygen, low water levels, and overabundant vegetation.

Although the surviving largemouth bass had a high growth rate, reproductive success was

poor (assessed by the number of young of year counted); therefore, a viable fishery could

not be reestablished (Benton and Douglas, 1996).

Chemical Contamination in the Upper Ocklawaha Basin

Recent studies have begun to determine the extent of contamination in Lakes

Apopka and Griffin and in an area of reclaimed muck farms on the north shore of Lake

Griffin known as the Emeralda Marsh Conservation Area (EMCA). Lake Apopka

exhibits elevated levels of organochlorine pesticides (OCPs) in soil, fish (ATRA, 1997;

Marburger et al., 1999), and alligators (Heinz et al., 1991; Guillette et al., 1999a and

1999b). In 1997, a risk assessment was conducted for a muck farm on the north shore of









Lake Apopka that showed soil concentrations ranging from 136 ppb for endrin (760 ppb

for dieldrin) to 20,000 ppb for DDT (6,000 ppb forp,p '-DDE) and 36,000 ppb for

toxaphene (ATRA, 1997). A study at the EMCA showed soil concentrations of

prominent OCPs (p,p '-DDE, dieldrin, and toxaphene) to be over 3,000 ppb, 500 ppb and

40, 000 ppb, respectively, with a total OCP load of over 68,000 ppb (Marburger et al.,

2002). The EMCA study also determined concentrations of OCPs in various tissues of

several fish species. Black crappie (Pomoxis nigromaculatus) generally showed less total

OCP accumulation than largemouth bass or brown bullhead catfish (Ameiurs nebulosus).

Concentrations in largemouth bass ovary and fat reached over 4,000 ppb and 17,000 ppb,

respectively for total DDT derivatives; 100 ppb and 700 ppb, respectively for dieldrin;

and 4,000 ppb and 20,000 ppb, respectively for toxaphene (Marburger et al., 2002).

Selection of Animal Model

The demand for a sports-fishing industry and environmental relevance as an

indicator species make the Florida largemouth bass an excellent model for this and

several historical studies. Marburger et al. (2002) indicated largemouth bass as

bioaccumulating large amounts of OCPs from the environment. This can be attributed to

top predator status of largemouth bass and persistence of OCPs in the environment due to

their lipophilicity. In addition to high contaminant concentrations, depressed sex steroids

(17p-estradiol (E2) and 11-ketotestosterone (11-KT)), reversed sex steroid ratios, and

reduced survival of fry have been observed in fish from these areas (Benton and Douglas,

1996; Marburger et al., 1999). Because largemouth bass are an important sport-fish in

Florida and nationally, much effort is being placed into the restoration of a viable fishery

in the EMCA and Upper Ocklawaha River Basin (Benton et al., 1991; Benton and









Douglas, 1994; Marburger et al., 1999). For these reasons, the Florida largemouth bass

was chosen as an animal model for laboratory studies of OCP exposure.

Largemouth Bass Reproductive Cycle

In Florida, largemouth bass reach sexual maturity at 1 year of age and generally at

a length of 250 mm (Chew, 1974). They are synchronous spawners with a spawning

season that can extend from January to April (Clugston, 1966). It is generally accepted

that the act of spawning is triggered by an increase in water temperature in the spring to a

range of 20 24C. Carr (1942) found that peak spawning occurs in the Gainesville,

Florida area in late March.

Teleost fish, such as the largemouth bass, follow a well-established reproductive

cycle that is influenced by environmental (temperature and photoperiod) and endogenous

(hormone) factors. These factors stimulate the hypothalamus to secrete gonadotropin-

releasing hormone (GnRH), norepinephrine (NE) and other neuropeptides that act to

stimulate the pituitary to secrete the primary teleost gonadotropins GTH-I and GTH-II

(Van Der Kraak et al., 1998). GTH-I and GTH-II have varying roles but will generally

stimulate gonadal sex steroid hormone production in preparation for spermiation in males

and oocyte maturation, vitellogenesis, and ovulation in females. Increasing E2

concentrations in females stimulates the liver to produce vitellogenin, a

phosphoglycolipoprotein that serves as a yolk precursor in oviparous vertebrates (Wahli

et al., 1981). Vitellogenin produced in the liver is released into circulation to travel to the

gonad where it is sequestered as a nutrient source in developing oocytes. The

vitellogenin gene has an estrogen responsive element in the promoter region and is

transcribed in response to an estrogen-estrogen receptor (ER) complex (Wahli et al.,

1981). Both males and females have copies of the vitellogenin gene, however it takes a









threshold concentration of E2 (only normally achieved in females) to produce measurable

levels of vitellogenin (Wallace, 1970; Ryffel, 1978; Hori et al., 1979; Wahli et al., 1981;

Copeland et al., 1986). Because males do not normally produce vitellogenin, it has been

used in many studies as an indicator of exposure of males to estrogenic compounds

(Pelissero et al., 1993; Sumpter and Jobling, 1995; Shelby et al., 1996; Monosson et al.,

1996; Celius and Walther, 1998; Smeets et al., 1999; Madsen et al., 2002; Okoumassoun

et al., 2002a and 2002b).

A study performed by Gross, et al. (2002) on hatchery-reared bass in Florida

characterized the annual cycles of circulating sex steroid hormones, vitellogenin, and

gonad development. Circulating sex steroid hormones begin to rise prior to the spawning

season. For male bass, 11-KT was the only sex steroid observed to show a strong

seasonal pattern and had a peak concentration of about 2,800 pg/ml in February. E2 was

detected in males but at concentrations of about one-third of females. Female bass

showed three distinctly seasonal hormones: E2, testosterone (T), and progesterone (P). E2

showed the strongest pattern with circulating levels nearly twice those of T at a peak

concentration of 4,000 pg/ml in February. Circulating vitellogenin levels closely

mimicked those of E2, rising in November and peaking in January at about 6 mg/ml.

Progesterone reached a peak concentration in early April corresponding with peak

spawning activity, indicating a possible role in ovulation. 11-KT was detected in

females, but at levels nearly one-half that of males. Similar peak concentrations of sex

steroids were found in wild-caught female bass from Lake Woodruff, Florida (Timothy S.

Gross, unpublished data).









Corresponding gonad weight as a percentage of body weight or gonadosomatic

index (GSI) begins to rise in October for both females and males, peaking in February to

March at approximately 6% and 2%, respectively (Gross et al., 2002). Increasing GSI

correlates with an increase in gonad maturation as seen histologically for both sexes

(Nelson, 2001). In females, much of the dramatic increase in gonadal weight is due to

the sequestering of a large amount of vitellogenin in developing oocytes (Wallace and

Selman, 1981; Van Der Kraak et al., 1998). A similar study of Florida largemouth bass

performed in Texas showed a peak GSI between March and April (Rosenblum et al.,

1994).

OCP Contaminant Effects on Fish Reproduction

Although some reproductive parameters such as circulating sex steroids, gonad

development, and reproductive success in fish (Marburger et al., 1999 and 2002;

Gallagher et al., 2001) and American alligators, Alligator mississippiensis, (Gross et al.,

1994; Guillette et al., 1994; 1995; 1999a; 1999b) have been shown to be abnormal in the

contaminated areas of the Ocklawaha basin, there are few studies linking OCPs as a

causative agent.

Many OCPs are thought to be endocrine disrupters and much research has been

focused in this area. Exposure to some OCPs by painting them onto the surface of red-

eared slider turtle, Trachemys script elegans, (Willingham and Crews, 1999) and

alligator (Gross et al., 1994) eggs showed thatp,p '-DDE can cause sex reversal,

producing females when incubated at male producing temperatures. p,p '-DDE has also

been shown to be a potent anti-androgen in rats (Kelce et al., 1995) by binding to the

androgen receptor (AR); therefore, preventing the transcription of testosterone resulting

in demasculinization. Similar results have been found in some fish species such as









guppies, Poecilia reticulata, (Baatrup and Junge, 2001; Bayley et al., 2002), white

sturgeon, Acipenser transmontanus, (Foster et al., 2001), and goldfish (testes only),

Carassius auratus, (Wells and Van Der Kraak, 2000). Alternatively, p,p '-DDE can also

bind to the ER in Hartley guinea pigs (Danzo et al., 2002), channel catfish, Ictalurus

punctatus, (Nimrod and Benson, 1997), rainbow trout, Onchorhynkus mykiss (Matthews

et al., 2000), and largemouth bass, (Garcia et al., 1997; Larkin et al., 2002). Other

prominent OCPs such as methoxychlor, DDT, dieldrin, and endosulfan, are also able to

bind to the ER of human, mouse (Mus musculus), chicken (Gallus gallus), green anole

(Anolis carolinesis) and rainbow trout (Matthews et al., 2000; Tollefsen et al., 2002a).

Matthews et al. (2000) demonstrated that methoxychlor, p,p '-DDT, o,p '-DDE, p,p '-DDE,

a-endosulfan, and dieldrin are able to bind weakly to GST-ER fusion proteins from

human, mouse, chicken and green anole, as well as, to displace all E2 from GST-ER

fusion proteins from a rainbow trout demonstrating a high binding affinity.

The ability of any exogenous compound to bind a sex steroid hormone receptor

and agonize and/or antagonize the action of an endogenous hormone can severely affect

normal endocrine function and ultimately lead to a decrease in reproductive success.

Normal androgen concentration and action are critical for development of the male fish

gonad and spermatogenesis. Chemicals able to bind the ER can increase or decrease

estrogen function. In males, even weak binding of a substance to the ER could induce an

upregulation of estrogen function because males generally show only a low-level

endogenous estrogen expression. This could lead to the production of vitellogenin in

males, possibly coupled with blocked AR functions, potentially resulting in

demasculinization (lowered testosterone level and testosterone-induced functions) and









feminization (increased estrogen-like function). Vitellogenesis in cultured hepatocytes

from male fish has been shown to occur with exposure to o,p'-DDT ( Sumpter and

Jobling, 1995; Smeets et al., 1999; and Leafios-Castafieda et al., 2002), tamoxifen,

nonylphenol, bisphenol A (Sumpter and Jobling, 1995), and phytoestrogens (Pelissero et

al., 1993). Some in vivo studies have shown vitellogenin production induced in male and

female fish exposed to OCPs (Okoumassoun et al., 2002b; Madsen et al., 2002; Leafios-

Castafieda et al., 2002). In females, competitive binding of the ER could reduce

estrogen function, however, the exogenous compound will likely have a much lower

potency than the endogenous estrogen. Inhibitory binding to the ER could lead to

impaired gonadal development and decreased vitellogenesis, ultimately ending in poor

egg quality and decreased reproductive success.

Fish and alligators living in areas with high OCP levels demonstrate depressed

circulating sex steroid levels (Guillette et al., 1995; Crain et al., 1997; Guillette et al.,

1999a, 1999b; Marburger et al., 1999; Foster et al., 2001; Sepulveda et al., 2002). This

could be a result of a disruption in the hypothalamus-pituitary-gonad axis. No data was

found on gonadotropin level changes in response to OCP exposure, however, studies on

fish exposed to bleached kraft mill effluent have shown concomitant decreases in

gonadotropin and sex steroid levels (Van Der Kraak et al., 1992; McMaster et al., 1995).

Another possibility for OCP action includes inducing or inhibiting the activity of

liver biotransformation enzymes that could increase/decrease the metabolism and

excretion of any xenobiotics, as well as, endogenous hormones. An intensely used

marker of induced cytochrome P450, a major class of phase I biotransformation enzymes

and specifically CYP1A, is the activity of ethoxyresorufin-O-deethylase (EROD). High









OCP levels can induce EROD in the white sturgeon (Foster et al., 2001), Nile tilapia

(Oreochromis niloticus) (Zapata-Perez et al., 2000), rat (Wade et al., 2002), and deer

mouse (Peromyscus maniculatus) (Dickerson et al., 1999). Dickerson et al. (1999) and

Foster et al. (2001) also noted a decrease in circulating sex steroids in the deer mouse and

white sturgeon, respectively, correlating with the increase in EROD activity. Also, a

study on English sole (Parophrys vetulus) from Puget Sound, an area contaminated with

PCBs and PAHs, showed significant negative correlations between mixed function

oxygenase (MFO) induction and estradiol concentrations (Johnson et al., 1988). Because

CYP1A is not the major isoform for metabolizing sex steroid hormones in fish, increased

EROD activity is not likely related to the decreases in sex steroid concentrations

(Snowberger and Stegeman, 1987). On the other hand, CYP3A is known to hydroxylate

sex steroids in fish (Kullman et al., 2000). While no studies were found that had tested

CYP3A induction/inhibition in fish exposed to pollutants, studies have shown it is

induced in human hepatocytes and rats exposed to OCPs (You et al., 1999; Coumoul et

al., 2002). More research is needed to elucidate the mechanism behind the correlation of

biotransformation enzyme induction and decreased sex steroid levels. Increased

metabolism and clearance of sex steroids may be due to the heightened activity of major

metabolizing enzymes.

Research Significance

This research aims to provide information on dosing efficiency and the

reproductive response of largemouth bass to p,p '-DDE and dieldrin exposure, two

predominant OCPs found in the Upper Ocklawaha River Basin. Single chemical

exposures were performed to assess the potential contribution of each pesticide to overall

reproductive function and steroidogenic declines known to occur when fish are






10


environmentally or experimentally exposed to contaminant mixtures. Endpoints were

measured on multiple levels of biological organization (biochemical, organ, and whole

body) to gain insight into the ecological relevance of the response measured.














CHAPTER 2
EXPOSURE OF LARGEMOUTH BASS TO p,p '-DDE AND DIELDRIN BY SLOW-
RELEASE MATRIX PELLET

Introduction

The most prominent organochlorine pesticides (OCPs) found in soils of the

Ocklawaha River Basin are DDT derivatives, toxaphene, and dieldrin (Marburger et al.,

2002). These chemicals bioaccumulate in Florida largemouth bass (Micropterus

salmoidesfloridanus) and other top predators, such as the alligator (Alligator

mississippiensis), and are suspected causes for decreased reproductive success of these

animals due to impaired endocrine function (Gross et al., 1994; Guillette et al., 1994;

1995; 1999a; 1999b; Gallagher et al., 2001).

p,p '-DDE and dieldrin were selected to characterize single chemical dose-response

relationships for reproductive function in Florida largemouth bass. Characterization of

the reproductive effects of two major OCP contributors to this system will aid in

determining if it is the actions of either of these chemicals alone that is a causative agent

for depressed sex steroid concentrations and will also assist in future studies determining

their contribution to the larger complex mixture that exists in the area.

The objectives of this study were to evaluate the slow-release pellet dosing method

and to determine potential dose effects ofp,p '-DDE and dieldrin on reproductive function

in largemouth bass. Health and reproductive parameters were measured on multiple

levels of biological organization: organism (weight, length, and condition factor (K)),

organ (gonadosomatic index (GSI) and hepatosomatic index (HSI)) and biochemical









(circulating hormones). This diversity of endpoints should provide insight into the

biological significance of the doses tested.

Specific Aims

1. Create internal doses of eitherp,p '-DDE or dieldrin in largemouth bass prior to the
onset of spawning season and evaluate the actual dose achieved.

2. Evaluate the effects ofp,p '-DDE or dieldrin dose on health (HSI and K) and
reproductive (GSI and circulating sex steroid hormones) biomarkers.

3. Determine the biological relevance of the dose response.

Null Hypothesis

Reproductive and health parameters will not vary across treatments regardless of

pesticide or dose administered.

Alternative Hypothesis

Because these chemicals have been shown to be endocrine active and wildlife

exposed to them exhibit decreased circulating hormone concentrations, it is hypothesized

that exposure of largemouth bass top,p '-DDE or dieldrin in a laboratory setting will also

cause a decrease in circulating hormones which, in turn, will alter gonadal development,

resulting in a decreased GSI. The doses administered were within the range of what is

found in large, seemingly healthy adult bass in the wild. Therefore, no effects on weight,

length, or K were predicted.

Materials and Methods

Experimental Animals

Florida largemouth bass of one to two years of age were obtained from a fish

hatchery (American Sports Fish, AL) in December 2001. The fish were transferred to the

United States Geological Survey Biological Resources Division Center for Aquatic

Resource Studies (USGS-BRD-CARS) facility where they were housed in 6,116 liter









concrete runs (366 cm x 183 cm x 91 cm) equipped with a flow-through system supplied

by on-site pond water and aeration. On Day 0 of the experiment (January 11, 2002) bass

had a mean SD weight, length, and K of 141.9 22.6 g, 213.1 + 10 mm, and 1.46 +

0.11, respectively, indicating fish were healthy and of reproductive size.

Water quality was measured as dissolved oxygen, temperature, and pH twice a

week. Water quality parameters were all within acceptable ranges for the duration of the

experiment: dissolved oxygen ranged from 4.28 to 10.20 mg/L (measurements were

taken at 8 a.m. when oxygen levels would be expected to be at the lowest), temperature

ranged from 11.7 to 24.0 C, and pH ranged from 7.5 to 8.6. Fish were fed "Floating

Finfish Silver" feed (Zeigler Bros., Inc. Gardners, PA) ad libitum twice a week.

Experimental Design

Sixty-day release pellets were inserted intraperitoneally with a steel trochar into

each of 25 fish per treatment on the eleventh of January, 2002. Treatments included 3

doses ofp,p-DDE (2.5 mg, 5.0 mg, and 10.0 mg), 3 doses of dieldrin (0.25 mg, 0.5 mg,

and 1.0mg), placebo (matrix only) and sham (no pellet inserted), total of 200 fish. Each

fish was also implanted with an intraperitoneal pit tag (Trovan Corp., Bel Air, MD) to

identify individuals for repeat measurements. Fish from all treatments were randomly

divided into two concrete runs holding 100 fish each. Each run contained fish from all

treatment groups. Sex ratios within a tank were unknown because the sex of a bass is

undistinguishable externally during this time of year.

Chemicals and Dosing

The organochlorine pesticides 2,2-bis(4-chlorophenyl)-1,1-dichloroethylene (p,p '-

DDE, Lot # 09020KU, 99.4% purity) and 1,2,3,4,10,10-hexachloro-6,7-epoxy-

1,4,4a,5,6,7,8,8a-octahydro-1,4,5,8-dimethanonaphthalene (dieldrin, Lot # 077H3578,









91.2% purity) were obtained from Aldrich Chemical Company (Milwaukee, WI). The

pesticides were shipped to Innovative Research of America, Inc. (Sarasota, FL) where

they were incorporated into 60-day slow-release matrix pellets at doses of 2.5 mg, 5.0 mg

and 10.0 mgp,p'-DDE and 0.25 mg, 0.5 mg and 1.0 mg dieldrin, per pellet. Cholesterol,

lactose, celluloses, phosphates and, stearates where used as carriers and chemical binders

in the pellets. Concentrations of chemical in the pellets were calculated, based on

average weight of the fish used in the study and lipid content of the gonad, to cover the

range of internal doses found in the Ocklawaha system.

Blood and Tissue Collection

Weight, to the nearest gram (g), and length in millimeters (mm), were measured

and blood was collected from each fish at Days 0 (1/11/02), 30 (2/10/02), and 60

(3/12/02). Condition factor was calculated for each time point (K = weight/length3 x

100). At each collection, approximately 1 ml of blood was obtained from the caudal vein

using a heparinized 20-gauge needle, dispensed into a 3 ml heparinized vacutainer,

labeled and stored on ice until centrifuged. Blood samples were centrifuged at 1,000 x g,

4C for 20 minutes to separate red blood cells from plasma. Plasma was removed with a

transfer pipette, placed in a cryovial and stored at -80C.

At the end of exposure (Day 60) all fish were sacrificed with a blow to the head.

Liver and gonads were removed and weighed to the nearest 0.01g for determination of

GSI (gonad weight/body weight x 100) and HSI (liver weight/body weight x 100). Liver,

gonad, muscle, and blood samples were collected for contaminant analysis from 2 fish

per treatment, and gonads only from a subset of five females per treatment. Female

gonads were also processed for histological analysis of reproductive stage. A cross-

section of one ray of each collected gonad was placed in a histological cassette and fixed









in 10% buffered formalin. Gonad, liver, and muscle samples were wrapped in tin foil and

blood remained in the glass vacutainer; all were frozen at -20C until OCP analysis by

gas chromatography-mass spectrometry (GC-MS).

Gonad Histology

Tissue staining with hematoxylin and eosin (H&E), sectioning, and slide

mounting was performed by Histology Tech Services (Gainesville, FL). Ovaries were

observed under a light microscope at 40X and stage of sexual maturation was assigned

(Sepulveda, 2000). Briefly, stage 1 ovaries are undeveloped with mostly primary phase

follicles. Stage 2 ovaries are previtellogenic with primary and secondary phase follicles,

but have no vitellogenic follicles. Stage 3 ovaries are early vitellogenic with some

vitelline granules in follicles of varying size and no fully developed eggs. Stage 4 ovaries

are late vitellogenic with a majority of follicles containing numerous vitelline granules

and fully developed eggs are present.

Analysis of Circulating Sex Steroid Hormones

Blood plasma from largemouth bass was analyzed for 11-KT and E2 using a

previously validated 3H radioimmunoassay (RIA) method (Gross et al., 2000). All

samples were assayed in duplicate and values were reported as pg/ml of plasma.

Standard curves were prepared in phosphate buffered saline plus gelatin and sodium

azide (PBSGA) with known amounts (15, 30, 60, 125, 250, 500, 1000, and 2000 pg) of

radioinert E2 (ICN Biomedicals, Costa Mesa, CA) or 11-KT (Sigma Chemicals, St.

Louis, MO) and 15,000 cpm of 3H-E2 or 3H-11-KT. Each plasma sample (50 [l) was

extracted twice with diethyl ether prior to RIA analysis. Diethyl ether (4 ml) was added

to each sample, tubes were vortexed and the ether was evaporated off, leaving behind the

solid phase precipitate. The procedure was repeated in the same test tube to insure









evaporation of all liquid plasma. PBSGA buffer and antibodies against sex steroid

hormones were then added to the sample tube and incubated overnight at 4C.

Antibodies were purchased from ICN Biomedicals (E2) or Helix Biotech, Richmond, BC,

Canada (11-KT). Following the antibody incubation, unbound antibody was removed by

addition of dextran-coated charcoal and centrifugated for 10 min at 1,000 x g. Excess

free antibody was stripped out of solution by binding to the charcoal and pelleting at the

bottom of the test tube. Four hundred [l of sample supernatant was removed and added

to a scintillation vial with 4 ml of Scintiverse scintillation cocktail (Fisher Scientific,

Pittsburgh, PA). Samples vials were then placed in a liquid scintillation counter (Pachard

Tricarb, Model 1600) and counted for two minutes each. The minimum concentration

distinguishable from zero (mean SE) was 94 14 pg/ml for E2 and 29 13 pg/ml for

11-KT. Cross reactivities of the E2 antiserum (produced and characterized by T.S. Gross,

University of Florida) with other steroids were: 11.2% for estrone, 1.7% for estriol, and <

1.0% for 17a-estradiol and androstenedione. Cross reactivity of the 11-KT antiserum

with other steroids was: 9.65% for testosterone, 3.7% for a-dihydrotestosterone, and <

1.0% for androstenedione.

Analysis of Largemouth Bass Tissues for OCPs

Chemical analysis was carried out at the Center for Environmental and Human

Toxicology, University of Florida. Briefly, largemouth bass tissues (whole liver, half of

the gonad, and left and right side muscle fillets) were homogenized to eliminate any

concentration variability within the tissue due to differences in lipid content. A portion

of the sample (2-5 g) was extracted into ethyl acetate. The sample was purified using

C18 and NH2 SPE (solid phase extraction) cartridges. Total OCP content was

determined by GC-MS, according to EPA method 8270 (EPA, 1983). Readings were not









lipid normalized. Samples were analyzed multiple times in full scan mode for analyte

identification and in selected ion mode (SIM) for quantitation to improve sensitivity.

Percent recovery ranged between 75 and 100% with a limit of detection of 0.75-1.5 ng/g.

Analysis of Largemouth Bass Blood Plasma for OCPs

Blood plasma concentrations ofp,p '-DDE and dieldrin were determined by a

competitive enzyme linked immunosorbent assay (ELISA) (Abraxis, LLC. Warmister,

PA; DDE kit Lot#3C06437, Cyclodiene kit Lot #3C0643). In a competitive ELISA, a

known amount of substrate (in this case p,p '-DDE or dieldrin) is immobilized to the

bottom of a 96-well plate. The sample to be tested is added, along with the specific

antibody for the substrate. The substrate in the sample competes with the substrate bound

to the plate for conjugation to a known amount of primary antibody. All sample and

excess primary antibody are washed from the plate leaving the immobilized substrate

with a percentage bound to the primary antibody. A secondary antibody is then added

that is conjugated to a horseradish-peroxidase enzyme. This antibody only binds to the

primary antibody and not to substrate that is not bound. Excess secondary antibody is

then washed from the plate and a substrate for the enzyme linked to the secondary

antibody, hydrogen peroxide, is added. When the hydrogen peroxide solution is cleaved

by the horseradish-peroxidase, a blue color results. This reaction is stopped after a

certain period of developing time by addition of a dilute acid solution. The plate is then

read at 450 nm for quantification of the amount of color in each well, which is inversely

proportional to the amount of substrate in the sample.

First, all kit components and samples were first brought to room temperature.

Each sample was diluted 1:50 in 10% methanol for DDE and 1:100 for dieldrin in 25%

methanol. Concentrated wash buffer (5X) provided in the kit was diluted 1:5 with









distilled water. Standards for DDE were made by addingp,p '-DDE stock solution

(10,000 ppb) to a 1:50 dilution of plasma from untreated bass with 10% methanol as per

instructions. Standards for dieldrin were included in the kit and required only 1:100

dilution with 25% methanol in a glass test tube. Parafilm was placed over each test tube

and vortexed to mix. On a 96 well plate, 25 al of each standard and sample was added to

separate wells. Standards and samples were run in duplicate. Next, 100 ul of rabbit anti-

DDE/anit-cyclodiene antibody was added to each well. Wells were covered with tape

and contents of the plate were mixed by moving the plate in a circular motion on the

bench top for approximately 30 s. The plate was then incubated at room temperature for

60 min. Following incubation, the covering was removed and the contents dumped by

vigorous shaking. Wells were then washed three times with 250 ul 1X wash solution,

followed by blotting off excess wash with a paper towel. Next, 100 ul of anti-rabbit-

horseradish peroxidase secondary antibody was added to each well, covered, and allowed

to incubate at room temperature for 30 min. Following incubation, the covering was

removed, contents dumped, and wells washed 3 times in 1X was solution as before. One

hundred ul of hydrogen peroxide color solution was then added to each well and

incubated at room temperature for 20 min. Fifty ul of acidic stopping solution was then

added to each well. The plate was read at 450 nm on a MRX Microplate Reader (Dynex

Technologies) within 15 min of application of the stopping solution. Results are given as

ng/ml.

The limit of detection forp,p '-DDE was 62.5 ppb. Cross reactivity of thep,p '-

DDE antibody were 46%p,p '-DDD, 16% o,p '-DDD, 10%p,p '-DDT, and 3.2% o,p '-

DDE. The limit of detection for dieldrin was 25 ppb. Cross reactivity of the cyclodiene









antibody were 150% endosulfan, 58% heptachlor, 26% aldrin, 26% chlordane, and 8.2%

toxaphene. High cross reactivity to other chemicals did not influence the analysis for this

experiment as all fish were from an uncontaminated environment and had low to no

background exposure to other chemicals.

Statistical Analysis

All statistical analyses were performed using Statistical Analysis System (SAS)

software, version 9. The original experimental design called for statistical analysis

comparing each treatment group by ANOVA and a multiple comparison procedure. The

Univariate procedure was run on all data sets to determine if the data was normally

distributed. ANOVAs were then performed and significance was set at a = 0.05.

Duncan's Multiple Range test followed as a multiple comparison procedure to determine

which treatment groups differed. Results are presented as means + SD.

As detailed in the results section, dosing was inconsistent within treatment groups.

Therefore, further statistical analysis was done by linear regression of each parameter

against the blood plasma chemical concentration of each individual fish (SAS, a = 0.05).

The independent variable was blood plasma concentration of either chemical and the

dependent variables were weight, length, K, HSI, GSI, blood plasma E2 and 11-KT

concentrations. Percent change in hormone concentrations from Days 0 to 30 and from

Days 30 to 60 were also regressed against Day 30 and Day 60 blood plasmap,p '-DDE

and dieldrin concentrations, respectively. Individual fish that exhibited increased

concentrations of both p,p '-DDE and dieldrin were considered outliers and removed from

the data set because possible interactions of these pesticides remain unknown.









Results

Largemouth Bass

Survivorship over the course of exposure was 95.5% excluding one incident that

occurred on January 22 where water flow was lost in both tanks for approximately 10

hours overnight and 15 fish died due to dissolved oxygen levels falling below 3 mg/L.

This incident brought survivorship down to 88%. No long term adverse effects on

surviving fish were observed. Weight, length and K steadily increased over the course of

60 days for males and females (Figures 2-1 and 2-2) and did not vary between treatments.

Health Parameters: K and HSI

Condition factor did not vary among treatments in females or males over the course

of the study. Female Day 60 HSI did not vary between treatment groups or from

controls; however, males in the dieldrin 1.0 mg treatment group had a significantly higher

HSI compared to all other treatments (Figure 2-3).

Gonad Histology

Histological analysis of female gonads taken at Day 60 indicated that 89.2% of

females were in late vitellogenesis (stage 4) with the remaining in early vitellogenesis

(stage 3) (Figure 2-4).

Reproductive Parameters: GSI and Circulating Sex Steroids

GSI

GSI, determined on Day 60, did not differ between treatment groups in neither

females nor males (Figure 2-5). Average GSI was 4.0 1.7% (n = 65) and 0.66 0.16%

(n = 90) for females and males, respectively.









17p-Estradiol

In female bass, the placebo and DDE 2.5 mg treatment groups consistently showed

higher, though not always significant, circulating E2 concentrations. At Day 0 the

females in the placebo group had a significantly higher E2 concentration compared to

DDE 5.0, DDE 10.0, and all dieldrin treatment groups (Figure 2-6a). The E2

concentrations in these treatments, however, did not differ from the Sham treatment. At

Day 30, females in the placebo and DDE 2.5 treatment groups had a significantly higher

E2 concentration as compared to all other treatments (Figure 2-6b). By Day 60,

variability within treatment groups had increased and only females from the DDE 2.5

group demonstrated a significantly higher E2 concentration than DDE 10.0 and dieldrin

0.5 females (Figure 2-6c).

In male bass, the DDE 5.0 mg treatment group consistently demonstrated higher

circulating E2 concentrations as compared to all other treatments. At Days 0 and 30 all

treatments except DDE 5.0 exhibited no significant difference from each other (Figure 2-

7a,b). At Day 60 variability within treatment groups had increased. Males in the DDE

5.0 treatment group had a significantly higher circulating E2 concentration as compared

to males in the sham and dieldrin 0.5 treatments (Figure 2-7c).

11-Ketotestosterone

In female bass, most treatment groups demonstrated much higher variability in

circulating 11-KT concentrations as compared to controls. At Day 0, females in the DDE

5.0, dieldrin 0.5, and dieldrin 1.0 had a significantly higher 11-KT concentration than

sham, placebo, DDE 2.5, and DDE 10.0 treatments (Figure 2-8a). By Day 30, only

females in the DDE 5.0 treatment group had significantly higher 11-KT concentration









(Figure 2-8b). At Day 60 all treatment groups had similar 11-KT concentrations (Figure

2-8c).

In male bass, placebo and DDE 2.5 treatment groups demonstrated a consistently,

though not always significant, higher circulating 11-KT concentration as compared to

sham and other treatment groups. At Day 0, males in the DDE 2.5 treatment group had

significantly higher 11-KT concentrations than males in the sham group (Figure 2-9a).

At Day 30, 11-KT concentrations were significantly higher in the DDE 2.5 treatment

group as compared to all treatments except placebo (Figure 2-9b). Day 60 11-KT

concentrations for males in the DDE 2.5 treatment group remained significantly higher

than sham, DDE 10.0 and all dieldrin treatments (Figure 2-9c).

In Vivo Treatment Dosing Consistency

Contaminant analyses of gonad (n = 7 per treatment), liver, muscle, and blood (n

= 2 per treatment), revealed that either the pellets within each treatment group differed or

the pellets did not release a consistent dose. Some fish had exposure to bothp,p '-DDE

and dieldrin indicating some pellets may have been cross-contaminated (Figure 2-10;

Table 2-1). Also, examination of the peritoneal cavity from each fish at the end of the

study demonstrated that pellets had varying degrees of dissolution (i.e. some pellets

remained whole while others completely dissolved). Consequently, gonadal

concentrations reached in each treatment group were highly variable. Despite the high

variability in gonadal dose, mean p,p '-DDE gonadal concentrations demonstrated an

apparent dose-response (Figure 2-10a). The range of gonadal p,p '-DDE concentrations

achieved in this study included the range of gonadal p,p '-DDE levels found in 1996 of 3

year old bass in the Emeralda Marsh Conservation Area (EMCA), 1300 to 4200 ppb

(total DDT derivatives) (Marburger, 2002). On the other hand, gonadal dieldrin doses









were not specific for their target dose (i.e. the low dieldrin treatment group received a

mean dose higher than that of the high dieldrin treatment group) and the range achieved

greatly exceed that which is found in the wild for gonadal dieldrin, 70 to 130 ppb

(Marburger, 2002; Figure 2-10b).

Because tissue (liver, gonad, muscle) was analyzed from only two fish per

treatment, it was impossible to determine individual exposures for every animal.

However, a blood plasma sample was obtained from every fish at three time points.

Therefore, a linear regression analysis (SAS) was performed to determine how well blood

concentrations correlated with specific tissue concentrations for each chemical. In

females and males, p,p'-DDE blood concentrations were significantly correlated to liver

(female liver = -8 + 1.5(blood), r2= 0.8345; male liver = -46 + 1.4(blood), r2= 0.8807)

and gonad (female gonad = 120 + 10(blood), r2= 0.9520; male gonad = 0.4 +

10.5(blood), r2= 0.7520). However, muscle only had a significant relationship with

bloodp,p '-DDE concentration in females (female muscle = 36 + 0.25(blood), r2 = 0.5858;

male muscle = 9 + 0.5(blood), r2 = 0.4041) (Figure 2-11). For dieldrin, female and male

blood concentrations were significantly correlated to liver (female liver = -5 + 0.6(blood),

r2= 0.9986; male liver = 52.5 + 0.3(blood), r2= 0.9650) and gonad (female gonad = -51 +

5.2(blood), r2 = 0.9262; male gonad = 412 + 1.5(blood), r2 = 0.6824), but muscle dieldrin

was only correlated with blood dieldrin in males (female muscle = -9 + 0.3(blood), r2 =

0.8527; male muscle = 22.7 + 0.14(blood), r2= 0.9120) concentrations (Figure 2-12).

Because a significant linear relationship between blood and tissue concentration existed,

a method for analyzing blood plasma concentrations was established. However, caution

should be taken in extrapolating blood plasma concentrations to gonadal concentrations,









especially dieldrin, as some linear regression models showed a weak goodness of fit,

explaining less than half of the variability of the data set.

New ELISA detection kits forp,p '-DDE and dieldrin were developed by Abraxis,

LLC (Warminster, PA) for detection in largemouth bass blood plasma. This method was

validated in our lab by analyzing a pool of blood plasma samples from treated bass in

parallel by GC-MS and the new ELISA method. A linear regression was used to

determine the correlation of results obtained by the two methods for both pesticides

(Figure 2-13). It can be seen in the graph of this data that the ability of the ELISA to

detect concentrations below 100 ppb is limited. All control fish, most treated fish at Day

30, and some treated fish at Day 60 register OCP concentrations below 100 ppb

indicating they are at, near, or below the limit of detection of the ELISA kits. However,

correlation between GC-MS and ELISA methods was highly significant and linear for

bothp,p '-DDE and dieldrin. Therefore, ELISA was used to analyze blood samples from

every treated fish, and 2 males and 2 females from control groups (mean concentrations

of these fish were substituted in as the exposure level for all other control fish), at Days

30 and 60 in order to assess the achieved dose over time.

Health Parameters: K and HSI (Regression)

At Day 0, all fish were healthy with an average condition factor of 1.46 0.11.

Condition factor increased steadily over the course of the exposure period, as body

weight increased mostly due to the developing gonads (Figure 2-1c and Figure 2-2c).

HSI was determined at Day 60 and did not show a significant correlation with blood

plasma concentration of eitherp,p'-DDE (Figure 2-14) or dieldrin (Figure 2-15).

Average HSI SD for all fish at Day 60 was 3.02 0.16% (n = 156).









Reproductive Parameters: GSI and Circulating Sex Steroids (Regression)

GSI

GSI was determined on Day 60 and neither males nor females showed a significant

relationship with blood plasma concentration of either p,p'-DDE (Figure 2-16) or dieldrin

(Figure 2-17). Average GSI SD was 4.0 1.7% (n = 65) and 0.66 0.16% (n = 90) for

females and males, respectively.

17p-Estradiol

On Day 0 (January 11) female and male bass had mean SD E2 concentrations of

527 278.6 and 264 106.4 pg/ml, respectively (Figure 2-18). E2/1 1-KT ratio at Day 0

was 2.5 + 1.9 for females and 0.5 + .2 for males.

In fish treated withp,p '-DDE, both Day 30 circulating E2 and percent change in E2

from Days 0 to 30 showed a significant negative relationship with increasing blood

plasma concentration (r2 = 0.2815 and 0.1401, respectively) in females (Figure 2-19)

while no relationship between Day 30 E2 and blood plasmap,p '-DDE existed in males

(Figure 2-20). Circulating E2 also showed a significant negative relationship with blood

plasmap,p'-DDE (r2 = 0.1588) at Day 60 in females (Figure 2-21a). However, the GSI

and percent change in E2 from Days 30 to 60 showed no relationship with blood plasma

concentration (Figure 2-21b,c). Again, males exhibited no relationship between blood

plasma p,p '-DDE and E2 (Figure 2-22).

In fish treated with dieldrin neither females nor males showed a significant

relationship with blood plasma concentration at Day 30 (Figures 2-23 and 2-24). At Day

60, female circulating E2 (r2 = 0.0933), but not percent change in E2 from Days 30 to 60,

showed a significant negative relationship with blood plasma dieldrin (Figure 2-25a,b).

This decrease in E2 was not correlated with GSI (Figure 2-25c). Males showed no









relationship between E2 and blood plasma dieldrin concentrations at Day 60 (Figure 2-

26).

11-Ketotestosterone

Day 0 circulating 11-KT concentrations were 301 + 211.6 and 731 345.9 pg/ml,

respectively for females and males (Figure 2-18).

In fish treated with p,p '-DDE, Day 30 circulating 11-KT concentrations showed no

relationship with blood plasma concentration in females (Figure 2-19) or males (Figure 2-

20). Also, at Day 60 neither females nor males showed a relationship between 11-KT

and blood plasmap,p '-DDE or GSI (Figures 2-21 and 2-22).

At Day 30, neither sex of fish treated with dieldrin showed a significant

relationship between blood plasma concentration and circulating 11-KT (Figures 2-23a

and 2-24a). However, males did show a trend of decreasing 11-KT and a significant

negative relationship between blood plasma dieldrin and percent change in 11-KT from

Days 0 to 30 (r2 = 0.1627) (Figure 2-24). At Day 60, females again demonstrated no

relationship between blood plasma dieldrin concentration and 11-KT (Figure 2-25). Male

circulating 11-KT had a significant negative relationship with blood plasma dieldrin at

Day 60 (r2 = 0.1610) and no relationship with GSI. Also, no relationship existed between

blood plasma dieldrin concentration and percentage change in 11-KT from Days 30-60

(Figure 2-26).

Discussion

The data collected from this study indicates that time-release pellets are not an ideal

dosing route for largemouth bass and represents a common problem in ecotoxicology

research: applied vs. attained dose. Many studies simply apply a dose and do not perform

the necessary analysis to determine 1) how much of the applied dose was attained, 2) if









the applied dose reached the target tissue, and 3) the variability of the attained dose

within treatment groups or replicates.

This study demonstrated the extreme variability inherent to the pellet dosing

method (Figure 2-10). Tissue samples were taken at two timepoints during the study to

examine tissue difference in bioaccumulation/chemical partitioning, but also served as an

indicator of variability of dose within treatment groups. Originally, the study design

called for statistical analysis (ANOVA) to be run by treatment group. ANOVA and

regression analyses were in agreement and showed no biologically significant differences

between treatment groups for weight, length, K, HSI or GSI. However, the two analysis

methods resulted in different conclusions of dose effects on circulating sex steroids.

ANOVA analysis indicated high doses ofp,p '-DDE and all doses of dieldrin significantly

reduced E2 levels in females at Day 30, as compared to placebo, but not to sham,

however, this relationship is similar to the differences in hormone levels between

treatments at Day 0 (Figure 2-6). Regression analysis demonstrated a relationship of

depressed E2 concentrations in females at Day 30 only with increasing p,p '-DDE dose

(Figure 2-19). Day 60 data analyzed by ANOVA only shows a reduction in E2 at the

high p,p '-DDE and medium dieldrin doses as compared to the low p,p '-DDE dose and

not controls (Figure 2-6c). On the other hand, regression analysis showed a relationship

with decreased E2 concentrations only at the highest doses of both pesticides (Figure 2-

21). ANOVA analysis of male sex steroid levels also was hindered by the high

variability of sex steroid concentrations within a treatment group. No clear relationships

between dose and 11-KT concentration could be drawn. Means were clearly lower in the

dieldrin treatments, but were not significantly different from controls (Figure 2-9).









However, regression analysis demonstrates a significant negative relationship between

dieldrin dose and 11-KT concentrations. Also apparent from the ANOVA analysis was

that fish in some treatment groups had consistently higher/lower sex steroid

concentrations as compared to all other treatment groups. This may have been due to

differences in relative sexual maturity or size of the fish, however, there were no

differences among treatments in size (ANOVA and regression), regression of GSI against

sex steroid concentrations showed no relationship (Figure 2-21, 2-22, 2-24, 2-25), and

histologic analysis of the gonad at Day 60 showed that almost 90% of females were at the

same reproductive stage (Figure 2-4). The cause of this variability is unknown.

Analysis based only on the applied dose (implant concentration) could lead to

erroneous conclusions. The problem lies in that the "high" doses were not consistently

higher than the "low" or "medium" doses, especially in the case of dieldrin where

gonadal dose indicated the "high" treatment group received a lower dose than the "low"

treatment group. If a relationship between OCP dose and circulating E2 in females and

11-KT in males truly does exist, it would be impossible to determine from a "by

treatment" statistical analysis due to the variability in dose within a treatment group. Not

to be ignored, however, is the variability in hormone data within the control groups,

especially as hormone concentrations are at peak levels.

Like many biological endpoints, circulating hormone concentrations show a wide

inter-individual variability making statistical modeling difficult. The data sets of

hormone concentrations generated from this study do not appear to fit well to a linear

model. r2 values of statistically significant relationships ranged from 0.15 to 0.28

indicating the biological significance of these relationships is weak. Further statistical









analysis with a non-linear model is warranted. However, the results of this study suggest

thatp,p '-DDE and dieldrin may be able to depress sex steroid synthesis in largemouth

bass exposed to high doses.

In general, this study provides evidence thatp,p '-DDE and dieldrin are endocrine

modulators in the Florida largemouth bass at concentrations above what was found in

three year old female bass gonadal tissues collected from contaminated areas of the

Ocklawaha River Basin in 1996 (Marburger et al. 1999; 2002). Another study of OCP

concentrations in female largemouth bass ovaries was performed in 2002 and showed

mean ovarian concentration of total OCPs to be approximately 40,000 ppb, 10-fold

higher than previously reported, indicating concentrations achieved in the present study

were considerably low (Sepulveda et al., 2003). However, the two studies show

approximately the same level of dieldrin in bass ovaries. This conflicting data can most

likely be resolved by evaluating the time periods the studies were conducted in.

Marburger et al. (1999, 2002) collected samples in 1996 of fish that had been stocked in

1993, while Sepulveda et al. (2003) collected samples in 2002 from fish with a mean age

of six years. The difference in exposure time would allow for the bioaccumulation of

larger concentrations of OCPs over time.

In this study, neither chemical induced a change in any health parameter (weight,

length, K, and HSI), indicating that acutely toxic doses were not achieved. Of note is that

circulating levels of E2 and 11-KT were approximately 2-fold less than reported for the

same age, pond-reared fish in the Gross et al. (2000) study and was most likely due to

stress induced by captive holding conditions (i.e. unnatural housing and high stocking

density). Fish were stocked at a density of 1 bass per 61 liters of water. At this density,









largemouth bass remain healthy but will not spawn. For reference, spawning enclosures

built on the edge of a pond at the USGS-CARS facility are stocked at an approximate

density of 1 fish per 2500 liters. The ratio of E2 to 11-KT, however, was within normal

ranges, greater than 1 for females and less than 1 for males, indicating that these fish

were sexually mature (Marburger et al., 1999). The average size of the fish used in this

study indicates they were approximately 1-2 years of age and most likely going through

their first reproductive cycle. This may have increased the variability in sex steroid

concentrations.

Overall, increasing p,p '-DDE concentration over time caused a decrease in

circulating E2 levels in females, with no other health or reproductive effects and was not

correlated with relative gonad size. p,p '-DDE concentrations reached by Day 30 of

exposure were enough to stifle the seasonal increase in E2, the magnitude of which

increased with increasing dose as shown by the relationship ofp,p '-DDE dose to the

percent change in E2 concentration from Days 0 to 30. Still demonstrating a dose-

dependant depression at Day 60, E2 levels in females was not correlated with GSI. This

is most likely because the gonad begins to mature in October and reaches a considerable

size (approximately 5% body weight) by January when this study was initiated (Gross et

al., 2000). Peak GSI of about 6% occurs is in February. A relevant dose ofp,p '-DDE

was most likely not achieved by the time peak GSI was reached and therefore could have

little to no effect on this parameter. A study of male guppies (Poecilia reticulata)

demonstrated a decrease in GSI when fed a total of 15 and 150 tg ofp,p '-DDE over 30

days; the guppies fed 1.5 tg showed no change in GSI (Baatrup and Junge, 2001),

however, no contaminant analysis was done in this study so attained dose in unknown.









p,p '-DDE also had no effect on circulating 11-KT levels in males or females in this study

despite its anti-androgenic properties in mammals (Kelce et al., 1995) and its ability to

bind the AR of some fish species (Wells and Van Der Kraak, 2000). It is possible that

p,p '-DDE may not bind the AR in largemouth bass. If it was a competitive inhibitor,

testosterone levels would be expected to decrease providing less substrate for 11-KT

synthesis and therefore a dose dependent decrease in circulating 11-KT. In comparison,

low p,p '-DDE doses (liver concentration 400 ppb, 2 to 4 times less than what was

achieved in this study) in male summer flounder (Pa, ili, hi.,\ dentatus) injected IP

showed no change in GSI or hormone concentrations (Mills et al., 2001). Conversely,

white sturgeon (Acipenser transmontanus) from the Columbia River with liverp,p '-DDE

concentrations over 700 ppb had severely depressed testosterone and 11-KT

concentrations (Foster et al., 2001). Similarily, largemouth bass collected from the

Emeralda Marsh and the St. John's River had depressed sex steroid concentrations (E2 in

females and 11-KT in males), as well as, decreased GSI (Marburger et al., 1999;

Sepulveda et al., 2002). These fish however, were exposed to multiple pesticides and

other chemicals and any mixture effects are unknown. Difference between species most

likely contributes to apparently varying effects ofp,p '-DDE.

Dieldrin had no statistically significant effects on circulating sex steroid levels in

either sex at Day 30 of exposure. However, the percent increase in 11-KT concentration

from Day 0 to Day 30 was significantly less in dieldrin-treated versus untreated males,

and although not statistically significant, 11-KT levels tended to decrease with increasing

doses of dieldrin. By Day 60, females and males treated with dieldrin had lower

circulating E2 and 11-KT, respectively, than controls. GSI was unaffected, probably due









to the timing of the exposure, and was not correlated with 11-KT concentrations. No

studies of in vivo dieldrin-only exposure in fish were found to compare the results to.

However the studies mentioned above from the Emeralda Marsh and the St. John's River

also demonstrated elevated levels of dieldrin in largemouth bass tissue and depression in

sex steroid hormones (Marburger et al., 1999; Sepulveda et al., 2002). This may be due,

in part, to the effects of dieldrin.

The mechanisms of action forp,p '-DDE and dieldrin as endocrine modulators in

fish are unknown and many possibilities exist. First, either of these chemicals could bind

to the plasma sex steroid binding protein (SSBP) and disturb endogenous hormone

binding making it available in the blood for biotransformation and elimination. This is a

more likely mechanism for dieldrin because its effects include decreases in both

hormones assayed. However, two studies on rainbow trout blood plasma have shown

little to no ability for dieldrin or DDT derivatives to bind the SSBP effectively (Milligan

et al., 1998; Tollefson, 2002b).

Second, either pesticide may be able to inhibit aromatase activity. Aromatase in

fish converts testosterone to E2 and 11-KT. Inhibition of this enzyme would lead to a

decrease in E2 in females and 11-KT in males, as is the case for dieldrin. Evidence of this

mechanism has been demonstrated in alligators in OCP contaminated areas of Lake

Apopka (Crain et al., 1997). However, p,p '-DDE exposure in a male rat study

demonstrated increased aromatase activity (You et al., 2001).

Third, these OCPs may be able to induce enzymes responsible for

biotransformation of sex steroids thereby eliminating them at a faster rate than they are

produced. This mechanism stands alone in that it is not directly involved in the endocrine









system and demonstrates a classic toxic action of inducing catabolism. The catabolic

pathway for sex steroids has not been fully characterized in fish but is believed to include

the cytochrome P450 isoform 3A (Stegman and Hahn, 1994). Studies on dieldrin in

human hepatoma cells (Coumoul et al., 2002) and rats (You et al., 1999) have shown that

increasing OCP load can induce CYP3A. However, species differences make

extrapolation of this data to fish uncertain. A marker of liver function and CYP induction

in general, EROD, has shown increased activity with increasing OCP load and related

decreases in circulating sex steroids in several studies (Dickerson et al., 1999; Zapata-

Perez et al., 2000; Foster et al., 2001; Wade et al., 2002), however, a relation of EROD to

sex steroid metabolism has not been established.

Lastly, p,p '-DDE and dieldrin could disrupt normal endocrine function in fish by

interfering with the feedback pathways of the hypothalamus-pituitary-gonad axis. p,p '-

DDE and dieldrin have both demonstrated the ability to bind estrogen and androgen

receptors (Kelce et al., 1995; Danzo, 1997; Nimrod and Benson, 1997; Matthews et al.,

2000; Wells and Van Der Kraak, 2000; Anderson et al., 2002) and could therefore bind to

these receptors anywhere in the feedback pathway. Binding of high concentrations at the

hypothalamus and pituitary would inhibit further production of GnRH and GTH,

respectively, effectively shutting down the signal to the gonad for sex steroid synthesis

with some temporality between exposure and effect. This is the mechanism by which the

normal annual cycling of hormones occurs (Janz and Weber, 2000). Low sex steroid

levels serve as a stimulus for further sex steroid production (positive feedback) and is sex

specific for which hormone is upregulated. However, when plasma hormone

concentrations reach a threshold concentration, the signal for steroid synthesis is shut off









(negative feedback) and hormone levels slowly return to baseline. If these pesticides are

able to mimic E2 and/or 11-KT action, then it is possible they could bind to receptors in

the brain and induce the hypothalamus or pituitary to shut down sex steroid synthesis. A

human hypothalamic cell line study examined the effects of OCPs (methoxychlor and

chlorpyrifos) on this pathway and found low levels of exposure stimulates GnRH release,

theoretically by stimulating sex steroid receptors initiating positive feedback (Gore,

2002). However, Van Der Kraak et al. (1992) demonstrated decreased GTH in white

sucker fish (Catostomus commersoni) exposed to bleached kraft pulp mill effluent

(BKME). Although BKME does not contain OCPs it does contain some chlorinated

chemicals that are thought to be endocrine active. It is clear that more research is needed

to elucidate the mechanism by whichp,p '-DDE and dieldrin exposure leads to decreased

plasma sex steroid concentrations.

This study did not include any endpoints that would indicate the functionality of

p,p '-DDE or dieldrin as enzyme inducers/inhibitors or estrogen/androgen mimics.

Aromatase, CYP activity, circulating vitellogenin, GnRH, and GTH could serve as

indicators of the biochemical actions of these pesticides. In order to draw accurate

conclusions about how OCPs cause a depression in sex steroid levels, it is imperative to

understand by which mechanisms they are acting. However, this study did not aim to

elucidate the mechanism of action for these chemicals. It serves only as evidence that

p,p '-DDE and dieldrin exposure at environmentally relevant levels cause a decrease in

circulating sex steroids in largemouth bass. This study also addressed the accuracy of

dosing fish by time-release pellets and characterized relevant endpoints of exposure and

their dose response top,p '-DDE and dieldrin. The results show that two of the






35


predominant chemicals found in the contaminated areas of the Ocklawaha River Basin do

contribute to the depressed sex steroid levels and altered reproductive success of

largemouth bass exposed in the wild. The data presented here will be used to design

future studies aimed at determining the effects of these pesticides in binary, ternary, and

complex mixtures, similar to what is found in the environment.












300-, a


S200-
100-


100-



0-


Day 0


Day 30


Day 60


300-, b


E
E 200-


S-
C0
. 100-


Day 0


Day 30


Day 60


3-C


0



0
'2-





0
L-
0


Day 0 Day 30 Day 60
Figure 2-1. Female body condition over time. Box plots of weight (a), length (b), and
condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 63). Box
plot contains the 25th to the 75th quartile, line in box indicates the median,
whiskers extend to the minimum and maximum value.










300- a


S200-


100-
~ o-


Day 0


Day 30


Day 60


300- b


E
E 200-


100-
. 100-


Day 0


Day 30


Day 60


3- C


L




C 1-
0

2-
0
0


U'
Day 0 Day 30 Day 60
Figure 2-2. Male body condition over time. Box plots of weight (a), length (b), and
condition factor (c) for all fish in the study at 0, 30, and 60 days (n = 93). Box
plot contains the 25th to the 75th quartile, line in box indicates the median,
whiskers extend to the minimum and maximum value.


~-









4-
a a a


2-
1

o,. II I ,o ",



5b b
S4- a b
S a,b a,b a,b ab a,b


2-il i


<(>a~ .-~P 3 ,0 -0 (, e,

Figure 2-3. Day 60 HSI. There was no significant difference in HSI among treatment
group for females (a). Males in the Dieldrin 1.0 mg treatment group showed a
significant increase in HSI over Sham and DDE 2.5 mg treatments (b).
Sample size per treatment ranged from 7 to 11 for females and 7 to 16 for
males. Treatments with the same lower case letter were not significantly
different.















CA


GV












CA


Figure 2-4. Histological section of stage 4 vitellogenic female gonad viewed at
= cortical alveoli, GV = germinal vesicle, YV = yolk vesicle.















7.5-


S 5.0-

2.5
LL 2.5-


~~oO <:' < <
CV 6. <;.ot~O

9\ a\ .a.


1.0-



0.5-


Ic~oo ~O :, 'o o G, c
O'QP (P N Z'
CV o. <;\
,e .


Figure 2-5. Day 60 GSI. There was no significant difference in GSI among treatment
group for females (a) or males (b). Sample size per treatment ranged from 7
to 11 for females and 7 to 16 for males. Treatments with the same lower case
letter were not significantly different.












1500 a


1000- ab a ,b
Sb b b b
E I
500-









1500oo b
a

21000- b
2b b
b b b

500-









1500- C a

a,b

ab ab


I e
500








Figure 2-6. Female circulating estradiol at Days 0 (a), 30 (b), and 60 (c). Placebo and
DDE 2.5 treatments showed consistently higher plasma E2 concentrations than
all other treatments. Sample size per treatment ranged from 7 to 11.
Treatments with the same lower case letter were not significantly different.












1000- a


S750-

.a,b
.2 5oo-
Va

w 250-









1000- b


E 750-




S250










a,b
E 750-
a,b

500- a,b a,b a,b


W. 250







Figure 2-7. Male circulating estradiol at Days 0 (a), 30 (b), and 60 (c). DDE 5.0 mg
treatment showed a consistently higher mean than all other treatments.
Sample size per treatment ranged from 7 to 16. Treatments with the same
lower case letter were not significantly different.













1000-
0 b
o
S 750- a,b b

O-


U







1 a
00








1000 b

0
750-
a a,b
w ) 500-
a a

250-









1000-
o
750-

0 500 a a a
0) 500- a
a a
0a
250-







Figure 2-8. Female circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Treatment
groups demonstrated high variability compared to controls. By Day 60, no
difference existed between treatment groups and controls. Sample size per
treatment ranged from 7 to 11. Treatments with the same lower case letter
were not significantly different.













151

0
-a,
C
0


a5
i 101

a,1
0-


a)
0

#-A

0
o-
0 a

(U
^:


IP QO 3 I
2 o: o~~
~"~ ~ 0n~`


C
0


aC

i-
4 0.
0-

a,


a, b


I a a a--
M a


,... o? o<. o.0 0 .



Figure 2-9. Male circulating 11-KT at Days 0 (a), 30 (b), and 60 (c). Placebo and DDE
2.5 mg treatment groups demonstrated consistently higher 11-KT. At Day 30
and 60 all dieldrin treatments and higher DDE treatments demonstrated
decreased 11-KT. Sample size per treatment ranged from 7 to 16. Treatments
with the same lower case letter were not significantly different.


0 b C3
CP~,O












DDE
Target
dose


I ICP
Kbt
SC3


Treatment


- Dieldrin
SdTarget
dose


s
Q) 5


Treatment

Figure 2-10. Day 60 mean female gonadal dose. p,p '-DDE (a) and dieldrin (b)
concentration + SD for each treatment group (n = 6) compared to the target
dose for each treatment. Also included is the mean gonadal concentration of
either OCP found in the Eustis property of EMCA.


,0o


4000- b


3000
Q.
Q.

* 2000


1000


IQoo
Z Q5.










Table 2-1. GC-MS results of tissue samples for bioaccumulation measurements at Day
30 of exposure. Blue and green numbers highlight fish treated withp,p '-DDE
and dieldrin, respectively.
a


Treatment Sex
Sham M
F
Placebo M
F
DDE 2.5 F
F
DDE 5.0 F
F
DDE 10.0 M
M
M
Dieldrin 0.25 M
F
Dieldrin 0.5 F
M
Dieldrin 1.0 M
M


DDE
gonad
10.00
18.60
23.59
20.62
7808.63
4960.30
2475.52
10813.95
5108.70
2953.26
700.93
365.85
1521.21
36.20
1024.94
182.08
49.56


DDE DDE
blood muscle
1.32
14.49
5.77 1.31
2.5 1.47
907.83 113.34
481.58 204.42
116.67 92.48
1006.45 348.37
316.07 318.06
390.70 111.61
278.57 48.47
101.56 19.88
171.95 26.28
7.15
270.59
6.38 1.40
13.24 5.23


Avg.
Gonad DDE
14.30

22.105


Std. Dev.


6.08

2.10


6384.47 2014.07

6644.74 737.67

2920.96 2204.06


Dieldrin Dieldrin Dieldrin Dieldrin Avg. Gonad
Treatment Sex Liver gonad blood muscle Dieldrin


Sham

Placebo

DDE 2.5

DDE 5.0

DDE 10.0


Dieldrin 0.25

Dieldrin 0.5

Dieldrin 1.0


11.03
13.23
11.4
14.95
13.11
8.73
55.20
129.19
11.63
13.44
8.64
143.81
223.92
153.51
748.35
127.34
114.48


20.00
34.33
31.45
36.82
32.72
38.02
567.83
1453.49
50.72
29.75
24.03
515.12
2442.71
1245.17
1249.38
1443.64
747.14


55.56
193.55



196.88
384.15


172.34
282.35


3.95
2.90
2.62
2.94
1.37
2.21
33.29
53.60
5.56
3.35
1.35
37.28
57.81
230.04
519.10
54.72
92.11


27.17 10.13


34.135


3.80


1478.92 1363.01


1247.28


2.98


1095.39 492.50


DDE
Liver
5.51
4.41
4.27
4.48
1733.85
349.01
198.41
1241.63
363.37
642.11
280.84
95.40
121.60
7.31
558.37
6.85
5.87


Std. Dev.


















15000- a


0.
5000-
"o

0
W 2500-
0


r2=0.9520


0 250 500 750 1000
DDE blood (ppb)


2
r =0.7520


0 100 200 300 400 50
DDE blood (ppb)


750- e


0 500-



o 250-
Q


r =0.8345


0 250 500 750 1000 1250
DDE blood (ppb)


400, C


S r=0.8807


100 200 300 400
DDE blood (ppb)


400, f


300-


S200-
E

o 100-
0


2

r 2=0.5858

I 250 500 750 1000
DDE blood (ppb)


r2=0.4041


U- I I I
0 100 200 300 400 50
DDE blood (ppb)


Figure 2-11. Regression analysis of gonad (a, d), liver (b, e) and muscle (c, f) DDE

concentrations against blood DDE concentrations in females (a c; n = 9) and

males (d f; n = 6). Significant linear relationships exist between blood

plasma and tissue DDE concentrations, except in male muscle.


"0

0)
W 5000-
1-


2000- b


&


> 1000-
W
0
0


300-


= 200-
E

o 100-
0


7500, d
















3000- a


Q.

2 2000-

0

I 1000-
2
2


r2=0.9262



100 200 300 400 500 600
Dieldrin blood (ppb)


400- b


0 300-


- 200


100-


n r


' 1 '0
r2=0.6824



250 500 750 1000 1250
Dieldrin blood (ppb)


500- e


S.400-

S300-

S200-

o 100-


r2=0.9986


1 100 200 300 400 500 600
Dieldrin blood (ppb)


.* r2=0.9650




250 500 750 1000 1250
Dieldrin blood (ppb)


200- C


r2=0.8527


200- f



U
100-

S100
c

Q: a,


U 100 200 300 40UU bU 600
Dieldrin blood (ppb)


r2=0.9120


U 2b0 bU00 b 1000
Dieldrin blood (ppb)


Figure 2-12. Regression analysis of gonad (a, d), liver (b, e), and muscle (c, f) dieldrin

concentrations against blood dieldrin concentrations in females (a c; n = 4)

and males (d f; n = 6). Significant linear relationships exist between blood

plasma and tissue dieldrin concentrations, except in female muscle.


-.

0 2000-
0

S1000-
Q


C.
Q.

U

E 100-
c


Q


12b5


3000, d
















( 1000-
LU
LU
500-
o 0-


0 250 500 750 1000 1250 1500
DDE GC-MS


750




C
2 500



S250


0 100 200 300 400 500 600
Dieldrin GC-MS

Figure 2-13. Linear regression of ELISA DDE (a) or dieldrin (b) results of pooled blood
samples against GC-MS results (n = 12). A highly significant and linear
relationship exists between the two methods.
















r2=0.0001


*


.


* U.- U


3-

4-
) 3-


E
2-

1-

0


U


* *
U


1000


r2=0.0026


0 1000 1500

Plasma DDE (ppb)


2000
2000


Figure 2-14. Linear regression of HSI against blood plasma DDE concentrations. No
significant correlation was found for females (a, n = 37) or males (b, n = 53).


250 500 7!

Plasma DDE (ppb)


6

5-

4-

3-1

2-


1-

0-


I


















4-
I
G 3-




1-


0-


r =0.0005

iU
U*
U *
I
ii U
*


Plasma Dieldrin (ppb)


U
I


r2=0.0815


0 100 200 300 400 500


Plasma Dieldrin (ppb)


Figure 2-15. Linear regression of HSI against blood plasma dieldrin concentration. No
significant correlation was found for females (a, n = 34) or males (b, n = 51).











10.0- a


7.5- r2=0.0071


5.0-


2.5


1000


Plasma DDE (ppb)


1.5- b


r2=0.0273
1.0-



0.5-
m.' ." U


1000


1500


2000


Plasma DDE (ppb)

Figure 2-16. Linear regression of GSI against blood plasma DDE concentration. No
significant correlation was found for females (a, n = 37) or males (b, n = 53).














r2=0.0028


-F


U U
U U


100 200

Plasma Dieldrin


300

(ppb)


400


r2=0.0024


i*
m U U


0 100 200 300 400 500

Plasma Dieldrin (ppb)


Figure 2-17. Linear regression of GSI against blood plasma dieldrin concentration. No
significant correlation was found for females (a, n = 34) or males (b, n = 51).


10.0- a














O


c-1
CL



0
E

I-
0)


1000-


750-


500-


250-


Estradiol


11-KT


c 1500- b
0
O


C
^ 1000-




S500-
E
0
I


UEstr
Estradiol


11-KT


Figure 2-18. Day 0 plasma hormone concentrations. Mean + SD (a) female, n = 59 (b)
male, n = 88.














SE2
11-KT


r2=0.2815


0 100 200 300

Blood Plasma DDE (ppb)


. E2
11-KT


Ip AA
SA A
. I


I *
---- t _____ i______A

^T,^--0


r= 0.0000


r2=0.1401

400


Blood Plasma DDE (ppb)



Figure 2-19. Day 30 Hormones Femalep,p '-DDE treated (n = 30). Linear regression
against DDE dose (a) circulating hormones (b) percent change in hormone
concentration from Day 0 to Day 30. Significant negative relationships exist
between DDE dose and circulating E2 and percent change in E2.


C
0
.1





) --
C

0
E
0
I


300-


200-


100-


-100-













a
c 2000-
0


.0--
o



C
O



E
0


* E2

A 11-KT


,A r =0.0468
A.
A A
A A A -A


LA

A


A A


-, A
EU


S r2=0.0008


Blood Plasma DDE (ppb)


. E2
11-KT
A 11-KT


S r2=0.0114


i t. r2=0.0329
U


Blood Plasma DDE (ppb)

Figure 2-20. Day 30 Hormones Malep,p '-DDE treated (n = 47). Linear regression of
(a) circulating hormones (b) percent change in hormone concentration from
Day 0 to day 30 against DDE dose. No significant relationships were found.


400-


300-


%*'*-*







57




C 1500 a .
C.2
S11-KT

< O 1000- .
E
Sr=0.1588
0 500- *
0 2 2=
E J_ --- --- --- ---.
-I- r=0.0240

0 250 500 750 1000
Blood Plasma DDE (ppb)


200- b
E2
11-KT
A *
0 100-

U r2=0.0128


i- '. r2=0.0482

-100
0 250 500 750 1000
Blood Plasma DDE (ppb)


1500- C








E
1000-



) 500- *

.. r2=0.0791

0.0 2.5 5.0 7.5 10.0
Female GSI

Figure 2-21. Day 60 Hormones Female p,p '-DDE treated (n = 35). Linear regression
of (a) circulating hormones (b) percent change in hormone concentration from
Day 30 to Day 60 againstp,p '-DDE dose and (c) ciculating E2 to GSI. A
significant negative relationship was found between DDE dose and circulating
E2.







58




S 2000- a
E2
c c 11-KT




E


0


0 250 500 750 1000 1250 1500
Blood Plasma DDE (ppb)
O bE2
o r""0.0116

0 250 500 750 1000 1250 1500
Blood Plasma DDE (ppb)


1000- b 2

A 11-KT

S500-
C A
Ac r=0;0526

0

Sr =0.0145., A
AA A A
-500 I I
0 250 500 750 1000 1250 1500
Blood Plasma DDE (ppb)


2000- C



I-P
S1000-

2 *. *
.=
r2=0.0043
0
0.00 0.25 0.50 0.75 1.00 1.25
Male GSI

Figure 2-22. Day 60 Hormones Malep,p '-DDE treated (n = 51). Linear regression of
(a) circulating hormones (b) percent change in hormone concentration from
Day 30 to Day 60 against dieldrin dose and (c) circulating 11-KT against GSI.
No significant relationships were found.













c 1500-
0


C





E
I- 1000-





0
"1


400-

300-

200-

100-


*E2
S11-KT


A r 2=0.0213

- --, r =0.0367
A A A
A A A A
50 100 150 20(

Blood Plasma Dieldrin (ppb)


* E2
A 11-KT


AA


A. 2


2
I ". r2=0.0097

! r2=0.0148
k i .
S* A* I A A


IUU-- I
0 50 100 150 200

Blood Plasma Dieldrin (ppb)

Figure 2-23. Day 30 Hormones Female dieldrin treated (n = 34). Linear regression of
(a) circulating hormones (b) percent change in hormone concentration from
Day 0 to Day 30 against dieldrin dose. No significant relationships were
found.












c 1500-



0.-. 1000-
o
C,
I 5 0 0 -
0
E


200-


E2
11-KT


A

A 2
A r 2= 0.0898
.089A
AN V I .


L I


A A Ar2 A0b
r-I0 063 I


0 50 100 150 2C
Blood Plasma Dieldrin (ppb)


E E2
11-KT


A *

S* r2=0.0137



100
0 50 100 150 200

Blood Plasma Dieldrin (ppb)

Figure 2-24. Day 30 Hormones Male dieldrin treated (n = 42). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day
0 to Day 30 against dieldrin dose. A significant negative relationship exists
between dieldrin dose and percent change in 11-KT from Day 0 to Day 30.














* E2
* 11-KT


*



i 0.0003
r ,00 2.1597


A' r =0.0003


0 100 200 300 400
Blood Plasma Dieldrin


500 600

(ppb)


E2
E2
A 11-KT



r2=0.0001


0 ~
.* r2=0.0007
-100-

-200 i
0 100 200 300 400 500 600
Blood Plasma Dieldrin (ppb)


- 1500- C



" 1000-

LU
S500-

E
U-


r2=0.1110


0.0 2.5 5.0
Female GSI


7.5 10.0


Figure 2-25. Day 60 Hormones Female dieldrin treated (n= 34). Linear regression of
(a) circulating hormones (b) percent change in hormone concentration from
Day 30 to Day 60 against blood plasma dieldrin and (c) circulating E2 against
GSI. A significant negative relationship between blood plasma dieldrin and
circulating E2 was found.


1500a
1500-i


- 1000-

e


() I -


300- b


200-

100-
i A
II



IZI


*


*







62



a
1500- a
o E2
S11-KT

1000-
CL ,


a 500- 2
r =0.2212
0 --
S:. r =0.0818
0 100 200 300 400 500 600 700
Blood Plasma Dieldrin (ppb)


200- b
A 11-KT


0 100- *

o2
Sr. =0.0096


r2=0.0675
100
0 100 200 300 400 500 600 700
Blood Plasma Dieldrin (ppb)


1500- C


1000
I- .


S500-
.-* r2=0.0030

0-- ------ i ------ i ---
0.00 0.25 0.50 0.75 1.00 1.25
Male GSI

Figure 2-26. Day 60 Hormones Male dieldrin treated (n = 51). Linear regression of (a)
circulating hormones (b) percent change in hormone concentration from Day
0 to Day 30 against dieldrin dose and (c) circulating 11-KT against GSI.
Significant negative relationships between circulating 11-KT and dieldrin
dose, as well as, between percent change in E2 from Day 30 to Day 60 were
found.













CHAPTER 3
ACCUMULATION OF DIETARY p,p '-DDE AND DIELDRIN BY LARGEMOUTH
BASS: A PILOT STUDY

Introduction

Large variability in the amount of chemical released from pellets placed

intraperitoneally (see Chapter 2) into largemouth bass led to the need for developing and

validating a new exposure method. The primary exposure route for fish to persistent

pesticides is accumulation through the food web (Woodwell et al., 1967; Macek and

Korn, 1970; Ruus et al., 2002). This research describes a pilot study aimed to determine

the accumulation rate ofp,p '-DDE and dieldrin in largemouth bass fed a contaminated

diet.

A review of the literature provided varying degrees of dietary accumulation of

either pesticide based on species and dose range tested. Previous studies on pesticide

uptake were performed by mixing a pesticide with an oil carrier and coating the feed

pellets with the oil solution. Fish were fed at a range of 0.5 4.0% body weight per day,

depending on species. No studies were found that addressed p,p '-DDE specifically, but

several evaluated DDT accumulation. When fed a DDT contaminated diet, rainbow trout

accumulated 20 24% of the DDT dose available over 140 days (Macek et al., 1970);

brook trout (Salvelinusfontinalis) accumulated 35% over 120 days (Macek and Korn,

1970); and Atlantic menhaden (Brevoortia tyrannus) accumulated 17 27% over 48 days

(Warlen et al., 1977). Dietary dieldrin accumulation was around 10% for rainbow trout

over 140 days (Macek et al., 1970) and striped bass, Morone saxatilis, over 84 days

(Santerre et al., 1997).









Another point of investigation in this study was the variation in accumulation rates

between two types of feed: sinking and floating. Application of pesticide to floating feed

is done by mixing the chemical with an oil carrier and using the oil as a top dressing on

the feed pellets. Sinking feed is manufactured by grinding up the feed pellets and mixing

the chemicals directly in with the pellet powder, then reconstituting the powder back into

pellets. The sinking pellets, presumably, have less pellet-to-pellet variation in achieved

chemical dose, however largemouth bass are usually top feeders and it was unclear if

they would consistently feed on sinking pellets.

Validation of a dietary exposure method was intended to provide useful

information on the accumulation rates ofp,p '-DDE and dieldrin, variability inherent to

the method, and difference between pellet type. This study was also intended to

determine uptake and elimination rates of these OCPs over time.

Specific Aims

* To determine the bioaccumulation rate and final whole body concentration ofp,p '-
DDE and dieldrin in largemouth bass when exposed orally for 30 and 50 days by
contaminated feed.

* To determine if an internal steady state dose is achieved after 30 and 50 days.

* To determine the variation in achieved whole body concentration for largemouth
bass, both within a tank and among replicates, fed different feed types: sinking or
floating.

* To determine if either chemical induced a change in circulating sex steroid
hormones at the achieved doses.

Null Hypothesis

Accumulation rates and circulating sex steroid levels will not vary between

individual fish or feed type.









Alternative Hypothesis

Because largemouth bass within each replicate are of varying size and have varying

degrees of aggressive food seeking behavior, accumulation rates will vary between

individual fish, treatment replicate, and feed type. Feeding aggression may be an

important factor for sinking feed as largemouth bass typically will not feed off the

bottom, introducing a greater possibility for variability in accumulation rates among

individual fish fed sinking feed. In addition, exposure of bass top,p '-DDE and dieldrin,

potential endocrine modulators, may decrease the concentration of circulating sex

steroids in treated versus control fish.

Materials and Methods

Experimental Animals

Largemouth bass, one to two years of age, were obtained from a fish hatchery

(American Sports Fish, AL) in November 2002. The fish were transferred to the United

States Geological Survey Biological Resource Division Center for Aquatic Resource

Studies (USGS-BRD-CARS) facility where they were weighed and measured, then

housed in groups of 14 fish in 700 L round tanks equipped with a flow-through system

supplied by on-site well water and aeration. On Day 0 of the experiment (December 5,

2002) bass had an average weight, length, and condition factor SD of 181.6 34.2 g,

233.2 13.3 mm, and 1.42 0.10, respectively, indicating fish were healthy and of

reproductive size.

Water quality was measured as dissolved oxygen, ammonia, temperature and pH

twice a week at a depth of approximately 15 cm, early in the morning (approximately 8 -

9 a.m.) when the dissolved oxygen is at its lowest. Water quality parameters were all

within acceptable ranges for the duration of the experiment: dissolved oxygen ranged









from 7.49 to 9.35 mg/L, ammonia content stayed below 1 ppm, temperature ranged from

17.4 to 20.8 C, and pH ranged from 7.8 to 8.1.

Chemicals and Feed

The organochlorine pesticides p,p '-DDE (2,2-bis(4-chlorophenyl)-1,1-

dichloroethylene, Lot # 09020KU, 99.4% purity) and dieldrin (1,2,3,4,10,10-hexachloro-

6,7-epoxy-l,4,4a,5,6,7,8,8a-octahydro-l,4,5,8-dimethanonaphthalene, Lot # 077H3578,

91.2% purity) were obtained from Aldrich Chemical Company (Milwaukee, WI). The

pesticides were shipped to Zeigler Brothers, Inc. (Gardners, PA) where they were

incorporated into feed pellets, sinking and floating for both chemicals, at concentrations

of 5 ppmp,p '-DDE and 1 ppm dieldrin. Contaminated sinking pellets were

manufactured by incorporating a measured amount of chemical into sinking pellets that

had been ground to a powder, then mixed thoroughly. The feed powder was then

reconstituted into pellets. Contaminated floating feed was manufactured by incorporating

a measured amount of chemical into a fish oil mixture that was used to top dress the

pellets. Oil mixture and pellets were mixed thoroughly in a mixer to achieve consistent

coating of all pellets. A control was also manufactured for each type of feed. The oil

control received a top dressing of pure fish oil and the sinking feed was ground and

reconstituted. Each chemical and feed type was tested in duplicate (Figure 3-1).

One sample of each control type and duplicate samples of all treated feed types

were sent to New Jersey Feed Laboratory, Inc. (Trenton, NJ) by Zeigler Feed, Inc. for a

chlorinated pesticide and PCB (polychlorinated biphenyl) screen analysis. All feed had a

background level of aldrin at 0.03 ppm. Sinking and floating controls had no other

detectable OCPs or PCBs. Sinkingp,p '-DDE (target 5 ppm) and dieldrin (target 1 ppm)

feeds had actual mean SD concentrations of 3.6 0.7 and 0.89 0.25, respectively.









Floating p,p '-DDE (target 5 ppm) and dieldrin (target 1 ppm) feeds had actual

concentrations of 3.9 0.3 and 1.07 0.05, respectively.

Although sinking feed was designed to deliver a more accurate and less variable

dose, it actually was farther from the target dose and more variable per pellet than

floating feed.

Feeding Rate

The average weight of largemouth bass in this study was used to determine the

feeding rate. Contaminated or control feed was administered to each tank at 5% mean

body weight for fish (14) within a tank per day (recommended rate from Rick Stout,

FFWCC-Fish Hatchery, Richloam, FL).

Day 30 and Day 50 Endpoints

Weight, to the nearest gram, and length in millimeters, were measured; blood was

collected; and gonads were removed and weighed from a subset of 3 fish per tank on Day

30 and all remaining fish (n = 11) on Day 50. Condition factor was calculated (K =

weight/length3 x 100,000). Approximately 1 ml of blood was obtained from the caudal

vein using a heparinized 20-gauge needle, dispensed into a heparinized vacutainer,

labeled and stored on ice until centrifuged. Blood samples were centrifuged at

approximately 1,000x g, 4C for 20 min to separate red blood cells from plasma. Plasma

was removed with a transfer pipette, placed in a cryovial and stored at -80C until

assayed for circulating sex steroids. Each fish carcass (and gonad for Day 50 fish) was

wrapped in aluminum foil, placed in a plastic bag and labeled. Carcass (n = 3 for both

Day 30 and 50) and gonads (n = 3, Day 50 only) were sent for GC-MS analysis ofp,p '-

DDE and dieldrin.









Analysis of Largemouth Bass Tissue for OCPs

Analysis was conducted by the Center for Environmental and Human Toxicology,

University of Florida. Briefly, the largemouth bass carcass/tissue was homogenized and

a portion of the sample (2-5 grams) was extracted into ethyl acetate. The sample was

cleaned-up using C18 and NH2 SPE (solid phase extraction) cartridges. Total OCP

content was determined by GC-MS, following EPA method 8270 (EPA, 1983). Samples

were analyzed multiple times in full scan mode for analyte identification and in selected

ion mode (SIM) for quantitation to improve sensitivity. Percent recovery ranged between

65 and 100% with a limit of detection of 0.75-1.5 ng/g.

Analysis of Circulating Sex Steroid Hormones

Blood plasma from largemouth bass was analyzed for 11-KT and E2 using a

previously validated 3H radioimmunoassay (RIA) method (Gross et al., 2000). All

samples were assayed in duplicate and values reported as pg/ml of plasma. Standard

curves were prepared in phosphate buffered saline plus gelatin and sodium azide

(PBSGA) buffer with known amounts (15, 30, 60, 125, 250, 500, 1000, and 2000 pg) of

radioinert E2 (ICN Biomedicals, Costa Mesa, CA) or 11-KT (Sigma Chemicals, St.

Louis, MO) and 15,000 cpm of 3H-E2 or 3H-11-KT. Each plasma sample (50 l1) was

extracted twice with diethyl ether prior to RIA analysis. PBSGA buffer and antibodies

against sex steroid hormones were added to the sample tube and incubated overnight at

4C. Antibodies were purchased from ICN Biomedicals (E2) or Helix Biotech,

Richmond, BC, Canada (11-KT). Following the antibody incubation unbound antibody

was separated out by addition of dextran-coated charcoal and centrifugation for 10

minutes at 1,000 x g. Binding to the charcoal in the pellet stripped excess free antibody

out of solution. Four hundred [l of sample supernatant was removed and added to a









scintillation vial with 4 ml of Scintiverse scintillation cocktail (Fisher Scientific,

Pittsburgh, PA). Samples vials were then placed in a liquid scintillation counter (Pachard

Tricarb, Model 1600) and counted for two minutes each. The minimum concentration

distinguishable from zero (mean SE) 148 + 18 pg/ml for E2 and 29 18 pg/ml for 11-

KT. Cross-reactivities (produced and characterized by T.S. Gross, University of Florida)

of the E2 antiserum with other steroids were: 11.2% for estrone, 1.7% for estriol, and <

1.0% for 17a-estradiol and androstenedione. Cross reactivity of the 11-KT antiserum

with other steroids was: 9.7% for testosterone, 3.7% for a-dihydrotestosterone, and <

1.0% for androstenedione.

Statistical Analysis

All parameters were analyzed using the Statistical Analysis System (SAS),

version 9. All data were run first by the Means Procedure to determine averages and

standard deviations for all treatment groups and replicates. All data presented in the

results are means SD. To determine any statistical significance between replicates,

feed type, and chemical, an analysis of variance (ANOVA) procedure, followed by the

multiple comparisons procedure, Duncan's Multiple Range Test, for all parameters (SAS)

to determine differences between replicates. Significance was declared at a = 0.05.

Results

Largemouth Bass

There was no mortality of largemouth bass throughout the duration of the study.

The condition factor at Days 30 and 50 were 1.45 + 0.09 and 1.42 + 0.10, respectively,

and did not differ between feed types (sinking or floating) or pesticide administered (p,p '-

DDE or dieldrin). This indicates a healthy population was maintained throughout the









exposure period. Largemouth bass grew an average of 3 0.4 mm and gained an average

of 2 0.8 g over the course of the study.

Feeding Rate

Largemouth bass used in this study did not eat at the recommended rate of 5% body

weight per day. Instead, pellets in the amount of 5% of tank's total body weight were

offered and the pellets remaining after approximately 10 min were counted. From the

number of remaining pellets, an estimate of total consumption was made for each tank on

each day (Table 3-1 and 3-2). Over time it became apparent that the fish were eating at

an approximate rate of 1% body weight per day, and so the feeding rate was adjusted to

1%. Total weight of pellets consumed over 30 or 50 days was used to determine total

exposure.

30-Day Exposure

p,p'-DDE accumulation

After 30 days exposure, bass in the control and dieldrin groups had a background

level of 7 1.7 ppb ofp,p '-DDE which did not vary among tanks, but did significantly

differ from fish fedp,p '-DDE contaminated feed types; sinking and floating (Figure 3-

2A).

Replicate 1 (Tank 3) for sinking p,p'-DDE contaminated feed was fed

approximately 115 |tg ofp,p '-DDE per fish over the course of 30 days. The whole-body

concentration was 495 + 144 ppb, equal to 81 15% accumulation. Replicate 2 (Tank 4)

was fed an average of 114 |tg ofp,p '-DDE per fish over the course of 30 days. However,

the whole-body concentration was 350 77 ppb, equal to only 54 3.7% accumulation.

The whole-body concentrations between replicates for sinkingp,p '-DDE contaminated









feed were significantly different from each other (Figure 3-2,A) but the percent of the

total dose accumulated was not significantly different between replicates (Figure 3-3).

Replicates 1 and 2 (Tanks 5 and 12) for floating p,p'-DDE contaminated feed

consumed an average of 109 and 106 |tg ofp,p '-DDE per fish, respectively, over the

course of 30 days. The whole-body concentrations were 341 51 and 347 41 ppb,

respectively, equating to 65 24.5% and 49 19.1% accumulation ofp,p '-DDE (Figures

3-2 and 3-3). The whole-body concentration and percentage of total dose accumulated

between replicates for floating p,p '-DDE contaminated feed were not significantly

different from each other.

Overall, one replicate of the sinking p,p '-DDE treatment was higher in whole-

body concentration and percent accumulation relative to all other replicates, sinking and

floating. The fish fed floating p,p '-DDE diets had a lower standard deviation from the

mean, indicating less variability among fish within a replicate.

Dieldrin accumulation

After 30 days exposure, bass in the control andp,p '-DDE groups had a

background level of 1.7 0.34 ppb of dieldrin, which did not vary among tanks, but did

significantly differ from all fish fed dieldrin contaminated feed; sinking and floating

(Figure 3-2B).

Replicates 1 and 2 (Tanks 6 and 10) for sinking dieldrin contaminated feed were

fed an average of 23 |tg of dieldrin per fish, respectively, over 30 days. Whole-body

concentration determined by GC-MS were 148 96 and 88 30 ppb, respectively,

equating to 74 24% accumulation for replicate 1 and 82 34% accumulation for

replicate 2. The whole-body concentration and percentage of dose accumulated were not









significantly different between replicates for sinking dieldrin contaminated feed (Figures

3-2, B and 3-3).

Replicates 1 and 2 (Tanks 2 and 8) for floating dieldrin contaminated feed were

fed an average of 19 and 21 |tg of dieldrin per fish, respectively, over 30 days. Whole-

body concentrations, as determined by GC-MS, were 114 7 and 120 6 ppb,

respectively, equal to accumulation rates of 93 16% for replicate 1 and 89 17% for

replicate 2. The whole-body concentration and percentage of dose accumulated did not

differ between replicates (Figures 3-2, B and 3-3).

Overall, the whole-body concentration and percent accumulation did not differ

between replicates and feed types, sinking or floating. However, the standard deviation

from the mean was smaller for replicates of the floating dieldrin diet.

GSI and sex steroids

GSI for males (n = 20, 0.42 0.07) and females (n = 16, 1.5 0.28) in all treated

groups was not significantly different from controls after 30 days exposure (analysis not

shown).

Blood plasma concentrations of sex steroids in both females and males did not

vary significantly by treatment after 30 days (data not shown). Because only 3 fish from

each treatment replicate were sampled, the sample size for either males or females ranged

from 0-3 depending on what was sampled for that replicate. No valid conclusions can be

drawn from the statistical analysis performed due to small sample size.

50-Day Exposure

The data from 30 days exposure indicated that there was less variability (lower

standard deviation from the mean) among fish within a tank, as well as among replicates,









for the floating pellet feed type. Use of sinking pellets was discontinued after 30 days

and fish were sacrificed. Contaminant analysis was run only for a subset of three fish

from each replicate of either a floating p,p '-DDE diet or a floating dieldrin diet at Day 50.

It was assumed that the background levels of either chemical in the control fish did not

change from Day 30 values and control bass were not analyzed for contaminants.

p,p'-DDE accumulation

After 50 days exposure, bass in the dieldrin group accumulated a background of

6.5 + 0.5 ppb ofp,p '-DDE, which did not vary among tanks, was not significantly

different from Day 30 controls, but did significantly differ from all fish fed floating p,p '-

DDE contaminated feed (Figure 3-4, A).

Fish from replicate 1 (Tank 5) were fed approximately 150 tg p,p '-DDE over the

course of 50 days. The whole-body concentration after 50 days was 314 30.5 ppb,

equal to an accumulation of 36 9% of the total dose administered. Replicate 2 (Tank

12) was fed 150 [tg p,p '-DDE per fish over 50 days and had a final whole-body

concentration of 337 41 ppb, equal to an accumulation of 48 15% of the total dose.

Analysis of whole-body concentration and percentage of total dose accumulated between

replicates revealed no significant differences (Figures 3-4, A and 3-5). Gonadal

concentration ofp,p '-DDE was determined to be 420 103.8 ppb for replicate 1 and 482

74.3 ppb for replicate 2 (Figure 3-6). Although gonad concentrations demonstrated

greater variability, means of the replicates were not significantly different. The gonadal

dose contributed approximately 0.5% and 1.0% of the whole-body concentration for

males and females, respectively.









Overall, a diet consisting of floating pellets contaminated with p,p '-DDE

produced a fairly consistent dosing method over the 50 day exposure time.

Dieldrin accumulation

After 50 days exposure, bass in thep,p '-DDE group accumulated a background

of 1.1 + 0.4 ppb of dieldrin which did not vary among tanks, was not significantly

different from Day 30 controls, but did significantly differ from all fish fed a diet of

floating dieldrin contaminated feed (Figure 3-4, B).

Each fish from replicates 1 and 2 (Tanks 2 and 8) were fed approximately 26 and

31 [g dieldrin, respectively, over the course of 50 days. The average whole-body

concentration after 50 days was 92 13 ppb for replicate 1 and 116 20 ppb for replicate

2, equating to accumulations of 61 + 15% and 64 14% of the total dose administered,

respectively. The whole-body concentration of dieldrin for fish in replicate 2 were

significantly higher than the fish in replicate 1, which is to be expected as replicate 2 was

fed approximately 20% more dieldrin (Figures 3-4A). The percentage of the total dose

accumulated did not differ significantly between replicates (Figure 3-5). Gonadal

dieldrin concentrations offish in replicates 1 and 2 were 196 64.9 and 547 144 ppb,

respectively (3-6). Corresponding to the whole-body concentration, gonadal dose of

dieldrin for fish in replicate 2 was significantly higher, nearly double that of replicate 1.

The gonadal dose contributed approximately 1.0% to the whole-body concentration of

replicate 1 and 3.5% to the total internal dose of replicate 2 with no apparent difference

between males and females.

GSI and sex steroids

GSI for males (n = 44, 0.40 + 0.11) and females (n = 22, 1.31 0.46) did not vary

among replicates or chemical administered.









Female sex steroids, both E2 and 11-KT, did not significantly differ between

replicates for any treatment. E2 concentrations were significantly lower for both p,p '-

DDE and dieldrin treatment as compared to the controls, however, there was no

difference in 11-KT for either treatment in females as compared to controls (Figure 3-7).

Floating control replicate 1 had three females with means of 386 82.4 pg/ml for E2 and

78 21.4 pg/ml for 11-KT. Floating control replicate 2 had four females with means of

404 61.9 pg/ml for E2 and 115 13.6 pg/ml for 11-KT. p,p'-DDE replicates 1 and 2

had five and three females with means of 215 86.6 and 153 38.1 pg/ml for E2 and 25

+ 13.3 and 38 46.8 pg/ml for 11-KT, respectively. Dieldrin replicates 1 and 2 had four

and three females with means of 125 100.1 and 158 86.3 pg/ml for E2 and 20 24.4

and 10 + 5 pg/ml for 11-KT, respectively.

Male sex steroids did not significantly differ between replicates of any treatment

and for bothp,p '-DDE and dieldrin treated males, E2 and 11-KT were significantly lower

than controls (Figure 3-8). Floating control replicate 1 had eight males with means of

289 40.8 pg/ml for E2 and 270 191.9 pg/ml for 11-KT. Floating control replicate 2

had seven males with means of 238 48.3 pg/ml for E2 and 251 + 102.6 pg/ml for 11-

KT. p,p '-DDE replicates 1 and 2 had six and seven females with means of 137 136.8

and 102 60.7 pg/ml for E2 and 79 92.2 and 66 15.7 pg/ml for 11-KT, respectively.

Dieldrin replicates 1 and 2 had seven and eight males with means of 73 30 and 78 48

pg/ml for E2 and 117 106 and 32 9.2 pg/ml for 11-KT, respectively.

Discussion

The results of this study suggest that diets containing OCPs provide an accurate

route of administration for largemouth bass with only moderate variability within

replicates. Sinking style feed produced greater variability both within and among









replicates for both pesticides and was most likely due to competition among fish for the

pellets before they sank to the bottom of the tank. This style of pellet is not conducive to

creating consistent doses in all fish within a tank. The floating style feed pellets showed

greater consistency in achieved whole body dose. Standard deviations from the mean, as

well as replicate differences, were lower in tanks fed floating style feed. The statistically

significant difference in whole body concentration seen at Day 50 in the dieldrin treated

groups occurred because of the difference in the total amount of dieldrin fed between the

tanks; the percentage of total dose accumulated did not differ. The consistency observed

within the floating feed type replicates indicates the superiority of floating feed over

sinking feed as an exposure route despite the suggestion that the method of preparing oil

dressed pellets is inherently less accurate. The remainder of the discussion will focus

only on the results obtained from fish fed a floating style diet.

The rapid uptake of bothp,p '-DDE and dieldrin from the diet of largemouth bass is

consistent with studies by Grzenda et al. (1971) that demonstrated rapid uptake for both

pesticides in goldfish fed contaminated diets. In these studies, even though exposure

continued past the initial uptake phase, accumulation rate showed a tendency to level off

after 30 days of exposure and tissues maintained a consistent dose, indicating the rate of

elimination was equal to that of the dosing (i.e. a steady state was reached). In Atlantic

menhaden, Brevoortia tyrannus, (Warlen et al., 1977), no equilibrium point was found

within a 60 day oral exposure to DDT. Studies on rainbow trout show that equilibrium

for both DDT and dieldrin did not occur until 140 days after initial dietary exposure

(Macek et al., 1970). Variability in absorption and elimination strategies may explain the

differences seen between fish species.









While this study measured whole body concentrations after just two time points, it

is clear that internal dose of either pesticide did not increase after Day 30 despite

continued daily dosing (Figure 3-9). Therefore, results suggest that largemouth bass may

undergo a rapid uptake phase followed by increased elimination. Absorption efficiency

from the intestine of fish can be greater than 50% for lipophilic compounds such as OCPs

(James and Kleinow, 1994). The percentage of the total dose accumulated at Day 30 in

this study was approximately 50% forp,p '-DDE and 90% for dieldrin. The accumulation

rates dropped by approximately 10% and 30%, respectively, by Day 50. This could be

due to a decrease in absorption efficiency or an increase in biotransformation and

elimination of the absorbed dose. Most likely, elevated elimination rates are the cause for

the decrease in overall accumulation; OCPs are known to induce liver biotransformation

enzymes at high doses in fish (Machala et al., 1998; Zapata-Perez et al., 2000; Foster et

al., 2001) and mammals (Dickerson et al., 1999; Sanderson et al., 2002; Leavens et al.,

2002).

Inconsistent with all other studies on accumulation of DDE/DDT and dieldrin in

various fish species was the total percentage of the accumulated dose. For DDT, uptake

was stated to be between 20 35% in several studies on rainbow trout, brook trout and

Atlantic menhaden over periods of 140, 120, and 48 days, respectively (Macek et al.,

1970; Macek and Korn, 1970; Warlen et al., 1977). The largemouth bass in this study

accumulated a total of 35 48% of thep,p '-DDE dose at Day 50, approximately 10%

more than previously reported. For dieldrin, studies with rainbow trout (Macek et al.,

1970) and striped bass (Santerre et al., 1997) showed only a 10% accumulation over 140

and 84 days, respectively, while the largemouth bass in this study accumulated 60 64%









of the total dieldrin dose after 50 days. Of mention is that the DDE/DDT and dieldrin

accumulation studies cited were generally of much longer durations and only reported

final percent accumulation, which may account for the smaller total accumulated dose

than reported here. The higher rate of accumulation and/or retention of lipophilic

compounds such as OCPs may also be due to the high fat content of largemouth bass,

especially at the time period in which this experiment was conducted. The lipid content

of largemouth bass is between 6 10%, increasing during the spawning season due to the

development of the gonad (Barziza and Gatlin, 2000; Brecka et al., 1996). Reported lipid

contents for rainbow trout and herring (Atlantic menhaden is in the herring family) are

most similar to the bass at approximately 6 and 9%, respectively while carp is 1 4%

(Ackman, 1995). Species variation in absorption efficiency may also play a large role.

Gonad concentrations of both OCPs had large variability within replicates, but

generally related to a consistent percentage of the whole-body concentration. p,p '-DDE

showed a greater partitioning into the female gonad than the male most likely due to the

high lipid content of the developing oocytes within the female gonad. Dieldrin did not

show such a difference in partitioning between the sexes, however, there was a replicate

difference due to different total doses of dieldrin fed over 50 days. It seems that as the

available dieldrin dose increases, the greater the percentage is that is stored in the fatty

tissue of the gonad. The sample size for the estimations of OCP concentrations within a

tank was small (n = 3) and there were not enough animals to determine differences

between sexes in one replicate. A larger sample size would aid in determining if a

dieldrin partitioning difference exists between sexes.









Nonetheless, the achieved whole-body concentrations of bothp,p '-DDE and

dieldrin demonstrated an effect on the endocrine system of treated fish. This manifested

as a decrease in circulating E2 in female and decreases in both E2 and 11-KT in males

after 50 days of exposure. From the data collected at Day 30, it is difficult to determine if

there were any trends in sex steroid levels due to treatment because of the uneven sex

ratio and small sample size collected. The GSI and circulating level of sex steroids from

Day 50 indicate that the fish used in this study were not sexually mature, as GSI levels

were 5 10 times less than normal and hormone levels were 10-fold less than what has

been shown for both male and female adult hatchery-raised bass (Gross et al., 2002). In

addition, GSI in males and females did not increase from Day 30 to Day 50. In fact,

GSIs from those time points did not significantly differ for males or females (Figure 3-

10) as normally occurs during this time period. The endocrine disruption seen in these

fish may not be of the same type or magnitude found in sexually mature fish containing

the same dose. Depressed sexual development may have been due to stress induced by

placement of fish in captive conditions.

Whole-body concentrations of treated fish, achieved by Day 30 and maintained

until Day 50, were approximately 375 ppb forp,p '-DDE and 100 ppb for dieldrin.

Gonadal concentrations determined for three fish per replicate were highly variable and

inconsistent between replicates for dieldrin (Figure 3-6). However, when the average

gonadal doses of pesticide achieved in this study are compared to gonadal doses of fish

found in affected areas of the Ocklawaha River basin, p,p '-DDE levels were lower by

10-fold and dieldrin levels were approximately equal (Marburger et al., 1999). (See

Chapter 2 discussion for theories of endocrine disruption by OCPs.)









In summary, exposing largemouth bass to OCPs by incorporation into an oil carrier

and subsequent coating of feed pellets is an effective, accurate, and reproducible dosing

method. Variability in achieved whole-body concentration and percentage of total dose

accumulated was minimal for treatment with bothp,p '-DDE and dieldrin. A rapid uptake

followed by a sustained whole-body concentration after 30 days indicates that study

length should be a minimum of one month. In addition, effects of these OCPs on

endocrine function during the onset of reproductive season were not manifest until after

50 days of continuous exposure, despite maintenance of approximately the same whole-

body concentration. This indicates that the study length should be around two months

and begin before January to obtain appropriate exposure duration to mimic chronic

exposure in the wild before the onset of spawning. Of future interest may be the study of

induction of biotransformation enzymes over the course of the exposure period. Further

studies examining OCP accumulation (absorption and elimination) over a longer period

of time should also be done to optimize the oral exposure method for largemouth bass.














8
Dieldrin
Floating






5
DDE
Floating


12 11
DDE Control
Floating Floating




9 10
Control Dieldrin
Sinking Sinking


7
Control
Sinking






6
Dieldrin
Sinking


4
DDE
Sinking




1
Control
Floating


3
DDE
Sinking




2
Dieldrin
Floating


Figure 3-1. Tank setup for preliminary feeding study. Each set of four tanks had a
divided well water and air supply line and separate drainage lines.









Table 3-1. Ten-day increment totals of feed eaten per tank. The total feed eaten by each
tank was totaled at Day 30 and used to estimate the average amount of
pesticide ingested by each fish.


Feed
Treatment Tank Days
1-10
(g)


Control
Floating
Dieldrin
Floating
DDE
Sinking
DDE
Sinking
DDE
Floating
Dieldrin
Sinking
Control
Sinking
Dieldrin
Floating
Control
Sinking
Dieldrin
Sinking
Control
Floating
DDE
Floating


1 196.3


Feed
Days
11-
20

174.3
174.3


Feed
Days
21-
30

175.3
175.3


Day
30
Total
Feed

916.6
916.6


Grams
per
Fish


Tank
DDE
(ltg)


36.7


2 182.4 150.8 138.9 805.2 32.2


Per
Fish
DDE
([tg)


Tank
Dieldrin
(ltg)


0 0


0 0 805.15


3 201.1 197.8 175.0 972.8 38.9 4863.9 194.6

4 192.4 201.4 175.3 962.8 38.5 4813.8 192.6

5 198.0 173.3 175.3 917.9 36.7 4589.7 193.6


6 199.9 193.4 174.0 960.7 38.4

7 200.8 187.7 170.3 947.2 37.9

8 189.2 163.8 175.3 881.4 35.3

9 194.6 200.5 175.3 961.6 38.5

10 191.0 195.6 172.8 946.1 37.8

11 200.8 172.2 175.2 921.1 36.8


0 0 960.7


0 0


0 0 881.4


0 0


0 0 946.1


0 0


12 192.9 161.2 175.3 883.5 35.3 4417.4 176.7


Per Fish
Dieldrin
(ltg)


32.2


38.4


35.3


37.8









Table 3-2. Ten-day increment totals of feed eaten per tank (floating only after Day 30).
The total feed eaten by each tank was totaled at Day 50 and used to estimate
the average amount of pesticide ingested by each fish.

Feed Feed Day Per
Days Days 50 Grams Tank Fish Tank Per Fish
Treatment Tank 31- 41- Total per DDE DDE Dieldrin Dieldrin
40 50 Feed Fish (tg) ((t g) (gg)
(g) (g) (g)
Control 1 106.7 103.3 1123.2 51.1 0 0 0 0
Floating
Dieldrin 2 73.6 74.7 930.5 42.3 0 0 536.7 25.6
Floating
DDE 5 90.4 87.8 1076.4 48.9 5381.8 244.6 0 0
Floating
Dieldrin 8 108.2 104.1 1096.2 49.8 0 0 677.3 30.8
Floating
Control 11 96.6 105.4 1109.3 50.4 0 0 0 0
Floating
DDE 12 102.1 90.2 1071.8 48.7 5358.9 243.6 0 0
Floating














750-
a
0.
. 500-
W b b
I.L

S250-


cccc
,, r r r N

.. .. .
,0 c^^,0 < < ,

"- "- G"o 0\d


150-



C
C 100


2- 50-



0
bb b

-\0 I 10 0
', ', <\o co


EI Sinking
- Floating


a1 Sinking
36 Floating


Figure 3-2. Day 30 whole-body concentration + SD ofp,p '-DDE (A) and dieldrin (B) in
largemouth bass fed a contaminated diet (n = 3 per treatment). All dosed fish
had significantly higher whole-body concentrations of their respective
chemical than the controls. One replicate fed sinkingp,p '-DDE contaminated
feed was significantly higher than all otherp,p '-DDE fed fish. There were no
significant differences among replicates fed dieldrin contaminated diets. Bars
with different letter designations indicate a significant difference.










I ISinking
Floating
a


DDE Dieldrin


Figure 3-3. Day 30, percentage of total dose accumulated + SD within each replicate (n
3 per replicate). There were no significant differences in the amount
accumulated between replicates of either feed type for dieldrin. There was a
significant difference between one DDE sinking diet replicate and one DDE
floating diet replicate. Bars with different letter designations indicate a
significant difference.







86



400- A a
a

300-
Q.
LLJ
0 200-
1.
Cx
100-

b b

^0 0









150- B


a
0.
0. 100-






c c
0o



oo0 o, o,_ o,_




Figure 3-4. Day 50 whole-body concentration + SDofp,p '-DDE (A) or dieldrin (B) in
largemouth bass fed a contaminated diet (n = 3 per replicate). All dosed fish
had significantly higher whole-body concentrations of their respective
chemical than the controls. There was no significant difference among p,p '-
DDE replicates. Whole-body concentration was significantly different
between dieldrin replicates. Bars with different letter designations indicate a
significant difference.












I Floating-1
m Floating-2


Dieldrin DDE


Figure 3-5. Day 50, percentage of total dose accumulated + SD within each replicate (n
3 per replicate). There were no significant differences in the amount
accumulated between replicates of either for eitherp,p '-DDE or dieldrin.
Bars with different letter designations indicate a significant difference.