<%BANNER%>

Effects of Organochlorine Contaminants on Hatchling American Alligator (Alligator mississippiensis) Growth


PAGE 1

EFFECTS OF ORGANOCHLORINE CONTAMINANTS ON HATCHLING AMERICAN ALLIGATOR ( Alligator mississippiensis ) GROWTH By JONATHAN J. WIEBE 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 2005

PAGE 2

Copyright 2005 by Jonathan J Wiebe

PAGE 3

This document is dedicated to Ralph Peter “J oey” Wiebe. Though I have not been able to see your face, your words, thought s, and style live on forever.

PAGE 4

iv ACKNOWLEDGMENTS I would like to thank my committee member s, Dr. Tim Gross, Dr. Dave Barber, and Dr. Franklin Percival, for their patienc e, understanding, and mo st importantly their interest in my project. Tim, I will never be able to truly express my thanks for all the opportunities that he has given me. I thank hi m for his counsel, beer making skills, and ability to know “almost” everything before it happens but, most of all I thank you for being my friend. Mom, I can’t say en ough about all of the love, support and understanding that she has provided. I thank her for being a great friend except for the following: Jon the Mexican baby, Stretch Mark s the Spot references, and Bulgur Wheat care packages. Cheryl, who is my all-time, fa vorite chick on this rock. I thank her for having a great attitude, closet neuroses, a nd removing that fishing hook. Janet, I cannot thank her enough for all of her help, guida nce, support, understanding and great food. Thanks for making me laugh at myself when I get… well the way that I get. Ruth, thanks for her supportive words of encouragement and wonderful sense of humor. Thanks to the many families that I call my own Smiths, Duncans, Greenans, Scarboroughs, Loverns, and Mitchells. All of you folks have showed tremendous support and kept me alive with your amazing hospitality and friendship. Heath, I thank him for his time, assistance as well as classic Arkansas st ories. Phil Wilkinson, Fr anklin Percival and Woody Woodward, I thank them for instilling in me an appreciation of alligators, southern jokes, and appreciation of fine BBQ cuisine. Dwayne Carboneau, I thank him for social commentary on not only alligator season but, lif e in general. Drs. Dan Sharp and Alan

PAGE 5

v Ealy, I thank them for providi ng time and assistance with my project. Finally, I thank all of my former and current lab mates: Trav is “Smitty” Smith, Carla “CW” Wieser, Jim “Roll Tide” Williams, Sherry “Lionheart” Bostick, Howard “Howie” Jelks, Nikki “Nicooola” Kernaghan, Shane “Prarie Boy” Ru essler, Alfred “Fredo” Harvey, Jessica “Gambusia Girl” Noggle, Kevin “The Stick” Johnson, Jessie “Piggy Girl” Grosso, Adro “Tweety Bird” Fazio, and James “The Tape Man” Basto. Your friendship, patience, and understanding throughout this MS e xperience are grea tly appreciated.

PAGE 6

vi TABLE OF CONTENTS page ACKNOWLEDGMENTS.................................................................................................iv LIST OF TABLES...........................................................................................................viii LIST OF FIGURES...........................................................................................................ix ABSTRACT....................................................................................................................... xi CHAPTER 1 LITERATURE REVIEW.............................................................................................1 Overview....................................................................................................................... 1 Organochlorine Contaminant Exposure and Endocrine Disruption in Alligators........2 Alligator Growth and Mortality in Re lation to Organochlorine Contaminants............4 Thyroid Structure...................................................................................................7 Thyroid Hormone Synthesis and Systemic Availability.......................................7 Thyroid Hormone Binding Proteins......................................................................9 Deiodination of Thyroid Hormones....................................................................10 Thyroid Hormone Availability and Synthesis among Oviparous Species.................12 Species-Differences in Thyroid Ho rmone Utilization and Regulation.......................13 Fish......................................................................................................................13 Amphibians..........................................................................................................13 Avian...................................................................................................................14 Physiological and Environmental In fluences on Thyroid Regulation........................15 Overview.............................................................................................................15 Reproductive and Thyroidal Seasonal Cycles.....................................................16 Nutritional Availability and Hibernation.............................................................18 Physiological and Environment Parameters Influence Growth...........................19 Effects of Organochlorine Contamin ant Exposure on Thyroid Regulation...............20 Overview.............................................................................................................20 Effects of Organochlorine Contamin ant Exposure on Alligator Thyroid Regulation........................................................................................................21 Thyroid Histology Alterati ons in Relation to Orga nochlorine Contaminant Exposure..........................................................................................................23 Influence of Organochlorine Contaminan t Exposure on Integrated Levels of Thyroid Hormone Regulation..........................................................................25 Thyroid Hormone Synthesis................................................................................25

PAGE 7

vii Thyroid Hormone Binding Proteins....................................................................26 Deiodination of Thyroid Hormones....................................................................27 Thyroid Hormone Excretion................................................................................28 Growth in Relation to p,p’-DDE, dieldr in, chlordane and toxaphene exposure.........30 Overview.............................................................................................................30 Experimental Data...............................................................................................31 Organochlorine Contaminant Exposure and Hatchling Alligator Growth.................34 2 MANUSCRIPT...........................................................................................................37 Introduction.................................................................................................................37 Materials and Methods...............................................................................................42 Egg Collection, Evaluation and Incubation.........................................................42 Clutch Selection...................................................................................................43 Animal Maintenance...........................................................................................44 Hatchling Morphometrics and Tissue Sampling.................................................44 Plasma Thyroid Hormone Validation Pro cedures (Total and Free Thyroxine)..45 Free T4 (FT4) Assay Procedures.........................................................................46 Total T4 (TT4) Assay Procedures.......................................................................46 Analysis of Chlorinated Analyt es from Alligator Egg Yolks.............................47 Statistics...............................................................................................................49 Results........................................................................................................................ .49 Clutch and Organochlorine Contaminant Parameters.........................................49 Hatchling Growth Rates......................................................................................50 Thyroid Hormones, Growth and Organochlorine Contaminants........................51 Discussion...................................................................................................................52 LIST OF REFERENCES...................................................................................................89 BIOGRAPHICAL SKETCH.............................................................................................98

PAGE 8

viii LIST OF TABLES Table page 2-1. Total length growth rate s among and within sites......................................................81 2-2. Snout-vent length growth rates among and within sites.............................................82 2-3. Head length growth rate s among and within sites......................................................83 2-4. Body weight growth rate s among and within sites.....................................................84 2-5. Hatchling alligator thyroid (TSI) and liver (LSI) somatic indices among sites.........85 2-6. Hatchling alligator thyroid somatic i ndices (TSI) within sites over time...................86 2-7. Hatchling alligator liver somatic i ndices (LSI) within sites over time.......................87 2-8. Multiple linear regression analysis of hatchling alligator growth rates,.....................88

PAGE 9

ix LIST OF FIGURES Figure page 2-1. Graphical interpretation of thyroid hormone biosynthesis.......................................61 2-2. Clutch fecundity and clut ch viability (site means)...................................................62 2-3. Clutch fecundity and clut ch viability (current study)...............................................63 2-4. Yolk OC concentrations. site means (a) and current study (b)................................64 2-5. Hatchling alligator growth parameters among sites over time.................................65 2-6. Hatchling alligator total lengt h (mm) within sites over time...................................66 2-7. Hatchling alligator snout-vent le ngth (mm) within s ites over time..........................67 2-8. Hatchling alligator head lengt h (mm) within sites over time...................................68 2-9. Hatchling alligator body weight (g) within sites over time......................................69 2-10. Hatchling alligator growth paramete rs (necropsy animals) among sites over time……...................................................................................................................70 2-11. Hatchling alligator total length (mm)(n ecropsy animals) within sites over time.....71 2-12. Hatchling alligator s nout-vent length (mm)(necropsy animals) within sites over time……...................................................................................................................72 2-13. Hatchling alligator head length (mm)(n ecropsy animals) within sites over time....73 2-14. Hatchling alligator body weight (g) (necr opsy animals) within sites over time......74 2-15. Hatchling alligator thyr oid weight (g)(necropsy animal s) within sites over time....75 2-16. Hatchling alligator live r weight (g) (necropsy animals) within sites over time.......76 2-17. Hatchling alligator to tal thyroxine(ng/ml)and free thyroxine (pg/ml) plasma concentrations among sites over time......................................................................77 2-18. Hatchling alligator tota l thyroxine (ng/ml) plasma c oncentrations within sites over time...................................................................................................................78

PAGE 10

x 2-19. Hatchling alligator free thyroxine (pg/ml) plasma concentrations within sites over time...................................................................................................................79 2-20. Graphical interpretation of factors that control the release of growth hormone.....80

PAGE 11

xi Abstract of Dissertation Pres ented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Master of Science EFFECTS OF ORGANOCHLORINE CONTAMINANTS ON HATCHLING AMERICAN ALLIGATOR ( Alligator mississippiensis ) GROWTH By Jonathan J Wiebe December 2005 Chair: Timothy S. Gross Major Department: Veterinary Medicine Alterations in alligator repr oductive and growth parameters have been reported in association with organochlorine (OC) contamin ated sites in central Florida. These data indicate reductions in egg and embryo quality as well as reductions in hatchling growth and survivability. Thyroid, a growth-regulating tissue, has been suggested as a key bioindicator of growth among seve ral species. In addition, severa l researchers have reported alterations in thyroid regulati on in relation to OC contaminant exposure. Previous field studies have reported alterations in alligator plasma thyroid hormone concentrations as well as several thyroid histolog ical parameters. However, thes e data were unable to relate plasma thyroid hormone (TH) concentra tions to alligator growth. Under captive conditions, preliminary data demonstrated th at hatchlings from high OC environments had hyperthyroid secretory patterns and acce lerated growth. The current study examined the same relationship; however an additional site with high OC contaminant concentrations was added in order to evaluate the effect s of OC contaminant exposure

PAGE 12

xii versus site as it relates to the observed a lterations in hatchling growth and thyroid regulation. In addition, a subset of hatchlings were sacrif iced bi-monthly to compare thyroid and liver weight (indicators of growth) with both hatchling external morphometrics and plasma TH concentrations over time. Though TH were shown to be bio-indicators of hatchling growth, no relationship was observed between OC contaminant exposure and hatchling alligator gr owth or plasma TH concentrations. These data suggest that hatchling alligator growth may be infl uenced by several key factors including an integrated e ndocrine network (GH, IGF-I, TH, corticoids), habitat degradation, as well as OC contaminant exposure.

PAGE 13

1 CHAPTER 1 LITERATURE REVIEW Overview During the 1980’s, significant reduc tions in American Alligator ( Alligator mississippiensis ) egg viability were observed on Lake Apopka (a site positioned at the headwaters of the Ocklawaha river basin with high organochlorine (OC) pesticide concentrations) in comparison with lake Woodruff, a national wildlife refuge with reduced concentrations of OC (Woodward, 1993; Rice et al., 1998). In addition, a severe (~ 90%) reduction in the juvenile alli gator population was observed on Lake Apopka (1981-1986) that was likely attributed to reproductive failure (Woodward, 1993). These observed reductions in juvenile survivability and adult reproductive success have been attributed in part to the influence of agriculture and anthr opogenic alterations specifically: extensive utili zation of organochlorine pe sticides by muck farming operations (i.e., ( 6,000 ha) of the lake’s northern we tland was converted for vegetable production), citrus crops, and e ffluent discharges from both the citrus processing plant and sewage treatment facility located at th e city of Winter Gard en (Woodward et al., 1993; Schelske and Brezonik, 1992). These envi ronmental alterations were compounded by the overflow of a wastewater pond located at the Tower Chemical facility which is adjacent to the Gourd Neck region of Lake Apopka (1980) consisting of high

PAGE 14

2 concentrations of sulfuric acid, DDT, dicofo l and several unidentified OC compounds in which by 1983, the EPA designated this faci lity’s property as a superfund site (Rauschenberger, 2004). Though several of thes e OC compounds were identified in yolk from alligator eggs, no direct association with reduced clutch viability was observed suggesting other cofactors (i .e., diet, population dynamics, specific OCP mixtures) might be involved and/or the devel opmental effects resulted from altered maternal physiology (caused by OC exposure) as opposed to dir ect embryotoxicity (Rauschenberger et al., 2004; Heinz et al. 1991). Therefor e, sites that have been hist orically impacted by varying degree of OC contamination (l akes Griffin and Apopka as well as the Emeralda Marsh Conservation Area) continue to demonstrat e coincident alterati ons in reproductive function and success as measured by sex steroi d biomarkers, sexual differentiation, clutch viability, embryonic mortality, post hatch su rvivability, and growth (Rauschenberger, 2004; Wiebe et al., 2002; Gross et al. 1994). Organochlorine Contaminant Exposure a nd Endocrine Disruption in Alligators Reductions in alligator reproductive su ccess as well as egg and embryo qualities have been observed in relation to sites with intermediate to high concentrations of OC contaminants (Rauschenberger, 2004; Masson, 1995). These chemicals have often been referred to as “endocrine di sruptors” or exogenous agents that interfere with the production, release, transport, metabolism, binding, action, or elimination of natural hormones in the body responsible for the main tenance of homeostasis and regulation of developmental processes (Rolland, 2000; Bruc ker-Davis, 1998). As some of these OC contaminants (i.e., p,p’-DDE) have been suggested to have positive and/or negative estrogenic or androgenic activity, plasma sex steroid concentra tions have been one of the principal biomarkers utilized to examine the relationship between exposure to OC

PAGE 15

3 contaminants and alterations in reproductiv e productivity. Gross et al. (1994) noted alterations in plasma sex steroids among juve nile alligators from lakes Apopka (high OC concentrations) and Woodruff (reference). Spec ifically, female juvenile alligators had significantly higher plasma estradiol concentr ations versus females from the reference site (Gross et al., 1994) In contrast, juvenile male alligators from lake Woodruff exhibited plasma testosterone concentra tions that were almost four times higher than males on lake Apopka (Gross et al., 1994). A similar inci dence of altered plasma testosterone concentrations in juvenile male alligators was reported by Guillette et al. (1999) among seven Florida lakes. In add ition, the author’s suggested a relationship between phallus size (a sex steroid-dependent tissue) as a bio-indicator of anti-androgenic or estrogenic contaminant exposure (Guillettte et al., 1999). Masson (1995) reported significant reductions in alligator clut ch viability (i.e. embryonic mortality) on lake Apopka (3.9%) versus conservation sites with low OC concentrations (71%). The author suggested that lake Apopka’s ex tremely variable, low clutch viability and hatch percentages confirmed the suggestion that a severe environmental problem exists at this lake site (Masson, 1995). Rice et al. (1998) observed that the majority of lake Apopka’s em bryonic mortality occurred during pre-egg deposition or in early incubation with the ne xt largest proportion of mortality occurring very late in incubation. These data continue to support several hypot heses: 1) maternal OC exposure alters reproductive regulation (a s demonstrated by alterations in plasma estrogen and testosterone concentrations) a nd, 2) the reported alterations in adult reproductive fitness as well as maternal-transfer of OC contaminants among yolk constituents appears to be related to th e observed increase in embryonic mortality.

PAGE 16

4 Alligator Growth and Mortality in Rela tion to Organochlorine Contaminants It has been suggested that many of the observed embryonic and post-natal alterations in offspring viability are the result in part of parental exposure to environmental contaminants (Guillette, 1995). Th is exposure is primarily associated with maternal transfer of lipophilic compounds (i.e., OCs) among yolk constituents to developing offspring (Rausche nberger et al., 2004; Wu et al., 2000). OC exposure has been suggested to alter hormones that contro l the course of devel opment and growth and may have the potential to alter differen tiation of major organ systems resulting in physiological and morphological changes (Rausc henberger et al., 2004; Wu et al., 2000; Guillette et al., 1995). Wiebe et al. (2001) repo rted significant alterations in alligator clutch viability and embryonic and post hatch survivabilities among sites of intermediate (Griffin) to high (Apopka and Emeralda Marsh) OC concen trations. These data were strengthened by Rauschenberger’s (2004) ex amination of the relationship between OC exposure and subsequent reductions in egg and embryo qualities under field and laboratory conditions. During 2000-2002 field co llections, eggs co llected from OC contaminated sites had higher fecundity, lower average clutch mass and reduced clutch viability in comparison with lake Lochl oosa, a site with determined low OC concentrations (Rauschenberger, 2004). Through the utilization of a captive adult alligator treatment study, populati ons (treated and control) we re orally dosed with ecorelevant doses of the four principal OC cont aminants identified from the previous field egg collection: DDT and metabolites (princip ally p,p’-DDE), dieldrin, chlordanes and toxaphene or vehicle control (Rauschenberg er, 2004). Though reduced clutch viability was observed in the treated vers us control clutches, the majori ty of the observed mortality was in the form of unbanded eggs which may represent either early embryonic mortality

PAGE 17

5 or lack of conception (Rotstei n et al., 2002). These data, fr om both field and laboratory, continue to suggest that overal l clutch survival appears to be related to total OC yolk or maternal burdens (Rauschenberger, 2004). Alterations in embryonic and hatchling growth as well as reduced post-hatch survivability in relation to OC exposure ha s been reported in the American alligator (Rauschenberger et al., 2004, Wiebe et al., 2002, Wiebe et al., 2001). It seems empirical that alterations in growth and survivability among animals in these OC contaminated environments would have ramifications at both site and population levels. Rauschenberger (2004) examined the incide nce of embryonic growth retardation and survivability in relation to OC exposur e utilizing an established embryo staging methodology (Ferguson, 1985). This eval uation not only examined embryonic morphological differences among sites over spec ific developmental time points but, also evaluated the histopathology of live and dead embryos from “best-case” (clutches with low mortality rates and low OC egg yolk concentrations) and “worst-case” (clutches with high mortality rates and high OC egg yolk conc entrations) clutches independent of site (Rauschenberger, 2004). These data demons trated several key points: 1) the youngest embryos sampled (calendar day 14 of artif icial incubation) showed the strongest relationship between OC egg concentrati ons and morphometric parameters, 2) morphology of live embryos was not consistent ly different among sites, except during calendar day 25 (timeframe signifies the middle of organogenesis and may be a more sensitive time period to OC exposure), 3) morphometry of live embryos was not significantly related to variati on in clutch mortality (i.e.., liv e embryos from clutches with high mortality rates develop similarly to thos e of low mortality rates) 4), cyclodienes

PAGE 18

6 (i.e., chlordane analytes) accounted for an av erage of 70% of the morphometric variation that could be attributed to OC variables which is surprising considering DDT and its metabolites compose an average of 66% of the total OC burden among all sites, 5) concurrent decreases in maturational age a nd mass of dead embryos in comparison with live embryos may have represented normal development up to a point at which the development stalled and the embryo eventu ally perished, or embryos could have developed at a much slower overall rate until the point at which they perished, and 6) no significant differences in histopathology we re observed among “best-case” and “worstcase” clutches. (Rauschenberger 2004). The principal mode of alligator embryonic exposure to OC contaminants has been suggested to occur via maternal transfer among yolk constituents. Several examples have demonstrated increased incidence of embryoni c mortality in relation to exposure to high concentrations of OC contaminants under both field and laboratory conditions. In addition, Rauschenberger (2004) detailed significant relations hips between OC exposure and subsequent reductions in embryonic gr owth and development. Therefore, OC contaminants are suggested to interfere w ith the regulation of critical growth and developmental time periods which may ultimate ly contribute to the observed increase in embryonic mortality on OC contaminated sites. These data demonstrate a critical need to better understand the physiologi cal role in regulating grow th and development among species exposed to OC contaminants. The thyroid is one of the principal regul atory tissues of growth and development among multiple taxonomic groups which has b een demonstrated to regulate diverse physiological endpoints including: metabolic rate, tissue differentiation and subsequent

PAGE 19

7 growth and development (Rousset and D unn, 2004). The two principal physiological actions of thyroid hormones consist of 1) regulation of cellular differentiation and development and, 2) regulation of metabol ic pathways (Rousset and Dunn, 2004). These general actions share a common integration in that changes in development and growth are due to both hormone modulation of meta bolism. In addition, ce llular differentiation changes inherently alter changes in gene e xpression, resulting in modulation of metabolic pathways (Rousset and Dunn, 2004). A detailed working knowledge of thyroid regulation is critical in understanding th e complex and integrated roles the thyroid plays in growth and development. Therefore, a literature review is provided which summarizes the principal factors that regulate thyroid function in cluding tissue structur e, thyroid hormone synthesis, availability, di stribution, and deiodination in both embryonic and post-natal life stages among several poi kilothermic as well as homeothermic species. Thyroid Structure The thyroid gland is a bil obular tissue that is organized into spherical follicles whose walls are composed of follicle cells that surround a central lumen filled with colloid (McNabb, 2000). Colloid is primarily composed of thyroglobu lin, a large protein which is constructed in the rough endoplasmic reticulum, glycosylated in the reticular lumen, and further post-translationally modified in the golgi apparatus of the follicle cell (Norman and Litwack, 1997). Thyroglobulin w ith its tyrosine residues provides the polypeptide backbone for the synthesis and st orage of thyroid hormones as well as an interim iodine storage area (McNab b, 2000; Norman and Litwack, 1997). Thyroid Hormone Synthesis and Systemic Availability The biosynthesis and secretion of thyr oid hormones requires four principal components including: thyroglobulin, thyrope roxidase, hydrogen pe roxide and iodide.

PAGE 20

8 Initially, dietary iodide is absorbed from the intestine and transferred from systemic circulation across the basal lateral membrane of the follicle cells utilizing an ATP-driven Na+ Iactive transport (Norman and Litwack, 199 7). The sequestered iodide is oxidized to iodine via thyroperoxidase enzymatic act ivity in the presence of hydrogen peroxide (principal electron acceptor) at the cell/colloid interface (McNabb, 2000). Concurrently, follicle cells synthesize thyroglobulin which cont ains select tyrosyl residues that will ultimately be iodinated and coupled to form either monoiodotyrosyls (MIT) or diiodotyrosyls (DIT) residues and stored as colloid (Norman and L itwack, 1997). In total, the catalyzing action of thyroperoxidase is re quired for the oxidation of iodide, iodination of the thyroglobulin tyrosyl residues and the coupling of the MIT and DIT tyrosyls (i.e., thyronines) which based on the coupling co mbination produces either triiodothyronine (T3) or thyroxine (T4) (Norman and Litwack, 1997). Systemic TH availability is regulated utilizing a classic negative feedback mechanism among the hypothalamic-pituitary-t hyroid (HPT) axis (Norman and Litwack, 1997). As thyroid hormones occupy their nuclear receptors in the anterior pituitary, it suppresses the transcriptional synthesis of pr eproTSH in the thyrotropes of the anterior pituitary (Norman and Litwack, 1997). U nder conditions of reduced T4, negative feedback is reduced on thyrotropes of the anterior pituitary (M cNabb 2000; Norman and Litwack, 1997) Thyroid-releas ing hormone (TRH) is secreted from the hypothalamus via the hypophyseal portal vessels inte racting with the anterior pi tuitary which results in the release of thyroid-stimulati ng hormone (TSH). TSH interacts with its 7 transmembrane, G coupled protein receptor on the thyroi d follicle cells (Norman and Litwack, 1997, Eales, 1984). As TSH is the most important cont rolling factor in iodine availability, the

PAGE 21

9 thyroid follicle will proceed to genera te free hormones from the stored hormones sequestered among thyroglobulin (Norman and Litwack, 1997). This is accomplished as the apical cell membrane engulfs the colloid by endocytosis and resulting cytoplasmic colloid droplets fuse with lysosomes to form phagolysosomes (Norman and Litwack, 1997). Thus, the internalized thyroglobulin molecules are subject to a variety of hydrolytic reactions leading to generation of free thyroid hormones and the complete degradation of the protein (Rousset a nd Dunn, 2004; Brown et al., 2004; McNabb, 2000; Norman and Litwack, 1997). Thyroid Hormone Binding Proteins Upon the release of TH from degraded thyr oglobulin, a system of plasma proteins that bind and distribute thyroid hormones is cr itical to counteract their loss from the vascular and interstitial compartments by permeation into cell membranes (Prapunpoj et al., 2002). These binding proteins are integral for systemic circulation due to THs high lipid solubility (Ric hardson et al., 2005; Prapunpoj et al., 2002). Albumin (ALB) and prealbumin or transthyretin (T BPA / TTR) are generally regarded as the two major T4 binding proteins throughout vertebrates; th ese having low binding affinity and high capacity (Licht et al., 1991). In add ition, many mammals possess thyroxine binding globulin (TBG), a separate high binding affi nity, low capacity binding protein that is responsible for the principa l portion of thyroid hormone binding (Licht et al., 1991). Thyroid hormone binding protei n(s) among vertebrate taxa de monstrate an evolutionary progression towards increasing thyroid hor mone distribution capacity during both developmental and adult life stages (Richards on et al., 2005). An exam ple of this can be observed in the binding protei n, transthyretin (TTR). TTR is transiently synthesized by the liver during the time of increased thyroid hormone concentrations (i.e., smoltification,

PAGE 22

10 metamorphosis and development) in fish, amphi bians, reptiles whereas it is synthesized by the liver during development and adult life st ages in eutherians and birds (Richardson et al., 2005). In crocodilians, TTR immunor eactivity has been detected in saltwater crocodile ( Crocodylus porosus ) serum on days 60, 68, 75 of egg incubation, and day 1 post-hatch, but not detected in serum at 6 months of age or a 3 year old animal. In addition, serum albumin was observed at all C. porosus age classes examined (Richardson et al., 2005). Prapunpoj et al. (2002) demonstrated that C. porosus TTR has higher binding affinity for T3 versus T4 suggesting that TTR was the principal transporter of T3 to the crocodilian brain. These data in conjunc tion with an observed higher percentage of amino acid sequence identity of C. porosus TTR to chicken TTR versus lizard TTR and, Chang et al. ( 1999) observation of avian TTRs having higher binding affinity for T3 versus eutherian TTR s suggest that the binding properties of C. porosus TTR are more evolutionarily similar to those of avian TTRs versus eutherian TTRs (Prapunpoj et al., 2002). Indeed, the se paration in evoluti onary functionality between eutherian, avian and poikilotherm thyr oid hormone regulation appears to be the eutherian’s ability to generate and regulat e thyroid hormones in a tissue-specific manner (i.e., the evolution of 5’ deiodinases) and th e utilization of additional binding proteins (i.e., TBG) which enhances t hyroid hormone regulation and distribution (Prapunpoj et al., 2002). Deiodination of Thyroid Hormones The delivery of the predominant circulating TH (T4) to specific target tissues (i.e., liver, choroid plexus) is critic al for the subsequent conversion of T4 to T3; which is considered the principal, biol ogically-active form of TH. Th e majority of systemic T3 availability for multiple taxa is generated via extrathyroidal mechanisms in these target

PAGE 23

11 tissues utilizing a process known as dei odination (Brown et al., 2004; McNabb, 2000). The process of deiodination is catalyze d by a family of selonoenzymes called deiodinases. These membrane-bound enzymes are located primarily in the microsomal fraction of tissue homogenates suggesting an endoplasmic reticulum and/or plasma membrane location (Hulbert, 2000). T4 is de iodinated by removal of iodine from the outer ring of the molecule (ORD) to produce T3 or the inner ring of the molecule (IRD) producing reverse T3 (rT3). ORD and IRD are catalyzed by three distinct deiodinases. Type I catalyzes both ORD and IRD by pref erentially removing phenolic and tyrosyl iodide. This type of deiodinase is probably located in all ti ssues but has especially high activity in the liver, kidney, thyroid tissue, and the central nervous system. Type II, catalyzes only ORD by removing only phenolic iodide and has been found in the central nervous system, brown adipose tissue, anteri or pituitary and placenta. Type III catalyzes exclusively IRD by removing only tyrosyl iodi de and is found in the central nervous system and the placenta (Shepherdley et al., 2002; Hulbert, 2000; Eales, 1984). The integrated nature of thyroid regulati on reflects a system principally regulated by classic endocrine feedback mechanisms In oviparous embryos, thyroid hormone synthesis and availability are governed by a developmentally-regulated system utilizing two sources: 1) maternal deposition in yolk (utilized during early stages of embryonic development) and, 2) embryonic endogenous sy nthesis (utilized duri ng later stages of embryonic development). The next section de tails the principal mechanism(s) that regulate oviparous embryo TH availability. In addition, a brief summ ary is provided to demonstrate species-differences in TH utilization and regulation.

PAGE 24

12 Thyroid Hormone Availability and Synthesis among Oviparous Species Thyroid hormone availability during embr yonic and early post-natal development in oviparous species has been principally investigated th rough the examination of TH synthesis, availability, compartmentalizati on, functionality, and util ization during several lifestages (Prati et al., 1992; Greenblatt et al., 1989; Tagawa a nd Hirano, 1987; Sullivan et al., 1987). The principal sources of thyroid hormones for developing oviparous embryos have been identified as maternal de position in yolk and endogenous synthesis by the embryo (Greenblatt et al., 1989). In sa lmonids, high-density lipoproteins (HDL) and vitellogenin (VTG), a yolk precurs or protein, have been identifi ed as the major carriers of thyroid and other hormones, vitamins, ions, a nd minerals from maternal circulation and subsequent sequestering in the yolk for the developing oocyte (Monteverdi and Di Giulio, 2000; Conley et al., 1997). In addition, Pra ti et al. (1992) suggested that TTR from chicken extra embryonic membranes may bi nd iodothyronines of maternal origin constituting the mechanism by which THs become available to the fetus before the onset of thyroid function. In an ex amination of the relationship between TH content and yolk mass, Sechman and Bobeck (1988) observed th at a linear increase in both T4 and T3 concentrations in oocytes was proportional to the weight of the yol k without changes in the iodothyronines content per 100 mg of yolk which indicated transfer of iodothyronines together with other yolk constituents as a principal source of TH for developing oocytes. Greenblatt et al. (1989) examined the compar tmentalization of both T4 and T3 in yolk and larvae in coho ( Oncorhynchus kisutsch ) and chinook ( O. tschawytscha ) salmon. These data demonstrated an asynchronous species difference in thyroid hormone utilization versus time between yolk reserv es and endogenous TH production (Sullivan et al., 1989). However, both species demonstrat ed a decreasing reliance on TH yolk

PAGE 25

13 reserves in step with an increase in endogenous TH producti on in relation to increasing larvae development (Sullivan et al., 1989). Species-Differences in Thyroid Horm one Utilization and Regulation Fish In teleosts, T4 has been reported as th e primary hormone released by the thyroid (Eales, 1985). Under TSH stimulation, Ea les (1985) reported a surge in both endogenously labeled and stable plasma T4 concentrations with no corresponding changes in plasma T3 concentrations. Kinetic studies have shown that about 80% of T3 in salmonids may reside in a slowly exch anging reserve pool, mainly represented by skeletal muscle (Brown et al., 2005). This c onstancy in plasma T3 concentrations is due at least in part to a rapid d ecrease in the proportion of av ailable plasma T4 peripherally monodeiodinated to plasma T3 (Eales, 1985) Though total thyroxine (TT4) and total triiodothyronine (TT3) plasma hormone concen trations have been shown to be highly correlated with their respective free plasma hormone concentrations, both percent free thyroxine (%FT4) and free tr iodothyronine (%FT3) plasma hormone concentrations demonstrated a negative correlation with TT4 and TT3 indicating that a smaller proportion of total hormone is free at higher total hormone concentrations (Eales and Shostak, 1985). In general, poik ilotherm plasma TH concentrati ons contrast with those of both Japanese Quail and humans where %F T3 exceeds %FT4, and are 3-5x higher than those reported in both trout and charr (Eales and Shostak, 1985). Amphibians Amphibian utilization of TH has been pr imarily reported duri ng several critical stages of metamorphosis (Galton and C ohen, 1980; Suzuki and Suzuki, 1980; Mondou and Kaltenbach, 1979). At stages V-XVIII (limb differentiation), plasma T4

PAGE 26

14 concentrations were undetecta ble suggesting that bullfrog ( Rana catesbeiana ) tadpoles were responsive to very low concentr ations of thyroid hormones (Mondou and Kaltenbach, 1979). During stage XIX (forelim b emergence) through stage XXI (tail resorption), a rapid increase was observed in both circulating plasma T4 and T3 concentrations (Suzuki and Susuki, 1981). In addition, the T3/T4 ratio of plasma TH concentrations suggested extrathyroidal deio dination during these stages of amphibian metamorphosis (Suzuki and Susuki, 1981). At the conclusion of metamorphosis (stages: XXIV – XXV), a rapid decline was observed in both plasma T3 and T4 concentrations in froglets of four months of age (Suzuki and Susuki, 1981). In adult frogs, low but detectable plasma T4 concentrations were observed (Mondou and Kaltenbach, 1979). Avian Birds possess the ability through the actions of thyroi d hormones to regulate and maintain thermal independence (i.e., home othermy) (Schew et al., 1996; McNabb, 1995). The initiation of avian thyroidal functi on is discriminatively observed among two separate modes of hatchling development: preco cial and altricial. Chicks of precocial species have dramatic peaks of plasma T3 a nd T4 concentrations at hatching, which is marked by the initiation of th ermoregulation. By contrast, altricial chick plasma TH concentrations are very low at hatching whic h is followed by a progressive increase by the time of the greatest endothermic imp rovements during nestling life (McNabb 2000; Olson et al., 1999). McNabb et al. (1 991) noted in Japanese quail ( Coturnix c. japonica ), a precocial species, that both plasma T4 and T3 concentrations as well as T3/T4 ratio increased following the chick’s penetration of the air cell. Thus, both TH release and utilization in quail increase concurrently with the beginning of pulmonary respiration and increased metabolic rate (McNabb et al., 1991) The proposed functionality of this rapid

PAGE 27

15 increase in TH release and utilization during the perinatal period proba bly institute a level of metabolic readiness and fi nal maturation of the nervous system (McNabb et al., 1991). In altricial species, a significan t increase in plasma T4 concen trations have been observed in the red-winged blackbird ( Agelaius phoeniceus ) from hatching to day 8 by which nestlings can achieve significantly large fact orial increases in both instantaneous and steady state rates of oxygen consumption in response to cold challenge (i.e., gradual cooling) versus their younger c ounterparts (Olson et al., 1999). In addition, early nestling blackbirds demonstrated increased plasma T3 concentrations which have been suggested to be important in the organization and matu ration of skeletal muscle essential for shivering thermogenesis (Olson et al., 1999). These data demonstrate the diverse and multifaceted roles that THs play in the areas of growth and development among severa l species. In additi on, thyroid regulation as well as growth and development have been reported to be signifi cantly influenced by several physiological and envir onmental parameters. Therefore, a review of the principal physiological and environmental effectors that have been re ported to influence thyroid regulation is provided. Physiological and Environmental Influences on Thyroid Regulation Overview Several studies have reported an inter-relation between physiological and environmental parameters and subsequent alterations in thyroid hormone regulation among a number of species (Kohel et al., 2001; Denver and Licht, 1991; Eales, 1985). Primarily, a seasonal, counter-r egulatory system involving pl asma T4 and testosterone (T) concentrations has been suggested among several poikilothermic species. In this system, plasma T4 generally increases in c onjunction with and beyond testis growth and

PAGE 28

16 subsequently regresses reproductive tissu es (Bona-Gallo et al., 1980). In addition, physiological and environmenta l factors such as: ambient and water temperatures, photoperiod, nutritional availability and hibernat ion have been reported to play critical roles in TH regulation among se veral poikilothermic and home othermic species (Kohel et al., 2001; Schew et al., 1996; Denver and Licht, 1991; Jallageas and Assenmacher, 1979). Reproductive and Thyroidal Seasonal Cycles Gonadal and thyroid seasonal cycles have been described for numerous reptile and avian species (Hulbert, 2000; Kar and Chandola-Sakalani, 1984; Licht et al., 1984; Bona-Gallo et al., 1980; Jallag eas et al., 1978). Bona-Gallo et al. (1980) examined both male and female cobra ( Naja naja ). In female N. naja plasma T4 concentrations were reported low in pre-vitellogenic animals, ro se significantly in vitellogenic and preovulatory animals and showed only a slight decline after ovulation (Bona-Gallo et al., 1980). Females demonstrated their greatest rise in plasma T4 con centrations during the peak of vitellogenesis but, these were obser ved to be much more variable than males (some values ranged up to 70 ng/ml) (Bona-Gallo et al., 1980). Male N. naja plasma T4 concentrations increased significantly in Marc h-April, coincident with rapid increase in testis weight however, plasma T4 concentratio ns demonstrated their greatest rise a full month after the peak in testis weight and plasma T concentrations (Bona-Gallo et al., 1980).These data suggest a distinct seasonality for plasma T4 concentrations in the male cobra as plasma T4 concentrations genera lly increased in conjunction with and beyond testis growth and subsequent regression (Bona -Gallo et al., 1980). Ja llageas et al. (1978) reported a strong inhibitory effect of elevat ed plasma T4 concentrations on sex steroid synthesis and secretion in male Peking ducks ( Anas platyrhynchos ) rather than LH concentrations suggesting that plasma T4 concentrations may be responsible for a

PAGE 29

17 seasonal state of reduced sensit ivity of the endocrine testis toward circulating LH. This suggestion, observed both in male Peking ducks and male teal ( Anas creeca ), was based on the observation that the highest concentr ation of plasma T4 coincided with a substantial decrease in circ ulating plasma T concentrati ons, whereas a transient rebound of plasma testosterone concentrations (Augus t/September) was associated with a decline in plasma T4 concentrations (Jallageas a nd Assenmacher, 1979; Jall ageas et al., 1978). Licht et al. (1985) noted a seas onal peak in plasma T4 concen trations in comparison with plasma T concentrations and follicle-stimula ting hormone (FSH) concentrations in the painted turtle ( Chrysemys picta ). Following emergence in mid-March to April, C. picta plasma T and FSH concentrations demonstrated a transient peak for about 2 weeks followed by a decline. In contrast, plasma T4 concentrations conti nued to progressively increase and did not peak until late May (i .e., the conclusion of reproductive activity). Licht et al. (1985) noted that pl asma T4 concentrations tended to fall more slowly or even remain relatively stable in spite of the obs erved decline in plasma T concentrations. Though a coincident regulatory pattern has been observed between plasma T4 and T concentrations, Licht et al. (1985) suggests that these se parate androgen and thyroid cycles may simply reflect independent or differential responsive ness of the gonads and thyroid to changing environmental stimu li in the temperate-zone reptiles. Several authors have experimentally dem onstrated the influence of both ambient temperature and photoperiod as it relates to te stosterone and thyroi d hormone synthesis and regulation (Jallageas and Assenmacher, 1979; Jallageas et al., 1978). In ducks and teal, cold environments have been shown to induce increased plasma T4 concentrations as well as moderate but, signifi cant inhibition of plasma T con centrations (Jallageas et al.,

PAGE 30

18 1978). However, these observed effects have no t been determined to be a clear inhibition of photogonadal response or merely an example of cold-induced hyperthyroidism increasing metabolic rate and subsequent inhibi tion of sex steroid secretion (Jallageas et al., 1978). Under artificial lighting conditions (20D: 4N), Wilson and Reinert (1999) noted that female tree sparrows ( Spizella arborea ) demonstrated both thyroid-dependent and thyroid-independent components that we re coincident with reproductive activity. Animals that received thyroidectomy (THX) demonstrated an inhibition of ovarian growth by 81 to 84% in comparison to (T HX) supplemented with T4 and controls. Interestingly, ovarian growth in THX anim als was still progressing whereas both THX supplemented with T4 and controls had co mpleted 40-50% of their postnuptial molt and significant ovarian reduction had occurred by day 84 of treatment (Wilson and Reinert, 1999). These data suggest that both temperat ure and delayed expression of absolute photorefractoriness (i.e., st ate of unresponsiveness to previously gonadostimulatory daylength which terminates breeding in many photoperiodic bird speci es) are associated with alterations in both reproductive and thyroid function (Wilson and Reinert, 1999; Jallageas and Assenmacher, 1979). Nutritional Availability and Hibernation Schew et al. (1996) examined the relati onship between food availability and TH regulation among precocial and altricial sp ecies. Initially, birds were placed on a maintenance diet (i.e., a limited ration of f ood was provided). Plasma T3 concentrations among both species were significan tly decreased not only compar ed to controls, but also compared to each species’ own values at the beginning of the restric tion period (Schew et al., 1996). Realignmentation (i.e., birds returned to ad libitum feeding), resulted in a rebound of plasma T3 concentrations among both species in comparison to controls

PAGE 31

19 (Schew et al., 1996). Upon emergence from their burrows, both male and female Desert Tortoise’s ( Gopherus agassizi ) demonstrated elevated plas ma T4 concentrations with increased feeding, activity (i.e., mating, locomo tion), and warmer temperatures (Kobel et al., 2001). Female tortoises exhibited a si ngle, dramatic increase in plasma T4 concentrations during the spring (i.e., warmer ambient temperatures and peak reproductive period) while males exhibited a long er plateau in plasma T4 concentrations throughout the summer (Kohel et al., 2001). Se llers et al. (1982) noted in the lizard ( Cnemidophorus sexlineatus ) significant increases in plasma T4 concentrations coincided with the entrance and emergence of hibernation. The author’s suggested that the observed increase in plasma T4 concentrations were th e result of decreased pe ripheral utilization of TH. Physiological and Environment Parameters Influence Growth Denver and Licht (1991) examined the inter-relationship between thyroid hormones, photoperiod, ambient temperature and growth utilizing slider turtles ( Pseudemys scripta ). Animals were treated by either sham, partial (PTX) or complete (TX) thyroidectomy (Denver and Licht, 1991) Significant reductions in plasma T4 concentrations and increased plasma TSH con centrations were observed in TX treatment versus sham. By 8 weeks (post-treatment), TX treatment had a significant reduction in both body mass and carapace length in comparison to sham treatment (Denver and Licht, 1991). Interestingly, partial groupings of sham PTX and TX treatment were maintained under either constant (30C am bient temperature, 27 1C wa ter temperature and constant light) or variable (40C to 24 C ambient temperature, 19C to 24 C water temperature and a 12L:12D photoperiod) environmental conditions (Denver and Licht, 1991). Under constant environmental conditions, growth rate s in the sham and TX treatments exhibited

PAGE 32

20 a significant decline whereas growth rate s of sham and TX animals under variable conditions declined only slightly by week 14 (Denver and Licht, 1991). These data demonstrate the profound influence of both phys iological and environment parameters on brain-pituitary-thyroid axis regu lation (Denver and Licht, 1991). OC contaminants have been reported to alter thyroid regulation producing deleterious effects in the areas of growth and development. As alligators have exhibited alterations in growth and survivability in re lation to OC exposure, a review is provided demonstrating reported alterations in TH synthesis, deiodination, delivery, activity, metabolism and excretion in relation to OC exposure. Effects of Organochlorine Contaminan t Exposure on Thyroid Regulation Overview Thyroid hormones are one of the princi pal regulators of di verse physiological endpoints including: metabol ic rate, oxygen consumption, tissue differentiation, and subsequent embryonic and post-natal growth an d development. However, these endpoints have been shown to be highly influen ced by a variety of physiological and/or environmental influences including but not limited to nutritional state, ambient temperature, photoperiod, and potentially coin cident counter-regulation by hypothalamicpituitary cascades involved in reproductive tissue development and subsequent reproductive quiescence. Current ly, environmental research has been examining the influence of introduced chemical compounds (i.e., environmental contaminants) which have been suggested to alter thyroid f unction, a growth-regula ting endocrine tissue (Brouwer et al., 1998). Many of the observe d actions of environmental contaminants have been reported to occur during embryonic development and sensit ive early life stages resulting in impaired reproduction and deve lopmental abnormalities in the offspring

PAGE 33

21 (Guillette, 1995). These chemicals have been referred to as “endoc rine disrupters” or exogenous agents that interf ere with the production, releas e, transport, metabolism, binding, action, or elimination of natural hormones in the body responsible for the maintenance of homeostasis and regulation of developmental processes (Rolland, 2000; Brucker-Davis, 1998). Due to the reported structural similarity among THs and chlorinated hydrocarbons (i.e., DDT, PCBs a nd dioxins), it has been hypothesized that these chemicals may elicit alterations in se veral areas of TH re gulation including: TH synthesis, deiodination, delivery, activity, metabolism and excretion (Brucker-Davis, 1998; Porterfield, 1994). Theref ore, OC contaminant exposure may contribute to the observed alterations in alligator embryoni c and hatchling growth, development and survivability. In order to examine this relations hip in greater detail, a detailed review is provided which 1) provides the current info rmation on alligator thyroid regulation and growth in relation to OC exposure, 2) pres ents reported alterati ons in both thyroid histology and regulation among se veral species in OC contam inated environments, 3) demonstrates the potential disruptive influence OC contaminants may have at all levels of thyroid regulation, and 4) provi des experimental data that demonstrate alterations in growth in relation to exposure by the four primary OC compounds identified among OC contaminated sites in central Florida: p,p’-DDE, dieldrin, chlordane and toxaphene. Effects of Organochlorine Contaminant Exposure on Alligator Thyroid Regulation American alligators ( Alligator mississippiensis ) have been considered a particularly suitable indicator species as th ey have been shown to bioaccumulate and biomagnify contaminants to levels equal to or greater than reported in birds and mammals (Crain and Guillette, 1998). Howeve r, an understanding of alligator thyroid function is limited as the principal data available is in relation to OC exposure

PAGE 34

22 (Gunderson et al., 2002; Crain et al., 1998). Crain et al (1998) noted a negative relationship with both plasma T3 and T4 concentrations and body size among male and female animals from lake Woodruff (low OC ). However, a general lack of correlation between plasma TH concentrations, sex and body size was observed in sub-adult alligators from both lakes Apopka and Okeechob ee (Crain et al., 1998). These data may potentially reflect altered re productive potential in these animals, as THs cooperatively regulate the reproductive activities of vertebra tes (Crain et al., 1998). Gunderson et al. (2002) and Hewitt et al. (2002) reported on su b-adult (0.9 to 1.5 m) alligator plasma T4 concentrations and quantitatively assessed sub-adult alligator thyroid function in sites with varying degrees of OC contamination in south Florida (Be lle Glade > WCA3A > Moonshine Bay). No obvious relationship was observed between body size and plasma T4 concentrations (Gunderson et al., 2002) Data generated from combined sampling years demonstrated that WCA3 A had significantly higher plasma T4 concentrations than either Belle Glade or Moonshine Bay (G underson et al., 2002). In addition, no differences in plasma T4 concentrations were observed between Belle Glade and Moonshine Bay (Gunderson et al., 2002). Howeve r, significant differences were observed between Belle Glade versus Moonshine Bay in epithelial width and colloid content (Hewitt et al., 2002). The author’s suggest an interrelation between the observed reduction in colloid content and reduced pl asma T4 concentration observed in Belle Glade animals. Therefore, reductions in the observed plasma T4 concentrations may be related to OC competition with TH for bindi ng proteins as well as elevation of UDP-GT enzymatic activity which induces T4 glucur onidation and subseque nt biliary hormone excretion. The inter-regulatory actions of both OC contaminant affinity for TH binding

PAGE 35

23 proteins and biliary TH excretion may ha ve led to the equivalent plasma T4 concentrations observed between Belle Glade and Moonshine Bay (Hewitt et al., 2002). Thyroid Histology Alterations in Rela tion to Organochlorine Contaminant Exposure Several field-oriented studies utiliz ing both qualitative and quantitative methodologies have provided insight as to the potential inte rrelation between environmental contaminant exposure and obser ved pathological th yroidal alterations among several species (Zhou et al., 1999; Moccia et al., 1986; Moccia et al., 1981; Sonstegard and Leatherland, 1976). Sonstegard and Leatherland (1976) noted that coho salmon ( Oncorhynchus kisutch ) from several Great Lakes had increased incidence of goiter (distinct growths located on the gill arches) and diffuse swelling at the base of the gill arches which is indicative of thyroid ne oplasia. Oblate or extremely elongated thyroid follicles with thickened, columnar shaped epithelial and extensive colloid vacuolation were observed among spawning coho ( O. kisutch ) and chinook ( O. tschawytscha ) salmon among the Great Lakes in comparison with the Fraser River (control site) (Moccia et al., 1981). In additi on, dense aggregations of t hyroid microfollicles were observed in many of the Great Lakes salmon (M occia et al., 1981). In order to assess the degree of observed thyroid hyperplasia, Mocc ia et al. (1981) deve loped a thyroid index for inter-lake and inter-species comparisons among the two salmon species. These data demonstrated a significant correlation betw een the thyroid index and observed goiter frequencies in the coho salmon (Moccia et al., 1981). The author’s reported a 13-fold difference in goiter frequency among Great La ke coho salmon populations (Moccia et al., 1981). Though the Great Lakes region has previously been documented with reduced iodine availability, the documented incidence of goiter has been reported to fluctuate over

PAGE 36

24 several years demonstrating goiters are not sole ly due to low iodine availability but, may be attributed to the presence of organo chlorine contaminants (principally: PCB congeners) in the environment (Moccia et al ., 1981; Sonstegard and Leatherland, 1978). In order to examine the direct effects of environmental contaminants and subsequent thyroid hyperplasia, Zhou et al (1999) quantitatively evaluated mummichogs ( Fundulus heteroclitus ) exposed to high sediment con centrations of PCBs, PAHs, DDT and various metals (Mercury, Lead, Coppe r, Zinc, Chromium, Cadmium) under both field and captive conditions. The author’s reported greater epithe lial height, larger follicles, and partially depleted colloid in fish from the contaminated site (PC) in comparison with the control site (TK) (Zhou et al., 1999). Both male and female fish from (PC) demonstrated a greater liver soma tic index (LSI) in comparison with animals from (TK) (Zhou et al., 1999). The author’s suggested that LSI may be utilized as a biomarker of extrathyroidal T4 conversion (Zhou et al., 1999). Fish (male and female) from PC demonstrated significan tly higher plasma T4 concentrations versus TK, which is different than what would typically be obser ved in goiterous fish (Zhou et al., 1999). No significant differences were observed in plasma T3 concentrations among PC and TK fish (Zhou et al., 1999). A captive reciprocal environment experiment was conducted utilizing animals and sediment from both c ontaminated and control environments (Zhou et al., 1999). These data suggest that the simulated PC envi ronment could elevate plasma T4 concentrations in TK fish, whereas an unpolluted environment could reduce plasma T4 and T3 concentrations in PC fish (Zhou et al., 1999). However, conditions of goiter as noted by Sonstegard and Leatherland (1978) were not observed in fish under field or experimental conditions (Zhou et al., 1999).

PAGE 37

25 Accumulation and biomagnification of high concentrations of lipophilic, polyhalogenated hydrocarbons has been suggested as an additive cause for the observed thyroid hyperplasia in several salmoni d species among the Great Lakes region (Sonstegard and Leatherla nd, 1978). Adult herring gull ( Larus argentatus ), a nonmigratory, piscivorous bird of the Great La kes region were utilized to quantitatively examine the incidence of thyroidal hyperpla sia in relation to dietary environmental contaminant exposure (Moccia et al., 1986). Great Lakes herring gulls demonstrated predominantly microfollicular follicles, en larged epithelial height, limited/no colloid versus established controls (Bay of F undy) which displayed normal thyroid morphology (Moccia et al., 1986). Many of the microfolli cular thyroids from Great Lakes herring gulls also had a severely hyperplastic epith elial component (Moccia et al., 1986). These data in conjunction with Moccia et al. ( 1981) demonstrated diffuse, microfollicular hyperplasia in both herring gulls and salmon in the Great Lakes region (Moccia et al., 1986). The author’s noted the increased preval ence of diffuse, microfollicular hyperplasia in most of the Great Lake collections and its absence in similar collections from the Bay of Fundy (control site) which are relatively free of environmental contaminants (i.e., lipophilic organohalogens) is consistent with the existence of thyr otoxic factors in the Great Lakes food chain (Moccia et al., 1986). Influence of Organochlorine Contaminant Exposure on Integrated Levels of Thyroid Hormone Regulation Thyroid Hormone Synthesis A wide variety of chemicals, drugs and ot her xenobiotics have been determined to affect thyroid hormone biosynthesis. A number of anions act as competitive inhibitors of iodide transport in the thyr oid, including perchlorate, th iocyanate, and pertechnetate

PAGE 38

26 (McNabb et al., 2004; Capen, 2001) In addition, several classe s of chemicals have been identified that inhibit the organification of thyroglobulin includ ing: 1) thionamides (thiourea, thiouracil, PTU), 2) alanine derivatives (sulfonamides), 3) substituted phenols, 4) and miscellaneous inhibitors (aminotri azole) (Capen, 2001). Many of these chemicals have been reported to exert their action by inhibiting thyroperoxidase, responsible for iodide oxidation to iodine, which results in the disruption of both iodination of tyrosyl residues in thyroglobulin and also the coupli ng reaction of iodotyrosines (i.e., MIT and DIT which form iodothyronines: T3 and T4) (Capen, 2001; McNabb, 2000). Thyroid Hormone Binding Proteins Concomitant reduction in plasma T4 con centrations has been reported in some cases to be an indication of compromised plasma transport sy stem for both ligands and of the presence of hydroxylated PHAHs on the TTR protein (Brouwer et al., 1998). Cheek et al. (1999) noted that hydr oxylated PCBs are potent ligands for TTR, having affinities in the 1 nM range, 50-fold grea ter than that of T4. TTR is a major T4 binding protein in the blood, and it shows in addition to the thyroxine binding sites a site that is complimentary to the DNA double helix, indi cating a possible relationship to the thyroxine nuclear receptor (Rickenbacher et al., 1986). The TTR molecule has two-fold symmetry, and the binding site is lined pr imarily with hydrophobic amino acid side chains that form polarizable pockets for hal ogen interactions (Ricke nbacher et al., 1986). In view of the highly hydrophobic/polarizab le nature of the TTR binding site, the author’s suggest that van der Waals / hydr ophobic interactions would be dominant in controlling the binding strength of biphenol compounds (Rickenbacher et al., 1986). Contaminants with the highest TTR bindi ng efficiencies were shown to have a para hydroxyl substituent flanked by two meta chlorines which is analogous to the

PAGE 39

27 diiodophenolic ring system in T4 (Rickenbacher et al., 1986). Va n den Berg et al. (1991) noted that chlorophenols demonstrated the highest level of competition for TTR binding utilizing a competition assay (i.e., radiolabelled T4, TTR versus individual contaminant). These data suggest that 1) inte raction with the T4 bi nding site is dependent on the degree of chlorination, 2) the combina tion of hydroxyl and chlorine groups is more competitive than either group separately, and 3) displacement of T4 from the binding site is by a competitive type of interaction (Van den Berg et al., 1991). The author’s noted that DDT isoforms such as p, p’-DDD, o, p’DDD as well as dicofol, in particular, were found to interact with TTR (Van den Berg et al., 1991). A large proportion of the chemicals with affinity for TTR appear to ha ve neurotoxic properties (Van den Berg et al., 1991). In addition, transthyretin has been reported as one of the few proteins identified in the cerebrospina l fluid (CSF) that is synthesi zed by the choroids plexus and may function in the transport of T4 through the blood-CSF barrier (Van den Berg et al., 1991). Therefore, chemicals inte racting with TTR may affect the transportation function of the choroids plexus with possible conseque nces on brain function (Van den Berg et al., 1991). Deiodination of Thyroid Hormones Iodothyronine deiodinase act ivity is principally respons ible for TH conversion in extrathyroidal tissues has been suggested as a more sensitive thyroidal index of contaminant exposure (Adams et al. 2000). Male plaice dosed (ip) with 5 ng PCB 77 / g body mass demonstrated reduced plasma T4 and T3 concentrations as well as increased hepatic T4 ORD activity during week one vers us week four post-exposure (Adams et al., 2000). Coimbra et al. (2005) not ed that Nile Tilapia receivi ng dietary treatments (0.1g Endosulfan / g -1 of food (EL), 0.5g Endosulfan / g -1 of food (EH), or 0.5g Arochlor

PAGE 40

28 1254 / g -1 of food (A)) demonstrated alterations in both plasma T4 and ORD activity (time points: days 21 and 35). Tilapia exposed to EL21 demonstrated lower plasma T4 concentrations than either EH (days 21 and 35), A (days 21 and 35), and control treatments (Coimbra et al., 2005). Plasma T3 concentrations were not significantly altered in any treatments (C oimbra et al., 2005). Liver DI ORD activity was found to be depressed by both EL treatments while liver D3 activity was found to be enhanced by the EL treatment in relation to time of expos ure (Coimbra et al., 2005). The observed changes in the activity of several deiodina ses could result in d ecreased plasma T3 availability (Coimbra et al ., 2005). The fact that plasma T3 concentrations remained unaltered, is probably indi cative of the prominent role of hepatic D2 activity and renal D1 activity, both of which remained stable (Coimbra et al., 2005). Thyroid Hormone Excretion Hepatic microsomal enzymes (specifically: uridine diphosphate glucuronsyltransferase UDP-GTs) play an important role in thyroid hormone economy/availability which is accomplished in part through glucuronidation (a ratelimiting step in the biliary excretion of T4) and sulfation (which utilizes phenol sulfotransferase for the excretion of T3) (Capen, 2001). Glucuronida tion and sulfation are responsible, in part, for the conversion/mob ilization of aglycones (parent compounds or phase I metabolites) into water-soluble conjugate s that can be subse quently excreted from the body (Parkinson, 2001). Sulfation and desu lfation appear to be very important pathways to regulate free TH concentrations in the fetal compartment (Brouwer et al., 1998). Since hydroxylated PCBs tend to accumu late in the fetal department, where sulfation is a major regulation pathway, it is hypothesized that the fetal regulation of free

PAGE 41

29 TH concentrations may be compromised by PHAHs which may have serious negative consequences for fetal and neonatal development (Brouwer et al., 1998). Several xenobiotics have been report ed to induce microsomal enzymes and disrupt function in rats in cluding: CNS-acting drugs (phe nobarbital) and chlorinated hydrocarbons (i.e., chlordane, DDT, and TC DD) and polyhalogenated biphenyls (PCB, PBB) (Capen, 2001). McClain et al. (1989) provided a detaile d assessment of hepatic T4UDP-glucuronyl transferase activity in phenobarb ital-treated rats. A significantly higher cumulative biliary excretion of 125I-labeled T4 was observed in rats orally treated with phenobarbital versus controls bile (McClain et al., 1989). The observed increase in biliary excretion was accounted for by an increase in T4-glucoronide resulting from increased T4 metabolism (McClain et al., 1989). This was consistent with enzymatic activity measurements which resulted in increased hepatic T4-UDP-glucuronyl transferase activity (McClain et al., 1989). In addition, hi stological alterations including: follicular cell hypertrophy followed by hyperplasia in associ ation with both a marked increase in biliary T4 excretion and sustained increases in TSH (McClain et al., 1989). These data are consistent with the hypothesis that the promotion of obs erved thyroid tumors in rats is not a direct effect of phenobarb ital treatment on the thyroid gl and but rather an indirect effect mediated by plasma TSH concentrations secreted from the pi tuitary secondary to the hepatic microsomal enzyme –induced increase of T4 excretion in the bile (McClain et al., 1989). In addition, signifi cant species differences in UDP-GT expression have been observed between rats and mice exposed to the PCB, Kanech lor-500 (K-500) (Kato et al., 2003). Though K-500 treatment resulted in a significant decrease in plasma T4 concentrations in both rats and mice, a si gnificant increase in UDP-GT activity was

PAGE 42

30 observed only in the rat (Kato et al., 2003). These data were further supported following K-500 treatment as gene expression of hepatic UDP-GT isoforms UGT1A1 and UGT1AG in the rat liver were enhanced prior to the decrease in plasma T4 concentrations as opposed to the mouse liver (Kato et al. 2003). Utilizing Gunn rats (UGT1A deficient) and Winstar rats (normal), Kato et al. (2 004) dosed both species with KC-500 and 2,2’,4,5,5’-Pentachloro biphenyl (PentaCB) examin ing deiodinase activity and additional mechanisms of biliary excretion of thyroid hormones. Plasma total T4 and free T4 concentrations were significantly de creased in both PCB treated species (Kato et al., 2004). In addition, type I deiodinase activit y (converts T4 to T3) in Winstar rats was significantly decreased by KC-500 but not by PentaCB, although in Gunn rats, it was significantly decreased by both PCB isoforms (Kato et al., 2004). These data led the author’s to suggest that PCB-mediated decr ease in plasma T4 concentrations does not occur through the induction of hepatic T4 glucuronidation enzymes (Kato et al., 2004). These conflicting reports rega rding UDP-GT activity prompted several authors to suggest potential mechanisms/factors that may i ndividually/collectively reduce plasma T4 concentrations including: displacement of T4 from transthyretin (TTR) binding by PCBs facilitating free T4 excretion in urine or bile alteration in the HPT axis, and/or increase in estrogen sulfotransferase, which efficien tly catalyzes the sulf ation of iodothyronines (Kato et al., 2004, McNabb and Fox, 2003). Growth in Relation to p,p’-DDE, diel drin, chlordane and toxaphene exposure Overview Several PCBs and organochlorine pesticid es (i.e., DDE, dieldrin, chlordanes, and toxaphene) have been suggested to alter thyroid regulati on in several species under experimental ( in-ovo and in-vivo ) conditions (Scollon et al., 2004; Nishimura et al., 2002,

PAGE 43

31 Willingham, 2001; Waritz et al., 1996; Jefferies and French, 1972). These OC pesticides have been previously identified in both al ligator maternal tissues and egg yolk which have been associated with alterations in alligator egg and embryo qualities as well as hatchling growth among several contaminated lakes and reclaimed agricultural properties in central Florida (Rauschenberger, 2004; Wiebe et al., 2002). TH regulation and alterations in thyroid histology in relation to OC exposure ha ve been primarily examined utilizing pharmacological dosing methodologies A consistent observation among several controlled treatment studies was thyroid gland histological alterations consisting of increases in overall thyroid weight, epithelial hyperplasia and colloid depletion in relation to exposure by several PCBs and/or OC pes ticides among several sp ecies (Fowles et al., 1996; Jefferies and French, 1972; Jefferies and Parslow, 1972; Fregly et al., 1967). Experimental Data As thyroid hormones are an integral component in embryonic and hatchling growth, the observed thyroidal alterations in relation to ch lorinated hydrocarbon exposure suggest the potential for s ubsequent growth alterations O’Steen and Janzen (1999) reported that plasma TH concentrations and re sting metabolic rate in hatchling snapping turtles ( Chelydra serpentina ) correlated with incubation temperature. As incubation temperature is strongly linked with sex determination in ma ny reptile species, compounds that mimic or antagonize steroid hormones may affect metabolism, TH concentrations, or growth rate (Willingham, 2001). Red Eared Slider ( Trachemys elegans ) eggs were topically treated prior to th e temperature-sensitive window of sex determination (Stage 17, from Yntema, 1968) with low, intermedia te, and high concentrations of either trans Nonachlor and chlordane or p,p’-DDE (Willingham 2001). Upon hatching, hatchling turtles were fasted for 28 days and subsequently re-fed ad-libitum for 14 days

PAGE 44

32 (Willingham, 2001). At the conclusion of a 28 day fast, the intermediate trans -Nonachlor group significantly lost mass in comparison with controls (Willingham, 2001). Following re-feeding, both the intermediate and high trans -Nonachlor groups significantly increased in mass (Willingham, 2001). The author suggests that the reduction in mass observed in several OC treatments may have elicited a temporal, hyperthyroid state in which yolk reserves were utilized more quickly, thus reducing overall mass. As was observed in Schew et al. (1996) following fasting and ad-libitum re-feeding, compensatory increases in mass were observed in several transNonaclor and p, p’-DDE treatment groups (Willingham, 2001). Janz and Bellward (1996) examined in-ovo exposure of 2,3,7,8tetrachlorodibenzo-p-dioxin (TCDD) in precoci al (chicken), semi-altricial (great blue heron) and altricial (pige on) species and subsequent alterations in growth and development (Janz and Bellward, 1996). In both chickens and great bl ue herons, no effect in plasma TH concentrations or hatchli ng growth and development was observed in relation to TCDD exposure (J anz and Bellward, 1996). Howeve r, pigeons exposed to TCDD demonstrated significan t reductions in both plasma TH concentrations and hatchling growth and development decrease s including: crown-rump length, wing length, and tibia length (Janz and Be llward, 1996). These data are re affirmed by the established temporal differences in TH maturation am ong precocial and altric ial species (McNabb 2000, Olson et al., 1999). Hatch ling Artic Glaucous Gulls ( Larus hyperboreous ) growth was assessed in relation to parental bird serum OC concentrations over a three year period (Bustnes et al., 2005). Adult female gu lls with high OC burdens spent significantly longer time periods in search of nutritional resources for th eir chicks (Bustnes et al., 2005). In addition, a significant negative relati onship was reported between chick growth

PAGE 45

33 and increasing adult OC serum concentrati ons of HCB, oxychlordane, p,p’-DDE, and several PCBs (Bustnes et al., 2005).. The author ’s suggest that there may be interactions between energy expenditure and different OC concentrations, and females with high OC concentrations may have fewer resources availa ble to provide for thei r chicks (Bustnes et al., 2005). In addition, significant reductions in weight were observed in juvenile (~ 37 day old) Nile Tilapia ( Oreochromis niloticus ) exposed to aqueous dieldrin (1.0 to 2.4 g/liter-1) for 30 days in comparison with controls (Lamai et al., 1999). Finally, Blanar et al. (2005) noted that juvenile Artic Charr ( Salvelinus alpinus ) orally dosed (1x) with toxaphene (10 g/g) demonstr ated decreased growth and overall body condition (k) as well as decreased muscle lipid and protein c ontent. These reports suggest the potential direct (i.e., feeding, injection, aqueous OC e xposure) and indirect (i .e., reduced parental fitness due to OC exposure) influences that OC contaminants may have to influence growth among several oviparous species. These data suggest that OC exposure can elicit alterations in both thyroid function and subsequent growth. Severa l field studies have reported severe alterations in both plasma T4 concentrations and thyroid hi stology in relation to OC contaminated environments among avian and several fi sh species (Rolland, 2000). In addition, controlled treatment stud ies utilizing either p, p’ DDE, di eldrin, chlordane, or toxaphene reported altered thyroid regulation and growth reduction. These data suggest that OC exposure may be related to the observed re ductions in alligator embryo and hatchling growth from OC contaminated sites in centr al Florida (Rauschenberger, 2004; Wiebe et al., 2001; Gross et al., 1994). In addition, severa l authors have reported modified alligator thyroid function in relation to OC exposure (H ewitt et al., 2002; Crain et al., 1998). These

PAGE 46

34 reported modifications have taken the forms of reductions in plasma T4 concentrations and changes in thyroid histol ogy compared with c ontrols. However, researchers must be keenly aware of both physiological (i.e., sex, age, nutritional availa bility, reproduction, hibernation) and environmental factors (i .e., ambient and water temperatures and photoperiod). These factors have been repor ted to vary thyroid regulation and may complicate data interpretati on regarding OC exposure and subsequent alterations in thyroid function. Therefore, a captive study providing a controlled, structured environment presents a more applicable m eans to test the relationship between OC exposure and subsequent differences in hatchling thyroid function and growth. Organochlorine Contaminant Exposure and Hatchling Alligator Growth Wiebe et al. (2002) evaluated hatchling alligator thyroid regulation and growth from several lakes in central Florida: lake s Apopka (high OC concentrations), Griffin (Intermediate OC concentrati ons), and lake Lochloosa (Low OC concentrations) under captive conditions for a period of 6 months. These experiment al conditions included: a restricted photoperiod (12D:12N), contro lled ambient and water temperatures, ad-libitum feeding twice a week, and restri cted number of animals per enclosure to limit stressful overcrowding. Though egg viability rates di d not differ among sites, lake Apopka hatchlings demonstrated a significantly hi gher growth rate and plasma T3 and T4 concentrations in comparison with lakes Gri ffin and Lochlooosa. These data suggest that lake Apopka hatchlings demonstrated a hypert hyroid secretory pattern resulting in an enhancement of hatchling grow th in relation to exposure to high OC concentrations. However, OC contaminants, due to their stru ctural similarity with THs, have been predominantly suggested to reduce TH system ic availability by competing for binding proteins. These conflicting data suggest the need for furt her examination of thyroid

PAGE 47

35 regulation among hatchling alligators exposed to OC contaminants. Specifically, hatchlings from a site of similar OC contam inants and concentrations (i.e., Emerelda Marsh Conservation Area) to lake Apopka shoul d be utilized in a comparative growth study. A comparison of hatchling thyroid regu lation and growth among several sites with high OC concentrations may provide furthe r insight (i.e., OC exposure versus sitespecific variables) into the observed hype rthyroid secretory pattern and accelerated growth rate observed in lake Apopka hatchli ngs. Therefore, a captive hatchling growth study was undertaken utilizing an imals from lakes Apopka, Griffin as well as Orange (a site with low OC concentrations) and Emeral da Marsh Conservation Area (Area #7) to assess if in-ovo exposure to high concentrations of OC contaminants elicits a hyperthyroid secretory pattern that accelerates hatchling alligator growth. The following hypotheses were tested by this study. Hypothesis #1 Ho: No change in hatchling growth rates will be observed among all sites in relation to high in-ovo OC contaminant exposure. Ha: In-ovo exposure to high concentrations of OC pesticides will ac celerate hatchling alligator growth rates in comparison with animals exposed to intermediate to low in-ovo OC concentrations. Hypothesis # 2 Ho: No change in hatchling TH secretory pattern will be observe d among all sites in relation to high in-ovo OC contaminant exposure.

PAGE 48

36 Ha: In-ovo exposure to high concentrations of OC contaminants will elicit a hyperthyroid secretory pattern in hatchling alligators that wi ll result in an accelerated growth rate in comparison with animals exposed to intermediate to low in-ovo OC concentration

PAGE 49

37 CHAPTER 2 MANUSCRIPT Introduction During the 1980’s, significant reduc tions in American Alligator ( Alligator mississippiensis ) egg viability were observed on La ke Apopka (high OC concentrations) in comparison with lake Woodruff, a national wildlife refuge (low OC concentrations) (Woodward, 1993; Rice et al., 1998). In addi tion, a severe (~ 90 %) reduction in the juvenile alligator populati on was observed on Lake Apopka (1981-1986) that was likely attributed to reproductive failu re (Woodward, 1993). The observe d reductions in juvenile survivability and adult reproductive success have been attributed in part to the influence of agriculture and anthropoge nic alterations specifically : extensive utilization of organochlorine pesticides by muck farming, c itrus crops, and effluent discharges from both the citrus processing plant and sewage trea tment facility located at the city of Winter Garden (Woodward et al., 1993; Schelske and Brezonik, 1992). These environmental alterations were compounded by the overflow of a wastewater pond located at the Tower Chemical facility, adjacent to the Gourd Neck region of Lake Apopka (1980), consisting of high concentrations of sulfuric aci d, DDT, dicofol and several unidentified OC compounds. This event resulted in the EPA designation of this property as a superfund site in 1983 (Rauschenberger, 2004). Though several of these OC compounds were identified in yolk from alligat or eggs, no clear association wi th reduced clutch viability was observed for specific OC contaminants (Rauschenberger et al., 2004, Heinz et al. 1991). Therefore, sites that have been histor ically impacted by varying degrees of OC

PAGE 50

38 contamination continue to demonstrate coincide nt alterations in reproductive function as measured by sex steroid biomarkers, sexual differentiation, clutch viability, embryonic mortality, post-hatch growth and survivab ility (Rauschenberger, 2004; Wiebe et al., 2002; Guillettte et al., 19 99; Gross et al. 1994). Guillette (1995) suggested that many of the observed embryonic and post-natal alterations in offspring viability are the re sult, in part, of parental exposure to environmental contaminants. OC exposure ha s been reported to alter hormones that control the course of growth and devel opment and may have the potential to alter differentiation of major organ systems re sulting in physiological and morphological changes (Rauschenberger et al., 2004; Wu et al., 2000; Guillette et al., 1995). Significant alterations in alligator clutch viability and embryonic and post-hatch survivability have been reported among sites of intermediate to high OC concentrations suggesting an interrelationship between in-ovo OC exposure and subsequent reductions in embryonic and hatchling survivability (Wiebe et al., 2001). The predominant exposure route for developing offspring would be maternal transfer of OC contaminants among yolk constituents (Rauschenberger et al., 2004; Wu et al., 2000). Rauschenberger (2004) noted that eggs collected from OC contaminated sites had higher fecundity, lower average clutch mass and reduced clutch viability in comparison with sites with low OC contamination (Rauschenberger, 2004). The observed alterations in embryo morphology appear to be in association with variation in OC contaminant burdens of eggs. In addition, OC analyte composition was determined to be equally as importan t as concentration, suggesting the importance of mixture com position (Rauschenberger, 2004). These data demonstrate a need to better understand th e physiological and/or chemically-induced

PAGE 51

39 mechanisms that may effect alligator grow th and development from OC contaminated sites (Rauschenberger, 2004; Wiebe, 2001). One of the principal regulators of growth and development among multiple taxonomic groups are thyroid hormones (TH) wh ich have been demonstrated to regulate diverse physiological endpoints including: metabolic rate tissue differentiation and subsequent growth and development (R ousset and Dunn, 2004). Several literature reviews have suggested that a lterations in thyroid function ma y be in relation to exposure to a variety of compounds including OC contaminants (Rolland, 2000; Brucker-Davis, 1998). Due to the structural similarities between THs and DDT, PCB’s and dioxins, these chemicals may act as weak agonists that have the potential to reduce/block thyroid hormone activity (Brucker-Dav is, 1998; Porterfield, 1994). French and Jefferies (1972) however, noted that pigeons fed low concentr ations of p,p’-DDE and dieldrin induced hyperthyroidism whereas higher doses of bot h OC contaminants caused hypothyroidism. Therefore, thyroid regulati on alterations due to OC c ontaminant exposure may have contributed to the observed variation in alligator embryo and hatchling growth and development. American alligators ( Alligator mississippiensis ) have been considered a particularly suitable indicator species as th ey have been shown to bio-accumulate and biomagnify contaminants to levels equal to or greater than reported in birds and mammals (Crain and Guillette, 1998). Howeve r, an understanding of alligator thyroid function is limited as the principal data available is in relation to OC exposure (Gunderson et al., 2002; Crain et al., 1998). Crain et al (1998) noted a negative relationship with both plasma T3 and T4 concentrations and body size among male and

PAGE 52

40 female lake Woodruff (low OC ) animals. However, a general lack of correlation between plasma TH concentrations, sex and body size was observed in sub-adult alligators from both lakes Apopka and Okeechobee (Crain et al., 1998). The author’s suggested that these data may potentially reflect altered re productive potential in th ese animals, as THs cooperatively regulate the repr oductive activities of verteb rates (Crain et al., 1998). Gunderson et al. (2002) and Hewitt et al. (2002) reported significantly higher plasma T4 concentrations among sub-adul t alligators exposed to inte rmediate OC contaminant concentrations versus animals from either high and low OC environments. In addition, the author’s observed no relationship between body size and plasma TH concentrations. The author’s suggested that the observed re ductions in plasma T4 concentrations from animals located at the site of high OC contamination may be related to OC competition with TH for binding proteins as well as el evation of UDP-GT enzymatic activity which induces T4 glucuronidation and s ubsequent biliary TH excretion. Alterations in thyroid regul ation in relation to OC expos ure have been reported to cause reductions in growth. Red Eared Slider ( Trachemys elegans ) eggs topically treated with trans -Nonachlor significantly lost mass in comparison with controls (Willingham, 2001). The author suggested that the reduction in mass may be the result of a temporal, hyperthyroid state in which yolk reserves we re utilized more quickly, thus reducing overall mass. A significant ne gative relationship was repor ted between Artic Glaucous Gull ( Larus hyperboreous ) hatchling growth chick a nd increasing adult OC serum concentrations of HCB, oxychlordane, p,p’DDE, and several PCBs (Bustnes et al., 2005). Juvenile Nile Tilapia ( Oreochromis niloticus ) exposed to aqueous dieldrin for 30 days demonstrated significant reductions in we ight in comparison with controls (Lamai et

PAGE 53

41 al., 1999). In addition, Blanar et al. (2005) reported that juvenile Artic Charr ( Salvelinus alpinus ) orally dosed with toxaphene demonstr ated decreased growth and overall body condition (k). These reports suggest the poten tial direct (i.e., feeding, injection, aqueous OC exposure) and indirect (i.e., reduced parent al fitness due to OC exposure) influences that OC contaminants may have to influe nce growth among several oviparous species. Several authors have reported altered al ligator thyroid functio n in relation to OC exposure (Hewitt et al., 2002; Crain et al., 1998). In addition, controlled treatment studies utilizing several OC contaminants reporte d both modified t hyroid regulation and subsequent growth reductions. These data sugg est that OC exposure may be related to the observed reductions in alligator embryo and hatchling growth from OC contaminated sites in central Florida (R auschenberger, 2004; Wiebe et al., 2001; Gross et al., 1994). However, researchers must be keenly aw are of both physiologi cal (i.e., sex, age, nutritional availability, reproduction, hibern ation) and environmental factors (i.e., ambient and water temperatures and photoperi od) which have been reported to alter thyroid regulation and may complicate data interpretation regarding OC exposure and subsequent alterations in thyroid function. Therefore, a captive study providing a controlled, structured environment presents a more applicable means to test the relationship between OC exposure and subse quent alterations in hatchling thyroid function and growth. Wiebe et al. (2002) evaluated hatchling alligator thyroid regulation and growth from lakes Apopka, Griffin, and Lochloosa under captive conditions for a period of 6 months. These experimental conditions incl uded: a restricted phot operiod (12D:12N), controlled ambient and water temperatures, ad-libitum feeding twice a week, and

PAGE 54

42 restricted number of animals per enclosure. Egg viability rates did not differ among sites. However, lake Apopka hatchlings demonstrat ed a significantly hi gher growth rate and plasma TH concentrations in comparison with lakes Griffin and Lochlooosa. These data suggest that lake Apopka hatchlings dem onstrated a hyperthyroid secretory pattern resulting in enhanced hatchling growth in re lation to exposure to high OC concentrations. OC contaminants, due to their structural sim ilarity with THs, have been predominantly suggested to reduce TH systemic availa bility by competing for binding proteins. Therefore, these conflicting data suggest a ne ed to compare hatchling thyroid regulation and growth among several sites with high OC concentrations to provide further insight (i.e., OC exposure versus site-specific variab les) into the observed hyperthyroid secretory pattern and accelerated growth rate observed in lake Apopka hatchlings. Therefore, a captive hatchling growth study was undertaken utilizing anim als from lakes Apopka and Griffin as well as lake Orange (a site of low OC concentrations) and Emerelda Marsh Conservation Area (Area #7) (a site of high OC concentratio ns) to assess if in-ovo exposure to high concentrations of OC c ontaminants elicits a hyperthyroid secretory pattern that accelerates ha tchling alligator growth. Materials and Methods Egg Collection, Evaluation and Incubation Clutches (n=10/site) were collected from lakes Apopka (N 28 35’, W 81 39’), Griffin (N 28 53’, W 81 46’), and Orange (N 29 30’, W 82 13’) as well as Emerelda Marsh Conservation Area (Area # 7)(N 28 55’, W 81 47’). Nests were located by aerial (helicopter) and ground (air boat) surveys. Clutches were collected and transported in their original nesting substrate. To provi de proper positioning for subsequent artificial incubation, a black mark was placed on top of each egg to indicate the original egg

PAGE 55

43 orientation in the nest. Eggs were evaluate d utilizing a bright li ght candling procedure (Lyon Electric, Chula Vista, CA, USA) in or der to observe the pr esence/absence of a calcium rich band (an indicator of developi ng embryos) encircling the midsection of each egg. Each clutch was evaluated by the followi ng measures: 1) clutch weight (Kg), 2) fecundity (total number of eggs in clutch ), 3) number of banded eggs (number of currently viable eggs in clutch), 4) number of unbanded eggs (number of eggs with no band which represents early embryonic mortality or lack of fertilization), and 5) number of damaged eggs (eggs that were cracked and leaking due to nest pr edators or collection error). Yolk was collected from one viable e gg per clutch to assess clutch age (Ferguson, 1985) as well as identify and quantify lipophilic OC pesticide concentrations. Following the initial clutch evaluation, the remaining ba nded, viable eggs from each clutch were transferred to an incubation pan (18.5” x 14” x 7”) containing moist sphagnum moss substrate. Clutches were maintained in an artificial incubation bu ilding (13’ x 11’ x 7.5’) at ambient temperatures of 31.5 C 1 C and 95% relative humidit y. Individual clutch viability (total number of hatchlings / total number of eggs collected) was assessed at the completion of hatching. Upon hatching, extern al morphometrics including: total length (mm), snout-vent length (mm), and head leng th (mm) (Wildlife Supply Co., Saginaw, MI, USA; Mitutoya Calipers, Japan) and we ight (g) (Ohaus, Inc., Pine Brook, NJ, USA) were collected on each animal. In additi on, a unique Monel web tag (National Band and Tag Co., Newport, KY, USA) was pr ovided to allow for individual animal identification. Clutch Selection Three to five clutches per site were sele cted, based upon specific selection criteria for this study. Clutch selecti on criteria included: 1) The cl utch must have at least 15

PAGE 56

44 hatchlings (as per the sample numbers required to satisfy the goals of the study), and 2) Clutches were selected based on site mean yolk OC pesticide concentrations among the four principal OCs (p,p’-DDE, dieldrin, ch lordane, toxaphene) (as variance in OC concentrations among sites limits the ability to test the direct effects of OCs on hatchling alligator growth). Hatchlings (n=15) were randomly selected fr om each study-related clutch. Prior to the studies onset, hatchlings (n =3/clutch) from all sites were sacrificed in order to establish baseline values of free and total T4 plasma concentrations, thyroid weight (g) (a suggested indica tor of thyroid activity), and liv er weight (g) (a suggested indicator of extrathyroidal conversions of THs) (McNa bb, 2004; Zhou et al., 1999). All remaining animals selected for this study received a corresponding microchip (Biomark, Inc., Boise, ID, USA) at the base of th e tail utilizing a trocar delivery system. Animal Maintenance Hatchlings (n=12) per clutch were main tained for a period of eight months. Each clutch was housed in a fiberglass tank (4’ x 2’ x 2’) (Rowland Fiberglass, Ingleside, TX, USA) with an aquarium heater and heat la mp to maintain uniform ambient and water temperatures. All clutches were fed a commer cial alligator diet (Burris Mill and Feed, Franklinton, LA, USA) ad libitum twice a week. Hatchling Morphometrics and Tissue Sampling Hatchlings were measured once a mont h for a period of eight months. These measurements include: total length (mm), s nout-vent length (mm), head length (mm) SVL, and weight (g). In a ddition, a 1.5 mL blood sample was taken from the cranial sinus. Whole blood was centrifuged at 1000 x g for 10 minutes. Plasma was aliquoted into several cryogenic vials (2 mL) and frozen at -80 C. Following the initial sampling date (Oct 2004), a subset of hatchlings (n=3 / clutch) from all sites were sacrificed on a

PAGE 57

45 quarterly schedule (Nov, Jan, Apr) to allow fo r a time series evaluation of thyroid and liver activity as it relates to hatchling morpho metrics and circulating free and total plasma T4 concentrations. Sacrificed animals were selected by random number generation to avoid researcher bias. Plasma Thyroid Hormone Validation Pr ocedures (Total and Free Thyroxine) Plasma samples from alligator hatchlings were analyzed for total thyroxine (TT4) and free thyroxine (FT4) using commerc ially available radioimmunoassay (RIA) procedures. The TT4 and FT4 analyses each utilized a monoclonal solid phase radioimmunoassay component system (MP Bi omedicals Costa Mesa, CA). For the TT4 analysis, samples (50 l) were assayed directly as per the component system instructions. For the FT4 analysis, sample (25 ul) were anal yzed as per the provide instructions. RIA analyses utilized iodinated (125I) ligand (L-thyroxine) and an tibody coated tubes. Each sample was analyzed in duplicate for both TT4, and FT4. Standard curves were prepared in buffer with known amounts of radioinert T4 (0, 2, 4, 8, 12, and 20 ug/dl) or FT4 (0, 0.34, 0.64, 133, 3.27, 10.18 ng/dl). The minimum concentration distinguishable from zero was 0.81 ug/dl for TT4 and 0.025 ng/dl for FT 4 and results were listed as ng/ml for TT4 and pg/ml for FT4. Cross-reactivities of the TT4 antiserum were; 30.9 % for Dthyroxine; 1.0% for 3,3,5 triiodo-thyronine ; and <0.1% for3,5-diodo-thyronine, 3,5diodo-tyrosine, 3-ido-tyrosine and phenytoin. Cross-reactivities of the FT4 antiserum were, 91.05 % for (D-thyroxine), 7.92% fo r 3,3,5-triiodo-rev-thyronine, 3.05 % for 3,3,5triiodo-thyronine, <1.0% for 3,3-diodo-thyron ine, and <0.1% for 3,5-diodo-tyrosine, 3iodo-tryosine, 5,5-diphenylhydantoin, sodium salicylate, acetylsalicylic acid and phenylbutazone. A pooled sample (approxima tely 550 ng/ml TT4 and 480 pg/ml FT4 was assayed serially in 10, 20, 30, 40, and 50 l volumes for Free-T4 and in 5, 10, 15, 20

PAGE 58

46 and 25 ul volumes for TT4. The resulting inhibiti on curves were parallel to the respective standard curve, with the test s for homogeneity of regression indicating that the curves did not differ. Further characterization of the assays involved measurement of known amounts (0, 2, 4, 8, 12, and 20 ug/dl) of TT4 in 25ul plasma or (0, 0.34, 0.64, 133, 3.27, 10.18 ng/dl) of TT4 in 50 of plasma. For TT4, mass recoveries were estimated as: Y=0.16 + 0.97X, R2=0.9018; and for free-T4: Y= 0.015 + 0.96X, R2=0.8814 (Y= amount of TT4, FT4 measured; X= amount ofTT 4, FT4 added). Interassay and intrassay coefficients of variation were 9.2 and 8.7 % respectively for plasma TT4, and 10.3 and 8.7% respectively for plasma FT4. Free T4 (FT4) Assay Procedures FT4 (representing <1% of av ailable T4) is considered the most biologically available form of thyroxine for cellular interaction. Plasma (50 L in duplicate) was added to solid-phase coated count tubes (MP Biomedicals, Costa Mesa, CA). 1.0 mL of 125I free T4 tracer was added to each tube. Tubes we re vortexed (< 10 seconds) and incubated in an IR-Autoflow C02 water-jacketed incubato r (Nuare, Plymouth, MA, USA) at 37 1 C for a period of 90 minutes. Contents of tubes were decanted and 1 mL of distilled water was added / decanted to rinse each tube Tubes were counted on a LKB-Wallac 1282 CompuGamma gamma counter (Per kinElmer, Boston, MA, USA). Total T4 (TT4) Assay Procedures TT4 (representing >99% of available T4) is reversibly associated with several binding proteins including transthyretin, thyroglobulin, and albumen. Plasma (25 L in duplicate) was added to solid-phase coated count tubes (MP Biomedicals, Costa Mesa, CA). 1.0 mL of 125I total T4 tracer was added to each tube. Tubes were vortexed (< 10 seconds) and incubated at room temperature (18 to 25 C) for a period of 60 minutes.

PAGE 59

47 Contents of tubes were decanted and counted on a LKB-Wallac 1282 CompuGamma gamma counter (PerkinElmer, Boston, MA, USA). Analysis of Chlorinated Analyt es from Alligator Egg Yolks Analytical grade standards for the follo wing compounds were purchased from the sources indicated: aldrin, -BHC, -BHC, lindane, -BHC, p,p’ -DDD, p,p’ -DDE, p,p’ DDT, dieldrin, endosulfan, endosulfan II, endos ulfan sulfate, endri n, endrin aldehyde, endrin ketone, heptachlor, heptachlor epoxide, hexachlorobenzene, kepone, methoxychlor, mirex, cis -nonachlor, and trans -nonachlor from Ultra Scientific (Kingstown, RI); cis -chlordane, and trans -chlordane from Supelco (Bellefonte, PA); oxychlordane from Chem Service, Inc. (West Chester, PA); o,p’DDD, o,p’DDE, o,p’DDT from Accustandard (New Haven, CT ); and toxaphene from Restek Corp. (Bellefonte, PA). A ll reagents were analytical grade unless otherwise indicated. Water was doubly distilled and deionized. Alligator egg yolk samples were analyzed for OCP content using methods modified from Holstege et al. [1] and Schenck et al. [2]. For extraction, a 2-g tissue sample was homogenized with ~1 g of s odium sulfate and 8 mL of et hyl acetate. The supernatant was decanted and filtered though a Bchner f unnel lined with Whatman #4 filter paper and filled to a depth of 1.25 cm with sodi um sulfate. The homogenate was extracted twice with the filtrates collected together. Th e combined filtrate was first concentrated to a volume of ~2 mL by rotary evaporation, then further con centrated until solvent-free under a stream of dry nitrogen. The residue was reconstituted in 2 mL of acetonitrile. After vortexing (30 s) the supernatant wa s applied to a C18 SPE cartridge (preconditioned with 3 mL of acetonitrile; Ag ilent Technologies, Wi lmington, DE) and was allowed to pass under gravity. This procedur e was repeated twice with the combined

PAGE 60

48 eluent collected in a culture tube. After the last addition, the cartridge was rinsed with 1 mL of acetonitrile which was also collected. The sample was then applied to a 0.5 g NH2 SPE cartridge (Varian, Inc., Harbor City, CA), was allowed to pass under gravity, and was collected in a graduated conical tube. Th e cartridge was rinsed with an additional 1 mL portion of acetonitrile which was also collected. The combined eluents were concentrated under a stream of dry nitrogen to a volume of 300 L and transferred to a GC vial for analysis. Analysis of the samples was perfor med using a Hewlett Packard HP-6890 gas chromatograph (Wilmington, DE) with split/splitl ess inlet operated in splitless mode. The analytes were introduced in a 1 L injection and separated across the HP-5MS column (30 m x 0.25 mm; 0.25 m film thickness; J & W Scientific, Inc., Folsom, CA) under a temperature program that began at 60 C, incr eased at 10 C/min to 270 C, was held for 5 min, then increased at 25 C/min to 300 C and was held for 5 min. Detection utilized an HP 5973 mass spectrometer in electron imp act mode. Identificatio n for all analytes and quantitation for toxaphene, was conducted in full scan mode, where all ions are monitored. To improve sensitivity, selected ion monitoring was used for the quantitation for all other analytes, except kepone. The a bove program was used as a screening tool for kepone which does not optimally extract with most organochlorines. Samples found to contain kepone would be reextracted and analyzed specifically for this compound. For quantitation, a five-point standard curve was prepared for each analyte (R2 0.995). Fresh curves were analyzed with each se t of twenty samples. Each standard and sample was fortified to contain a deuterat ed internal standard, 5 L of US-108 (120 g/mL; Ultra Scientific), added just prior to analysis. All samples also contained a

PAGE 61

49 surrogate, 2 g/mL of tetr achloroxylene (Ultra Scientif ic) added at homogenization. Duplicate quality control samples were prepar ed and analyzed with every twenty samples (typically at a level of 1.00 or 2.50 g/mL of -BHC, heptachlor, aldr in, dieldrin, endrin, and p,p’ -DDT) with an acceptable recovery rangi ng from 70 – 130%. Repeated analyses were conducted as allowed by matrix inte rferences and sample availability. Statistics Initial RIA data was analyzed and fit f our parameters logistic curve utilizing Beckman EIA/RIA ImmunoFit software (Fullerto n, CA). All statistics were performed on SAS version 9.1 for windows (SAS Institute, Inc., Cary, NC, USA). PROC GLM procedures including Tukey multiple comparison analysis was utilized to detect differences (p < .05) among hatchling extern al morphometrics, plasma thyroid hormone concentrations and OC contaminant concentrat ions between and within sites. Correlative analysis among growth rates, plasma thyr oid hormone concentration rates and OC contaminated concentrations was performed with PROC REG procedures (p < .05). Differences in thyroid and liver somatic indices were analyzed by the Wilcoxon Rank Sum Test in which the Kruskal-Wallis Test was utilized to determine significant differences among and within sites (p< .05). Results Clutch and Organochlorine Contaminant Parameters Clutches (n=40) were co llected from lakes Apopka (n=10), Griffin (n=10), Orange (n=12) and Emerelda Marsh Conser vation Area (n=8). Two principal clutch parameters were utilized to select clutches for the current study: fecundity and viability. A summary of all clutches co llected demonstrated site differences among both clutch fecundity and viability (p < .05) (Fig. 2-2) In specific, lake Apopka clutches had

PAGE 62

50 significantly reduced clutch viability in comp arison with the remaining sites. Selected clutches for the current study demonstrated si milar trends in clutch fecundity (p < .05) (Fig. 2-3). However, no differences were obser ved in clutch viability between sites for these select clutches (Fig. 24). Hatchling OC (specifically: total chlordane, total DDE, dieldrin, and toxaphene) exposur e was determined from a representative yolk sample per clutch. Total OC concentrations per site (i.e., all clutches and growth clutches) concentrations were distributed as follows: (EM>AP>GR>OR) (Fig. 2-4). Hatchling Growth Rates Hatchling growth morphometrics were m onitored monthly for a period of eight months. Multiple comparative analyses am ong sites demonstrated that lake Griffin hatchlings grew significantly larger in total length, snout -vent length, head length and weight (Fig. 2-5) (p < .05). Clutches within each site demonstrated similar trends in total length (Fig. 2-6), snout-vent length (Fig. 2-7) head length (Fig. 2-8) and weight (Fig. 29). Hatchlings (n=3/clutch/site ) sacrificed on a quarterly sc hedule to compare thyroid and liver weights to growth over time demonstr ated similar trends among (Figs. 2-10) and within sites in total length (Fig. 2-11), snout-vent length (F ig. 2-12), head length (Fig. 213), weight (Fig. 2-14), t hyroid weight (Fig. 2-15), and live r weight (Fig. 2-16) (p < .05). No differences were observed in thyroid somatic indices among sites (Table 2-5). However, significant temporal differences were observed in thyr oid somatic indices within sites (p< .05) (Table 2-6) Liver soma tic indices demonstrated several temporal significant differences among and within si tes (p < .05) (Tables 2-5 and 2-7). Mean growth rates were tabulated among and within sites to examine hatchling growth per day including: total length/day (Tab le 2-1), snout-vent length/day (Table 2-2), head length/day (Table 2-3) and weight/day (Table 2-4). Multiple comparative analyses

PAGE 63

51 of growth rates among clutches demonstrated several significant differences among total length, snout-vent length and head length rates (p < .05). However, analysis of growth rates among sites again demonstrated that lake Griffin hatchlings grew larger than the other sites (p < .05). A correlative analysis in which all clutches were independent of site demonstrated no differences in growth rates among all sampling dates (p < .05). These data present several isolated differences in growth rates within clutches which can potentially be attributed to inter-clutch va riability. The dominant inference taken from both correlative and multiple comparative anal yses continues to indicate no significant differences in hatchling growth among sites. Thyroid Hormones, Growth and Organochlorine Contaminants Thyroid hormones (specifically: total (TT4) and free (FT4) thyroxine) were utilized as bio-indicators of hatchling alligator growth. Mult iple comparison analysis of plasma TT4 concentrations over time dem onstrated an asynchronous secretory pattern among (Fig. 17) and within sites (Figs. 18). Similarly, plasma FT4 concentrations over time demonstrated an asynchronous secretory pattern among (Fig. 17) and within sites (Fig. 19). No significant alterations in either TT4 or FT4 plasma concentrations were observed over the eight month sampling peri od. In addition, no pair ed relationship was observed among either growth rates or any growth parameter during specific sampling dates and plasma thyroid hormone concentra tions. However, a review of monthly mean hatchling growth parameters and plasma t hyroid hormone concentr ation distributions demonstrate a temporal relationship between TH secretion and subsequent hatchling growth. Modifications in growth and plasma thyr oid hormone concentrations have been reported in association with OC contaminan t exposure among several species (Bustnes et

PAGE 64

52 al., 2005, Willingham, 2001). To examine the potential interactiv e nature of these experimental variables, a corre lative analysis was performe d utilizing hatchling growth rates, thyroid hormone rates and the four principal OC contaminants (i.e., total chlordane, total DDE, dieldrin and toxaphene). No si gnificant correlative relationships were observed (Table 2-8). Discussion The objective of the current study was to determine if in-ovo exposure to high concentrations of OC contaminants elic its a hyperthyroid secr etory pattern that accelerates hatchling alligator growth. This assessment was based on several reports indicating both alterations in thyroid function and/or subsequent grow th in relation to OC exposure under field and experimental c onditions. Rauschenberger (2004) reported alterations in embryonic alligator growth a nd development in relation to maternal OC exposure. In addition, several reports have related OC exposure to modified alligator thyroid histological parameters and regul ation (Gunderson et al., 2002; Hewitt et al., 2002; Crain et al., 1998). Wiebe et al., ( 2002) reported both h yperthyroid secretory patterns of THs and subsequent accelerated growth among hatchlings from high OC contaminated environments. In addition, c ontrolled treatment st udies utilizing OC contaminants (i.e., total chlordane, total DDE, dieldrin, and toxaphene) demonstrated altered growth in relation to OC contaminan t exposure (Blanar et al ., 2005; Bustnes et al., 2005; Willingham, 2001; Lamai et al., 1999). These combined data suggest an interrelation between OC exposure a nd modification of growth a nd growth-regulating factors such as thyroid hormones. However, the current study demonstrated no significant differences in hatchling alligator thyroid regulation or growth rates in relation to in-ovo OC exposure. Results of the current study may be attributed to the 1) inability to utilize

PAGE 65

53 clutches with low viability, 2) additional growth-regulating product( s) other than or integrated with THs that regulate hatchling alligator growth, or 3) non-OC contaminant influences including: maternal size a nd/or habitat and nutritional quality. Reduced alligator clutch viability has been reported within sites of intermediate to high concentrations of OC contaminants in central Florida (Rausc henberger, 2004; Gross, 1994). These data demonstrate growth retarda tion and subsequent mortality during both early and late embryonic development, and among hatchlings from high OC environments. However, low viability clutches were excluded from the current study due to clutch selection requirement s: study clutches required at least fifteen hatchlings in order to test the current studi es hypothesis. The elimination of these clutches from the current study likely removed hatchlings with an increased potent ial to demonstrate irregularities in growth and developmental regulation in rela tion to OCs and/or additional environmental stressors. Though differences in OC contaminant concentrations were observed (p<.05), clutches ut ilized in the current study demonstrated no significant differences in viability between sites. Several authors have reported alterations in alligator th yroid regulation in relation to OC contaminant exposure (Gunderson et al ., 2002; Hewitt et al., 2002; Crain, 1998). In addition, these reports stated that alligat ors from OC contaminated environments demonstrated a general lack of correlation between plasma TH c oncentrations, sex and body size. However, these studies were not ab le to eliminate several physiological and environmental factors (i.e., age, sex, photope riod, water and ambient temperatures and food availability) reported to influence thyroid regulation. In addition, these studies examined the relationship between OC cont aminant exposure and hatchling growth as

PAGE 66

54 well as plasma TH concentrations utilizing a si ngle point in time sampling procedure. As thyroid hormones have been reported to have a pulsatile secretory pattern, multiple sampling over time would appear to be pert inent when trying to relate plasma TH concentrations and growth to the hyper-variable influences of environmental contaminant exposure. Wiebe et al. (2002) correlated bot h plasma T3 and T4 concentrations with growth over time among hatchling alligators from sites of varying OC contamination under captive conditions. To more directly examine the relationship between OC exposure and alterations in hatchling grow th and thyroid regula tion, captive conditions were designed to limit the influence of phys iological and environmental influences on thyroid regulation. These condi tions included: 12L:12D phot operiod; constant ambient and water temperatures; restri cted pod size to avoid stressf ul overcrowding; documented hatchling age; and ad-libitum food availability. Results from the 2002 study demonstrated that hatchlings from high OC environments demonstrated a hyperthyroid TH secretory pattern and accelerated growth. Utilizing the comparable cap tive conditions, a temporal relationship between plasma TH concentra tions and hatchling alligator growth was observed over time in the current study. Add itionally, thyroid and liver weights as well as liver somatic indices were found to be repr esentative biomarkers of hatchling growth among and within sites over time. However, no relationship was observed between OC exposure and hatchling alligato r growth rates or plasma t hyroid hormone concentrations among or within sites over time. Therefore, future research may require examination of additional growth regulating endocrine pathwa ys when assessing th e potential influence of OC contaminant exposure on hatc hling alligator growth regulation.

PAGE 67

55 These conflicting data su ggest that THs may not be the principal growth regulating hormone influenced by OC contam inant exposure. McNabb (2000) noted that THs act permissively or indirectly, in con cert with the principal growth regulators: growth hormone (GH) and insulin-like growth factor I (IGF-I). In addition, THs have been reported to participate in highl y integrated growth regulation among both somatotropic and corticotropic axis’ (Khn et al., 2005, Kobel et al ., 2001, Elsey et al., 1990). Growth hormone (GH) is an essentia l regulator of growth with complex metabolic functions (Bjornss on et al., 2002). Pituitary GH secr etion reported to be under the dual control by two neuropeptides from the hypothalamus: GH releasing hormone (GHRH) which stimulates GH release and soma tostatin (SRIH) which has an inhibitory action (Renaville et al., 2002). However, plasma GH concentrations have been demonstrated to be influenced by a vari ety of hormones, growth factors, and environmental influences (Fig. 2-20). The anabolic and growth–promoting effects of GH are to a large extent mediated by the stimul ation and expression of insulin-like growth factor I (IGF-I) in the liver a nd peripheral tissues (Sjgren et al., 1999). The interactive (i.e., local and systemic) regulation demonstr ated between GH and IGF-I is known as the “dual effector theory of acti on.” (Bjornsson et al., 2002). Se veral reports have examined plasma IGF-I concentrations among reptilian models (Guillette et al., 1996, Crain et al., 1995, Crain et al., 1995). Thes e reports indicated that increased plasma IGF-I concentrations were coincident with e gg formation as oviparous species must compartmentalize growth-promoting substances and nutrients into the yolk and albumen of eggs (Guillette et al., 1996). In addition to IGF-I, maternal transfer of growth-

PAGE 68

56 regulating substances (i.e., GH, TH) appears to be critical for embryo development with implications on future hatchling growth and su rvival (Greenblatt et al., 1989). Therefore, maternal quality, which encapsulates animal health in relation to exposure to environmental stressors including OCs, continue s to appear to be a dominant regulatory factor in clutch growth and survival. OC contaminant exposure has been re ported to influence reproductive and developmental parameters among adult and juve nile alligators (Rauschenberger et al., 2004, Gross et al., 1994). These data suggest an integrated relationship between adult alligator exposure to multiple environmental stressors (i.e., OCs, water quality, nutritional quality) and subsequent alterati ons in clutch and ha tchling quality. Under captive conditions, the current study demonstrated that THs may serve as indicators of hatchling alligator growth utilizing multiple sampling procedures over time. However, no relationship was observed between OC expos ure and hatchling al ligator growth and thyroid regulation in the current study. These data suggest th at THs may not represent the principal endocrine pathway affected by OC contaminant exposure. Therefore, the null hypothesis which states that there will be no effects of in-ovo OC exposure on hatchling alligator growth or thyroid regulation must be accepte d. Future research efforts examining the relationship between hatchli ng alligator growth and OC exposure should incorporate an integrated eval uation of multiple endocrine path ways (i.e., GH, IGF-I, TH, corticoids), utilize multiple sampling techniques over time, and, when possible, limit the influence of reported physiological and envi ronmental parameters on growth regulation (Scollon et al., 2004).

PAGE 69

57 In addition to OC exposure, anthropoge nic habitat modifications have been suggested as potential influe ntial factors in the observed modifications in alligator reproductive and growth parameters among OC cont aminated sites. Schelske et al. (2005) provides details a chronology of habitat modi fication in the upper Ocklawaha river basin including: constructi on of the Beauclair canal and exte nsive levee systems, extensive citrus and muck farming operations, as well as municipal sewage discharge. These habitat modifications and the subsequent “back pumping” of phosphorous from muck farming operations is credited with creating marginal habitat with an extensive changes of both flora and fauna among this river system (Sch elske et al., 2005). Several reports have investigated the influence(s) of habitat m odification and other non-OC related influences on alligator clutch viability among OC cont aminated sites. Rauschenberger (2004) examined the incidence of alligator nutri tional deficiencies specifically: thiamine (Vitamin B1) deficiency, which has been sugge sted to increase embryonic mortality in relation to OC contaminant exposure. Resu lts from clutches collected from OC contaminated environments demonstrated that thiamine deficiency may be involved in decreased clutch viability. In addition, Mason (1995) suggested that changes in available nesting vegetation had the potential to redu ce alligator clutch viability through reduction in insulation as well as inappropriate moistu re content. Though alte rations in alligator reproductive and growth quality have been asso ciated with OC contaminant exposure, the tremendous influence of habitat modificati ons (i.e., habitat qual ity, water quality, nonindigenous species) on alligator growth, re production and surviv al should not be discounted.

PAGE 70

58 Though no significant differences were obser ved in hatchling alligator growth or thyroid regulation in relation to in-ovo OC exposure, these data do not discount the potential for growth alterati ons among wild alligator populat ions in OC contaminated environments. Significant variability in al ligator reproductive a nd growth regulation continues to be observed in relation to OC contaminated environments (Rauschenberger, 2004, Guillette et al., 1999, Gross et al., 1994). These data include: 1) gonadal modifications such as: altered plasma sex steroid concentrations and histological abnormalities, 2) increased fecundity 3) increased incidence of early and late embryonic mortality, as well as 4) growth disparities between and within OC contaminated sites versus control sites. In order to better re late the observed reductions in alligator reproductive and clutch qualities to OC exposur e, Rauschenberger et al. (2004) orally dosed captive adult alligators in reproductiv e groups (1 male:1 female) with an ecorelevant OC contaminant mixture. This expe rimental mixture was re presentative of OC isoforms concentrations anal ytically identified among yolks from OC environments in central Florida. Experimental clutches dem onstrated comparable reductions in clutch viability specifically: increased incidence of early embryonic mortality, which has been observed in wild clutches from OC environments. Data from the current study demonstrates that THs can be utilized as a bioindicator of hatchling alligator growth under captive conditions. Therefore, experimental control of established physio logical and environmental influences on thyroid regulation allowed for a more through examination of not only hatchling alligator growth but, the potentia l inter-relation between OC cont aminant exposure and subsequent growth and thyroid regulation. These data re present a more direct examination of the

PAGE 71

59 inter-relationship between OC exposure and altered hatch ling alligator growth and thyroid regulation. Though previ ous work reported a hyperthyro id secretory pattern and accelerated growth in hatchli ng alligators from high OC environments, the current study demonstrated no relationship between OC expos ure and subsequent alterations in growth or thyroid regulation(Table 2-8) These data suggest that ha tchling alligator growth is regulated by an integrated e ndocrine network (i.e., GH, IGF -I, corticoids) in which THs may not be the principal regulatory agent. In addition, the inability to utilize clutches of lower viability from OC contaminated site s may have restricted the incidence of observing growth and developmental alterations. Alligator reproductive and growth alte rations continue to be reported in association with OC contaminated sites. Previous data reporte d hyperthyroid secretory patterns and accelerated hatchling alligator growth in association with high OC contaminants. However, no relationship was observed between OC contaminant exposure and hatchling growth or t hyroid regulation in the cu rrent study. Though THs were deemed useful for monitoring hatchling alliga tor growth, they do not appear to be the principal growth regulatory factor. Future examination of both i ndividual as well as integrated regulatory relations hips between growth-regulating hormones / growth factors (i.e., GH, IGF-I, TH, corticoids) may prove more useful when trying to relate OC contaminant exposure to observed alterations in alligator growth. In conclusion, there has been a singular focus in associating the obser ved reductions in alli gator reproductive and growth parameters with OC contaminant expos ure. However, the significant influence(s) of environmental factors (i.e ., habitat modification as we ll as water and nutritional

PAGE 72

60 quality) should not be discounted when evalua ting alligator physiol ogy in relation to OC contaminant exposure.

PAGE 73

61 Figure 2-1.Graphical interpre tation of thyroid hormone biosynthesis. Taken from www.neurosci.pharm.otoledo.edu/MBC3320/thyroid.htm (11/04/05).

PAGE 74

62 Figure 2-2. Clutch fecundity and clutch vi ability (site means). Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p <.05). Viability GROREMAP (%) 0 20 40 60 80 100 a a a b Fecundity GR OR EM AP 0 10 20 30 40 50 60 a b a a

PAGE 75

63 Figure 2-3. Clutch fecundity and clutch viab ility (current study). Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p <.05). Viability GR OR EM AP (%) 0 20 40 60 80 100 a a a aFecundit y GR OR EM AP 0 10 20 30 40 50 60 a b a a

PAGE 76

64 Figure 2-4. Yolk OC concentrations. site means (a) and current study (b). Significant differences among sites were determ ined by Tukey Multiple Comparison Analysis (p <.05). YOLK OCP Concentrations (Site Means) ORGRAPEM (ng / g) 1 10 100 1000 10000 Total Chlordane DDTx Dieldrin Toxaphene a a b a b b a b b b a a a a a b b Yolk OCP Concentrations (Current Study) ORGRAPEM (ng / g) 1 10 100 1000 10000 Total Chlordane DDTx Dieldrin Toxaphene a b

PAGE 77

65 Figure 2-5. Hatchling alligator growth pa rameters among sites over time. Significant differences among sites were determ ined by Tukey Multiple Comparison Analysis(p < .05). Total Length SeptOctNovDecJanFebMarAprMay TL (mm) 0 100 200 300 400 500 600 GR OR EM AP a a a a a a b b b a b b b a a b b b a a b b b a a b b b a a b a b b a a b b b a a b a b b Head SeptOctNovDecJanFebMarAprMay Head (mm) 0 20 40 60 80 GR OR EM AP a a a a a b bb a b b b a b b b a b b b a a b a b b a a b b b a a b b a b a a b b a b Weight SeptOctNovDecJanFebMarAprMay Weight (g) 0 100 200 300 400 500 600 GR OR EM AP a b bb a b b a b a b b b a b b b a b b b a b b b a a b a b b a a b a b b a a b b a b Snout-Vent Length SeptOctNovDecJanFebMarAprMay SVL (mm) 0 50 100 150 200 250 300 GR OR EM AP a a b a b b a b b b a b b b a b bb a b b b a a b a b b a b b b a a b b b a a b b a b

PAGE 78

66 Figure 2-6. Hatchling alligator total length (mm) within sites over time. Significant differences among sites were determ ined by Tukey Multiple Comparison Analysis (p < .05). Griffin SeptOctNovDecJanFebMarAprMay TL (mm) 0 100 200 300 400 500 600 700 GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D a a b b c c b a b a a a b b a b a a a b b a b a a a b a a a a a b c c a a b a b c a a a a a a b b a b a b a a b b a a a b Orange SeptOctNovDecJanFebMarAprMay TL (mm) 0 100 200 300 400 500 600 700 OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5 a a a a a b c bb a a c a b a b a b b b b b a a b c d b c a b d a b c b c a b a b c a c a b c a b c a b c a a b a b a b b a b a a b a b Emerelda SeptOctNovDecJanFebMarAprMay TL (mm) 0 100 200 300 400 500 600 EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 a a b b c a a a a b a b a c b a b a b a c b b b a c b c b c a b a b a b a b a b a b b b a a a a a a a a a a Apopka SeptOctNovDecJanFebMarAprMay TL (mm) 0 100 200 300 400 500 600 700 AP-04-10 AP-04-W2 AP-04-W10 a b a a a b b a b a b a b a b a b a b a a a a a a a a a a c

PAGE 79

67 Figure 2-7. Hatchling alligator snout-vent length (mm) within sites over time. Significant differences among sites were determ ined by Tukey Multiple Comparison Analysis (p < .05). Orange SeptOctNovDecJanFebMarAprMay SVL (mm) 0 50 100 150 200 250 300 350 OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5 a a b b c c b c a b bb b a b b a b a b a b a b b b a b b b a b a a b c a b b c a a b c b c b c c a b a b a b a b a b b a b a b Griffin SeptOctNovDecJanFebMarAprMay SVL (mm) 0 50 100 150 200 250 300 350 GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D a a b b c c c a a a a a a a a a a a a a a a a b b a b a b a a a b b a a b a a a a a a b b a b a b a a b c a b c c a b a Emerelda SeptOctNovDecJanFebMarAprMay SVL (mm) 0 50 100 150 200 250 300 EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 a b b c d a a aa b a b a c b b a b b b c b a c b b c a b c a c a b b c a b a b a b b a a a a a a a a a a Apopka SeptOctNovDecJanFebMarAprMay SVL (mm) 0 50 100 150 200 250 300 350 AP-04-10 AP-04-W2 AP-04-W10 a b c a a b a a b a aa a a b b a a a a a a a a a a a a

PAGE 80

68 Figure 2-8. Hatchling alligator head length (mm) within sites over time. Significant differences among sites were determ ined by Tukey Multiple Comparison Analysis (p < .05). Griffin SeptOctNovDecJanFebMarAprMay Head (mm) 0 20 40 60 80 100 GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D a a a a a aa a a a a a a a a a a a a a a a a a a a a a a a a b b a a a b a a a a a a b b a b a a Griffin SeptOctNovDecJanFebMarAprMay Head (mm) 0 20 40 60 80 100 GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D a a a a a aa a a a a a a a a a a a a a a a a a a a a a a a a b b a a a b a a a a a a b b a b a a Emerelda SeptOctNovDecJanFebMarAprMay Head (mm) 0 20 40 60 80 EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 a a b bb b a a b a b b c a b a c b b b a c b b b a c b b c b c a c a b b c b a b a b b a a a a a a a a a a Apopka SeptOctNovDecJanFebMarAprMay Head (mm) 0 20 40 60 80 100 AP-04-10 AP-04-W2 AP-04-W10 a b c aa b a a b a a b a a b b a a b b a a a a a a a a a

PAGE 81

69 Figure 2-9. Hatchling alligat or body weight (g) within sites over time. Significant differences among sites were determ ined by Tukey Multiple Comparison Analysis (p < .05). Griffin SeptOctNovDecJanFebMarAprMay Weight (g) 0 100 200 300 400 500 600 700 800 GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D aa b b c aa a aa a b b a b a a b a b b a b a a b a b b a b a a b a b a a b a b a b b a a a a b b a b a b a a b b a b a a Orange SeptOctNovDecJanFebMarAprMay Weight (g) 0 100 200 300 400 500 600 700 OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5 a b b b b a b b b b a b b b b a a b c b c b c a a b b c b c c a a b b c b c c a a b b c b c c a a c c a b b a c a a b a b b b Emerelda SeptOctNovDecJanFebMarAprMay Weight (g) 0 100 200 300 400 500 600 EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 a a b b c d a b b b c b a c b b b a c b b b a c b b c a b b b b a b b b b a b a b a b a b a a a a a Apopka SeptOctNovDecJanFebMarAprMay Weight (g) 0 100 200 300 400 500 600 700 AP-04-10 AP-04-W2 AP-04-W10 a b c aa b a a b a a b b a a b b a a b b a a b b a a a a a a

PAGE 82

70 Figure 2-10. Hatchling alligator growth pa rameters (necropsy animals) among sites over time. Significant differences am ong sites were determined by Tukey Multiple Comparison Analysis (p < .05). Thyroid Weight SeptNovJanMarchMay Thyroid (g) 0.00 0.01 0.02 0.03 0.04 0.05 GR OR EM AP a a b b a b a a a a a a a a a a a a a a a a Liver Weight SeptNovJanMarchMay Liver (g) 0 2 4 6 8 10 12 14 GR OR EM AP a a a a a a a a a a a a a a a a a b a b a b Weight SeptNovJanMarchMay Weight (g) 0 100 200 300 400 500 600 700 GR OR EM AP a aa a a a b bb a a a a a a a a a b a b a b Snout-Vent Length SeptNovJanMarchMay SVL (mm) 0 50 100 150 200 250 300 350 GR OR EM AP a a a a a b b b aa aa a a a a a a b a b b Head Length SeptNovJanMarchMay Head (mm) 0 20 40 60 80 GR OR EM AP a a a a a aa a a a a a a a a a a a b a b b Total Length SeptNovJanMarchMay TL (mm) 0 100 200 300 400 500 600 700 GR OR EM AP a b a b a b a a b bb a a a a a a a a a b a b a b

PAGE 83

71 Figure 2-11. Hatchling alligat or total length (mm)(necropsy animals) within sites over time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Orange OR-04-12OR-04-13OR-04-BOR-04-W1OR-04-W5 TL (mm) 0 100 200 300 400 500 600 700 Sept Nov Jan Mar May a a a b b b a a a b a a b a a a a a a a b a b b a a a b a a b Emerelda EM-04-01EM-04-02EM-04-03EM-04-04EM-04-11 TL (mm) 0 100 200 300 400 500 600 700 Sept Nov Jan Mar May b a a a a a a b a b a b b a b b b b a a b b c b c c a a b a b a b b Apopka AP-04-10AP-04-W2AP-04-W10 TL (mm) 0 100 200 300 400 500 600 700 Sept Nov Jan Mar May a a b b a a b b a a a a a a a a a Griffin GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D TL (mm) 0 100 200 300 400 500 600 700 Sept Nov Jan Mar May a a b a b c b c c a a b a b b b a a a a a a a a a a a a a a a

PAGE 84

72 Figure 2-12. Hatchling alligator snout-vent length (mm)(necr opsy animals) within sites over time. Significant differences am ong sites were determined by Tukey Multiple Comparison Analysis (p < .05). Orange OR-04-12OR-04-13OR-04-BOR-04-W1OR-04-W5 SVL (mm) 0 50 100 150 200 250 300 350 Sept Nov Jan Mar May a a b a b b b a a b a b b b a a a a a a a a a a a a a a a Griffin GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D SVL (mm) 0 50 100 150 200 250 300 350 Sept Nov Jan Mar May a a b a b c b c c a a b a b b b a a a a a a a a a a a a a a a Emerelda EM-04-01EM-04-02EM-04-03EM-04-04EM-04-11 SVL (mm) 0 50 100 150 200 250 300 350 Sept Nov Jan Mar May a a b b c c d a a b a b a b b a b b b b a a b b c b c c a a a a a Apopka AP-04-10AP-04-W2AP-04-W10 SVL (mm) 0 50 100 150 200 250 300 350 Sept Nov Jan Mar May a b b a a b b a a a a a a a a a

PAGE 85

73 Figure 2-13. Hatchling alligator head lengt h (mm)(necropsy animals) within sites over time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Orange OR-04-12OR-04-13OR-04-BOR-04-W1OR-04-W5 Head (mm) 0 20 40 60 80 100 Sept Nov Jan Mar May a a b a b b a b a a a a a a b a b a b a b a b b a b a b a a a a a Emerelda EM-04-01EM-04-02EM-04-03EM-04-04EM-04-11 Head (mm) 0 20 40 60 80 100 Sept Nov Jan Mar May a b b b c a a a a a a a b b b b a a b b b b a a b a b a b b Apopka AP-04-10AP-04-W2AP-04-W10 Head (mm) 0 20 40 60 80 100 Sept Nov Jan Mar May a b b a a b b a a a a a a a a a Griffin GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D Head (mm) 0 20 40 60 80 100 Sept Nov Jan Mar May aaa b b a a a a a a a aa a a a a a a a a b a b a b b

PAGE 86

74 Figure 2-14. Hatchling alligat or body weight (g) (necropsy an imals) within sites over time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Griffin GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D Weight (g) 0 100 200 300 400 500 600 700 800 Sept Nov Jan Mar May b a a a a a a b a b b b a a a a a a a a a a a a a a a Orange OR-04-12OR-04-13OR-04-BOR-04-W1OR-04-W5 Weight (g) 0 100 200 300 400 500 600 700 800 Sept Nov Jan Mar May a a b b b b a a b a b a b b a a b a b b c c a b b b a b a a a a a Emerelda EM-04-01EM-04-02EM-04-03EM-04-04EM-04-11 Weight (g) 0 100 200 300 400 500 600 700 Sept Nov Jan Mar May a b b a a a a a a b a b b a b b b b a b b b b a a b a b a b b Apopka AP-04-10AP-04-W2AP-04-W10 Weight (g) 0 100 200 300 400 500 600 700 Sept Nov Jan Mar May a a b b a b b a a a a a a a a a

PAGE 87

75 Figure 2-15. Hatchling alligator thyroid weight (g)(necropsy animals) within sites over time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Griffin GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D Thyroid Wt. (g) 0.00 0.01 0.02 0.03 0.04 0.05 0.06 0.07 Sept Nov Jan Mar May a a b a b b b a a b a b b b a a a a a a a a a a a a a a a Orange OR-04-12OR-04-13OR-04-BOR-04-W1OR-04-W5 Thyroid Wt (g) 0.00 0.01 0.02 0.03 0.04 0.05 Sept Nov Jan Mar May a a a a a a a a a a a a a a a a a a a a a a a a a Emerelda EM-04-01EM-04-02EM-04-03EM-04-04EM-04-11 Thyroid Wt (g) 0.00 0.01 0.02 0.03 0.04 0.05 0.06 Sept Nov Jan Mar May a b b b b a a b b b b a b b b b a a b a b a b b a a a aa Apopka AP-04-10AP-04-W2AP-04-W10 Thyroid Wt (g) 0.00 0.01 0.02 0.03 0.04 0.05 0.06 Sept Nov Jan Mar May a a b b a a a a a a a a a a a a

PAGE 88

76 Figure 2-16. Hatchling alligator liver weight (g) (necropsy an imals) within sites over time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Griffin GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D Liver Wt (g) 0 2 4 6 8 10 12 14 16 18 Sept Nov Jan Mar May a a b a b b c c a a b b b b a a a a a a a a a a a a a a a Orange OR-04-12OR-04-13OR-04-BOR-04-W1OR-04-W5 Liver Wt (g) 0 2 4 6 8 10 12 14 16 18 Sept Nov Jan Mar May a a b a b b b a a b b b b a a a a a a a a a a a a a a a Emerelda EM-04-01EM-04-02EM-04-03EM-04-04EM-04-11 Liver Wt (g) 0 2 4 6 8 10 12 14 16 Sept Nov Jan Mar May a a b b c c d a a b a b a b b a a b b b b a b b b b a a a a a Apopka AP-04-10AP-04-W2AP-04-W10 Liver Wt (g) 0 2 4 6 8 10 12 14 Sept Nov Jan Mar May a a b a a a a a a a a a a a a

PAGE 89

77 Figure 2-17. Hatchling allig ator total thyroxine(ng/ml )and free thyroxine (pg/ml) plasma concentrations among sites ov er time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Total Thyroxine OctNovDecJanFebMarAprMay TT4 (ng/mL) 0 2 4 6 8 10 12 14 16 18 GR OR EM AP b c d a d b a b a b a a a a a a b b b a a a a b c a b a c a a a a a a b b Free Thyroxine OctNovDecJanFebMarAprMay FT4 (pg/mL) 0 1 2 3 4 5 6 GR OR EM AP b b b a a a b b b a b a b a a b b c a b b b a a b a b b a a a a a a a a

PAGE 90

78 Figure 2-18. Hatchling alligator total thyroxine (ng/ml) plas ma concentrations within sites over time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Griffin OctNovDecJanFebMarAprMay TT4 (ng / ml) 0 2 4 6 8 10 12 14 16 18 GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D a a a b b a a a a aa a a a a b a b a b a b a a a a a a a a a a a b a a a b a b a a aa a Apopka OctNovDecJanFebMarAprMay TT4 (ng/mL) 0 2 4 6 8 10 12 14 16 18 20 AP-04-10 AP-04-W2 AP-04-W10 a a b a a a a a a a a a a a a a a a a a a a b b Orange OctNovDecJanFebMarAprMay TT4 (ng / ml) 0 2 4 6 8 10 12 14 16 18 OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5 a a a a a a a a a a a a a a a b b a b a a b a a a a a a a a a a a a a a a a b a a a Emerelda OctNovDecJanFebMarAprMay TT4 (ng / ml) 0 2 4 6 8 10 12 14 16 18 20 22 EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 d b c c d a b a a a a a a a a a a a a a a a a a a a a a b a b a b b a a b a b a b b a a a a a a

PAGE 91

79 Figure 2-19. Hatchling alligator free thyroxine (pg/ml) plas ma concentrations within sites over time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05). Griffin OctNovDecJanFebMarAprMay FT4 (pg/mL) 0 1 2 3 4 5 6 7 GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D a b b a b b b a b a b a a b a a a a a a a b a b b a b c a b a a b c b c a a a a a a a a a a a a a a a Orange OctNovDecJanFebMarAprMay FT4 (pg/mL) 0 1 2 3 4 5 6 7 8 OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5 a a a a a a a a a a a b b b b a a a a a a b c a b c a b c a a a a a a a aa a a c b b b Emerelda OctNovDecJanFebMarAprMay FT4 (pg/mL) 0 1 2 3 4 5 6 EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 a a b b a b b a a a a a a a a a a a a a a a a a b a a b a b b a a a a a a a a a a a a a a a Apopka OctNovDecJanFebMarAprMay FT4 (pg/mL) 0 1 2 3 4 5 6 7 8 9 AP-04-10 AP-04-W2 AP-04-W10 a b b a a a a a a a a a a a a a a a a a a b b a

PAGE 92

80 Testosterone Estradiol GHRH Estradiol Neuropeptide Y GLP-I Estradiol Norepinephrine Galanin Somatotroph cells of anterior pituitary Growth Hormone Release Ration Size Protein Intake Starvation Acute Stress Chronic Stress TSH T3 T4 DOPA Dopamine Norepinephrine 5-Hydroxytryptamine Somatostatin-25 Somatostatin-28 SRIF IGF-I NPY NMA GH TRH SRIF GnRH Estradiol SRIF Bombesin Exercise Ovulation Temperature Daylength Seawater Adaptation CCK Figure 2-20.Graphical interpre tation of factors that cont rol the release of growth hormone. Adapted from Mommsen, 1998.

PAGE 93

81 Table 2-1. Total length growth rates among and within sites. Apopka Sept Oct Nov Dec Jan Feb Mar Apr May MEAN AP-04-10 0.92719 1.2397 0.8521 0.8392 1.061 1.0336 1.0406 0.9725 1.0382 1.0005 AP-04-W2 0.65300 0.8118 0.9669 0.9649 0.929 0.9695 0.9993 1.0239 1.0647 0.9314 AP-04-W10 0.87014 1.2942 1.3956 1.3069 1.2078 1.1868 0.9884 1.1727 1.2023 1.1805 MEAN 0.81678 1.1152 1.0715 1.037 1.0659 1.0633 1.0094 1.0564 1.1017 1.0375 Emeralda Sept Oct Nov Dec Jan Feb Mar Apr May MEAN EM-04-01 0.77604 1.1813 1.3056 1.2425 1.1461 1.0249 1.1103 1.0215 1.0639 1.0969 EM-04-02 1.35882 1.4240 1.5496 1.5557 1.4495 1.3829 1.3603 1.1997 1.2105 1.3879 EM-04-03 0.60000 0.8813 0.9617 0.8891 0.9475 0.9306 0.9593 0.9855 1.0178 0.9081 EM-04-04 0.71874 0.8774 1.0094 1.0137 0.9838 1.0014 0.981 0.9953 1.0308 0.9568 EM-04-11 1.30798 1.2455 1.2505 1.1403 1.062 1.0473 1.024 1.0319 1.0678 1.1308 MEAN 0.95232 1.1219 1.2154 1.1683 1.1178 1.0774 1.087 1.0468 1.0781 1.0961 Griffin Sept Oct Nov Dec Jan Feb Mar Apr May MEAN GR-04-51 1.05110 1.1677 1.3124 1.2542 1.1867 1.183 1.1959 1.1849 1.195 1.1923 GR-04-A 1.17706 1.1154 1.2462 1.1698 1.1075 1.0523 1.016 0.9912 1.0139 1.0988 GR-04-B 1.25966 1.4946 1.5755 1.4899 1.3532 1.4082 1.3217 1.2943 1.3036 1.389 GR-04-C 1.01199 1.1343 1.3977 1.362 1.2779 1.2944 1.2933 1.2191 1.2655 1.2507 GR-04-D 0.95602 1.3219 1.4598 1.3821 1.2988 1.3109 1.147 1.3192 1.3961 1.288 MEAN 1.09117 1.2468 1.3983 1.3316 1.2448 1.2498 1.1948 1.2017 1.2348 1.2438 Orange Sept Oct Nov Dec Jan Feb Mar Apr May MEAN OR-04-12 0.78744 1.0712 1.165 1.105 1.0405 0.999 0.9785 0.985 1.026 1.0175 OR-04-13 0.51497 0.8102 1.0353 1.0101 1.0077 1.0107 1.0663 0.9947 1.009 0.9399 OR-04-B 2.94167 1.4562 1.4706 1.4429 1.346 1.3432 1.3341 1.2683 1.2874 1.5434 OR-04-W1 0.77147 1.0547 1.1855 1.1606 1.1091 1.1101 1.062 1.0742 1.1135 1.0713 OR-04-W5 1.20455 1.3867 1.4210 1.3716 1.2823 1.3337 1.3012 1.2719 1.2736 1.3163 MEAN 1.24402 1.1558 1.2555 1.218 1.1571 1.1593 1.1484 1.1188 1.1419 1.1777

PAGE 94

82 Table 2-2. Snout-vent length growth rates among and within sites. Apopka Sept Oct Nov Dec Jan Feb Mar Apr May MEAN AP-04-10 0.41725 0.5400 0.4251 0.4391 0.513 0.5115 0.5123 0.4711 0.4985 0.4809 AP-04-W2 0.29293 0.3679 0.4742 0.4943 0.4633 0.4845 0.4915 0.5021 0.5209 0.4546 AP-04-W10 0.35694 0.6228 0.6875 0.5936 0.594 0.5738 0.5838 0.562 0.579 0.5726 MEAN 0.35571 0.5102 0.529 0.509 0.5234 0.5233 0.5292 0.5117 0.5328 0.5027 Emeralda Sept Oct Nov Dec Jan Feb Mar Apr May MEAN EM-04-01 0.35677 0.5489 0.6454 0.612 0.5553 0.5518 0.5424 0.4982 0.5136 0.536 EM-04-02 0.0239 0.4779 0.6486 0.6905 0.645 0.6256 0.6075 0.5411 0.5434 0.5337 EM-04-03 0.30013 0.4205 0.4855 0.4862 0.4772 0.4622 0.4723 0.4821 0.4935 0.4533 EM-04-04 0.31928 0.4267 0.5154 0.525 0.4989 0.5042 0.4818 0.4978 0.5114 0.4756 EM-04-11 0.70714 0.6343 0.6458 0.6089 0.543 0.5394 0.5204 0.5207 0.5309 0.5834 MEAN 0.34145 0.5017 0.5881 0.5845 0.5439 0.5366 0.5249 0.508 0.5185 0.5164 Griffin Sept Oct Nov Dec Jan Feb Mar Apr May MEAN GR-04-51 0.45901 0.5561 0.6835 0.6408 0.592 0.5932 0.5952 0.5848 0.5841 0.5876 GR-04-A 0.49338 0.5452 0.6354 0.5891 0.5478 0.5264 0.6536 0.4852 0.4899 0.5518 GR-04-B 0.51534 0.684 0.7696 0.7288 0.6537 0.6817 0.6582 0.6192 0.6277 0.6598 GR-04-C 0.41335 0.5216 0.6842 0.6815 0.6404 0.6308 0.6271 0.5895 0.6091 0.5997 GR-04-D 0.36947 0.5993 0.7288 0.6933 0.6338 0.6469 0.7693 0.6421 0.6752 0.6398 MEAN 0.45011 0.5813 0.7003 0.6667 0.6135 0.6158 0.6607 0.5842 0.5972 0.6078 Orange Sept Oct Nov Dec Jan Feb Mar Apr May MEAN OR-04-12 0.31765 0.4799 0.5794 0.5471 0.5018 0.4862 0.4729 0.4669 0.4829 0.4816 OR-04-13 0.31012 0.3886 0.5214 0.5114 0.4996 0.4954 0.5174 0.4926 0.4879 0.4694 OR-04-B 0.52337 0.6927 0.7193 0.7192 0.6548 0.658 0.6483 0.6109 0.6211 0.6497 OR-04-W1 0.36559 0.5127 0.5961 0.5916 0.5435 0.5632 0.5356 0.5445 0.5561 0.5343 OR-04-W5 0.51056 0.6264 0.6669 0.6606 0.6066 0.6115 0.6203 0.5984 0.5876 0.6099 MEAN 0.40546 0.54 0.6166 0.606 0.5613 0.5629 0.5589 0.5427 0.5471 0.549

PAGE 95

83 Table 2-3. Head length growth rates among and within sites. Apopka Sept Oct Nov Dec Jan Feb Mar Apr May MEAN AP-04-10 0.10576 0.1288 0.0943 0.0990 0.1227 0.1183 0.1178 0.11 0.1175 0.1127 AP-04-W2 0.08664 0.0973 0.1133 0.1126 0.1096 0.1122 0.1137 0.116 0.1232 0.1094 AP-04-W10 0.12184 0.1536 0.1593 0.1474 0.1394 0.1374 0.1388 0.134 0.1385 0.1411 MEAN 0.10475 0.1266 0.1223 0.1197 0.1239 0.1227 0.1235 0.12 0.1264 0.1211 Emeralda Sept Oct Nov Dec Jan Feb Mar Apr May MEAN EM-04-01 0.11993 0.1372 0.1445 0.1366 0.1283 0.1237 0.1216 0.1127 0.1196 0.1271 EM-04-02 0.11278 0.1434 0.1605 0.1619 0.1549 0.1477 0.1471 0.1319 0.1348 0.1439 EM-04-03 0.07988 0.11 0.1101 0.1104 0.1112 0.108 0.1109 0.1115 0.1172 0.1077 EM-04-04 0.15792 0.1075 0.118 0.1175 0.1152 0.1158 0.1142 0.1167 0.1202 0.1203 EM-04-11 0.1484 0.1422 0.1386 0.1300 0.1212 0.1175 0.1167 0.117 0.1219 0.1282 MEAN 0.12378 0.1281 0.1344 0.1313 0.1262 0.1225 0.1221 0.118 0.1228 0.1254 Griffin Sept Oct Nov Dec Jan Feb Mar Apr May MEAN GR-04-51 0.13007 0.1454 0.1532 0.1466 0.1412 0.1375 0.1373 0.1374 0.1397 0.1409 GR-04-A 0.05335 0.1392 0.1373 0.1371 0.1288 0.1223 0.1181 0.1149 0.1178 0.1188 GR-04-B 0.12613 0.159 0.1683 0.1600 0.1502 0.1551 0.1523 0.1439 0.1459 0.1512 GR-04-C 0.12601 0.138 0.1556 0.1511 0.1448 0.1325 0.1445 0.1368 0.1458 0.1417 GR-04-D 0.08718 0.1407 0.1511 0.1448 0.1402 0.1392 0.1375 0.1404 0.1501 0.1368 MEAN 0.10455 0.1445 0.1531 0.1479 0.141 0.1373 0.138 0.1347 0.1399 0.1379 Orange Sept Oct Nov Dec Jan Feb Mar Apr May MEAN OR-04-12 0.10978 0.1278 0.1300 0.1223 0.1168 0.1126 0.1095 0.1085 0.1139 0.1168 OR-04-13 0.06736 0.0980 0.1125 0.1144 0.1125 0.1105 0.1169 0.1099 0.1132 0.1061 OR-04-B 0.12861 0.1578 0.1572 0.1567 0.1499 0.1492 0.1472 0.14 0.1441 0.1479 OR-04-W1 0.03139 0.0911 0.1105 0.1136 0.1119 0.116 0.1141 0.1154 0.1215 0.1028 OR-04-W5 0.14353 0.1569 0.1543 0.1482 0.1409 0.1396 0.1398 0.1388 0.142 0.1449 MEAN 0.09613 0.1263 0.1329 0.1311 0.1264 0.1256 0.1255 0.1225 0.1269 0.1237

PAGE 96

84 Table 2-4. Body weight growth rates among and within sites. Apopka Sept Oct Nov Dec Jan Feb Mar Apr May MEAN AP-04-10 -0.142 0.3339 0.4028 0.5629 0.664 0.8235 0.9383 0.9884 1.2239 0.6439 AP-04-W2 -0.186 0.2751 0.5495 0.7342 0.6933 0.9298 1.1494 1.2807 1.483 0.7677 AP-04-W10 -0.192 0.4847 0.8331 1.0648 0.9643 1.157 1.4169 1.5286 1.7323 0.9989 MEAN -0.173 0.3646 0.5951 0.7873 0.7739 0.9701 1.1682 1.2659 1.4797 0.8035 Emeralda Sept Oct Nov Dec Jan Feb Mar Apr May MEAN EM-04-01 -0.096 0.4238 0.6937 0.8715 0.814 0.9782 1.1183 1.1024 1.3155 0.8024 EM-04-02 -0.053 0.565 0.9901 1.3877 1.2348 1.4202 1.7494 1.4623 1.6692 1.1584 EM-04-03 -0.060 0.2368 0.3948 0.5283 0.5246 0.6456 0.8348 0.9462 1.1025 0.5725 EM-04-04 -0.244 0.2415 0.5107 0.7018 0.6939 0.8735 1.0231 1.1602 1.3324 0.6992 EM-04-11 -0.125 0.3893 0.6335 0.7705 0.7107 0.8668 0.9828 1.1521 1.2867 0.7408 MEAN -0.116 0.3713 0.6445 0.8519 0.7956 0.9569 1.1417 1.1646 1.3412 0.7947 Griffin Sept Oct Nov Dec Jan Feb Mar Apr May MEAN GR-04-51 -0.085 0.4041 0.7652 1.0108 0.9695 1.2064 1.3834 1.5226 1.5575 0.9705 GR-04-A -0.196 0.3000 0.6267 0.8584 0.7982 0.9064 1.0218 1.0647 1.2059 0.7318 GR-04-B -0.048 0.5038 0.8912 1.1153 1.0485 1.3593 1.5635 1.6191 1.8298 1.0981 GR-04-C -0.097 0.4397 0.8715 1.1457 1.0766 1.3123 1.5465 1.5398 1.8769 1.0791 GR-04-D -0.034 0.5374 0.8623 1.1058 1.0394 1.2682 1.5675 1.7796 2.18 1.1451 MEAN -0.092 0.4370 0.8034 1.0472 0.9864 1.2106 1.4165 1.5051 1.73 1.0049 Orange Sept Oct Nov Dec Jan Feb Mar Apr May MEAN OR-04-12 -0.118 0.3779 0.6099 0.7557 0.7193 0.8099 0.8451 0.9966 1.1556 0.6836 OR-04-13 -0.102 0.2698 0.5399 0.6947 0.6913 0.8461 1.0955 1.0616 1.1855 0.698 OR-04-B -0.113 0.5184 0.8935 1.1375 1.0759 1.3295 1.6161 1.6289 1.8493 1.104 OR-04-W1 -0.127 0.2920 0.5567 0.7012 0.6772 0.8885 1.0224 1.1775 1.348 0.7263 OR-04-W5 0.008 0.4602 0.7101 0.9814 0.9218 1.1863 1.4461 1.5617 1.7317 1.0008 MEAN -0.09 0.3837 0.662 0.8541 0.8171 1.0121 1.2051 1.2853 1.454 0.8426

PAGE 97

85 Table 2-5. Hatchling alligator thyroid (TSI) and liver (LSI) somatic indices among sites over time. No differences were obser ved in TSI among sites. Temporal differences were observed in LSI among sites. Significant differences determined utilizing Wilkoxon analysis with the Kruskal –Wallis Test (p < .05). TSI among Sites LSI among Sites Date Chi Square Pr > Chi Square Date Chi Square Pr > Chi Square Sept 4.2855 0.2322 Sept 6.8526 0.0767 Nov 5.936 0.1148 N ov 17.0271 0.0007 Jan 1.1091 0.7749 Jan 6.1687 0.1037 Mar 4.8678 0.1817 Mar 9.786 0.0205 May 1.0038 0.8003 May 6.3759 0.0947

PAGE 98

86 Table 2-6. Hatchling alligator thyroid somatic indices (TSI) within sites over time. No significant differences were observed. Significant differences determined utilizing Wilkoxon analysis with th e Kruskal –Wallis Test (p < .05). Apopka Chi-Square Pr> Chi Square Sept 3.4667 0.1767 Nov 3.2000 0.2019 Jan 0.3556 0.8371 Mar 5.0667 0.0794 May 5.0667 0.0794 Emeralda Chi-Square Pr> Chi Square Sept 5.1434 0.2729 Nov 10.8945 0.0278 Jan 9.5667 0.0484 Mar 5.5667 0.2339 May 4.9333 0.2942 Griffin Chi-Square Pr> Chi Square Sept 9.1747 0.0569 Nov 6.8706 0.1429 Jan 4.7667 0.3121 Mar 7.2667 0.1224 May 9.7333 0.0452 Orange Chi-Square Pr> Chi Square Sept 4.2000 0.3796 Nov 8.9667 0.0619 Jan 10.7667 0.0293 Mar 2.5796 0.6304 May 7.0000 0.1359

PAGE 99

87 Table 2-7. Hatchling alligator liver somatic indices (LSI) within sites over time. No significant differences were observed. Significant differences determined utilizing Wilkoxon analysis with th e Kruskal –Wallis Test (p < .05). Apopka Chi-Square Pr> Chi Square Sept 3.2889 0.1931 Nov 3.2000 0.2019 Jan 4.6222 0.0992 Mar 0.8000 0.6703 May 5.9556 0.0509 Emeralda Chi-Square Pr> Chi Square Sept 10.4333 0.0337 Nov 4.4667 0.3465 Jan 4.6333 0.3270 Mar 10.833 0.0285 May 3.9000 0.4197 Griffin Chi-Square Pr> Chi Square Sept 12.0333 0.0171 Nov 8.1667 0.0857 Jan 11.7000 0.0197 Mar 7.9667 0.0928 May 3.7000 0.4481 Orange Chi-Square Pr> Chi Square Sept 7.1711 0.1271 Nov 9.2333 0.0555 Jan 5.3000 0.2579 Mar 2.5000 0.6446 May 2.7667 0.5976

PAGE 100

88 Table 2-8. Multiple linear regression analysis of hatchling alligator growth rates, thyroid hormone secretory rates and organochlor ine contaminant concentrations. No significant relationships were demonstrated (p < .05). Total Length Rate Snout-Vent Length Rate Head Length Rate Body Weight Rate Total Chlordane 0.1084 0.1108 0.3362 0.3072 Total DDTx 0.1129 0.0959 0.3938 0.1462 Dieldrin 0.1281 0.1246 0.3376 0.4492 Toxaphene 0.4905 0.6954 0.8230 0.5753 TT4 Rate 0.3704 0.7254 0.5308 0.8545 FT4 Rate 0.2137 0.1193 0.4314 0.1983

PAGE 101

89 LIST OF REFERENCES Bjornsson, B.T., Johansson, V., Benedet, S., Ei narsdottir, I.E., Hildahl, H., Agustsson, T., Jnsson, E. (2002) Growth hormone endocrinology of salmonids: regulatory mechanisms and mode of action. Fish Physiology and Biochemistry 27: 227-242. Blanar, C.A., Curtis, M.A., Chan, H.M. (2005) Growth, nutritional composition, and hematology of Artic Charr ( Salvelinus alpinus ) exposed to toxaphene and tapeworm ( Diphyllobothrium dendriticum ) larvae. Archives of Environmental Contamination and Toxicology 48: 397-404. Bona-Gallo, A., Licht, P., MacKenzie, D.S., Lo fts, B. (1980) Annual cycles in levels of pituitary and plasma gonadotropin, gonadal steroids, and thyroid activity in the Chinese cobra ( Naja naja ). General and Compara tive Endocrinology 42: 477-493. Brouwer, A., Morse, D.C., Lans, M.C., Sc huur, G., Murk, A.J., Klasson-Wehler, E., Bergman, A., Visser, T.J. (1998) Inter actions of persistent environmental organohalogens with the thyroid hormone system: mechanisms and possible consequences for animal and human health. Toxicology and Industrial Health 14(1/2): 59-84. Brown, S.B., Adams, B.A., Cyr, D.G., Eale s, J.G. (2004) Contaminant effects on the teleost fish thyroid. Environmental Toxicology and Chemistry 23(7): 1680-1701. Brown, S.B., Fisk, A.T., Brown, M., Vinella, M., Muir, D.C.G., Evans, R.E., Lockhart, W.L., Metner, D.A., Cooley, H.M. (2002) Dietary accumulation and biochemical responses in juveni le rainbow trout ( Oncorhynchus mykiss ) to 3,3’4,4’,5pentachlorobipheyl (PCB 126). A quatic Toxicology 59(3-4): 139-152. Brucker-Davis, F. (1998) Effects of envi ronmental synthetic chemicals on thyroid function. Thyroid 8(9): 827-856. Bustnes, J.O., Miland, O., Fjeld, M., Eriksta d, K.E., Skaare, J.U. (2005) Relationships between ecological variables and four orga nochlorine pollutants in an artic glacous gull ( Larus hyperboreous ) population. Environmen tal Pollution 136: 175-185. Capen, C.C. (2001) Cassarett and Doull’s Toxicology The Basic Science of Poisons 6th Edition. Chapter 21: Toxic Responses of the Endocrine System, Thyroid Gland (Follicular Cells). Ed. Curtis D. Klaassen. McGraw-Hill Medical Publishing Division. New York, NY..

PAGE 102

90 Chang, L., Munro, L.A., Richardson, S.J., Schr eiber, G. (1999) Evolution of thyroid hormone binding by transthyretins in bird s and mammals. Eur. J. Biochem. 259: 534-542. Cheek, A.O., Kow, K., Chen, J., MacLachlan, J.A. (1999) Potential mechanisms of thyroid disruption in humans: interacti on of organochlorine compounds with the thyroid receptor, transthyre tin, and thyroid-binding globu lin. Environmental Health Perspectives 107(4): 273-278. Conley, A.J., Elf, P., Corbin, C.J., Dubows ky, S., Fivizzani, A., Lang, J.W. (1997) Yolk steroids decline during sexual differen tiation in the alligator. General and Comparative Endocrinology 107: 191-200. Cooley, H.M., Fisk, A.T., Wiens, S.C., Tom y, G.T., Evans, R.E., Muir, D.C.G. (2001) Examination of behavior and thyroid histology of juveni le rainbow trout ( Oncorhynchus mykiss ) exposed to high dietary concentrations of C10 -, C11 C12 and C14 polychlorinated n -alkanes. Aquatic Toxicology 54: 81-99. Crain, D.A., Guillette, Jr., L.J., Pickford, D. B., Percival, H.F., Woodward, A.R. (1998) Sex-steroid and thyroid hor mone concentrations in juvenile alligators ( Alligator mississippiensis ) from contaminated and reference lakes in Florida, USA. Environmental Toxicology and Chemistry 17(3): 446-452. Denver, R.J. and Licht, P. (1991) Depende nce of body growth on thyroid activity in turtles. The Journal of E xperimental Zoology 258: 48-59. Eales, J.G. and Shostak, S. (1985) Free T4 and T3 in relation to total hormone, free hormone indices, and protein in plasma of Rainbow Trout and Artic Charr. General and Comparative Endocrinology 58: 291-302. Eales, J.G. (1984) The peripheral metabolism of thyroid hormones and regulation of thyroid status in poikiloth erms. Can. J. Zool. 63: 1217-1231. Elsey, R.M., Joanen, T., McNease, L., Kinl er, N. (1992) Growth rates and body condition factors of Alligator mississippiensis in coastal Louisiana wetlands: a comparison of wild and farm-raised juveniles. Comp arative Biochemistry and Physiology 103A(4): 667-672. Elsey, R.M., Joanen, T., McNease, L., Lan ce, V. (1990) Growth rate and plasma corticosterone levels in j uvenile alligators maintain ed at different stocking densities. The Journal of E xperimental Zoology 255: 30-36. Ferguson, M.W.J. (1985) The reproductive bi ology and embryology of crocodilians. In Biology of the Reptilia : Vol. 14 Development A : 329-491. Gans, C. (Ed.) New York, NY: John Wiley and Sons.

PAGE 103

91 Fowles, J.R., Fairbrother, A., Trust, K.A., Ke rkvliet, N.I. (1997) Effects of Arochlor 1254 on the thyroid gland, immune function, a nd hepatic cytochrome p450 activity in mallards. Environmental Research 75: 119-129. Fregly, M.J., Waters, I.W., Straw, J.A. ( 1967) Effect of isomers of DDD on thyroid and adrenal function in rats. Canadian Jour nal of Physiology and Pharmacology 46:5966. Galton, V.A. and Cohen, J.S. (1980) Action of thyroid hormones in premetamorphic tadpoles: an important role for thyroxine?. Endocrinology 107: 1820-1826. Greenblatt, M., Brown, C.L., Lee, M., Dauder, S., Bern, H.A. (1989) Changes in thyroid hormone levels in eggs and larvae and the iodide uptake of eggs of coho and chinook salmon, Oncorhynchus kisutsch and O. tschawytscha Fish Physiology and Biochemistry 6(5): 261-278. Gross, T.S., Guillette, L.J., Percival, H.F., Masson, G.R., Matter, J.M., Woodward, A.R. (1994) Contaminant-induced reproductive anomalies in Florida. Comparative Pathology Bulletin XXVI(4): November. Guillette, L.J., Jr. (1995) Endocrine disr upting environmental contaminants and developmental abnormalities in embryos. Human and Ecological Risk Assessment 1: 25-36. Guillette Jr., L.J., Woodward, A.R., Crain, D. A., Pickford, D.B., Rooney, A.A., Percival, H.F. (1999) Plasma steroid concentrati ons and male phallus size in juvenile alligators from seven Flor ida lakes. General and Co mparative Endocrinology 116: 356-372. Guillette, L.J., Jr., Crain, D.A., Rooney, A.A ., Pickford, D.B. (1995) Organization versus activation: the role of endocrine disrup ting contaminants (EDCs) during embryonic development in wildlife. Environmen tal Health Persp ectives 103: 157-164. Gunderson, M.P., Bermudez, D.S., Bryan, T.A., Crain, D.A., Degala, S., Edwards, T.M., Kools, S.A.E., Milnes, M.R., Guillette, Jr., L.J. (2002) Temporal and spatial variation in plasma thyroxine (T4) concentrations in juvenile alligators collected from lake Okeechobee and the northern Ever glades, Florida, USA. Environmental Toxicology and Chemistry 21(5): 914-921. Heinz, G.H., Percival, H.F., Jennings, M.L. (1991) Contaminants in American alligator eggs from lake Apopka, lake Griffin, and lake Okeechobee, Florida. Environmental Monitoring and Assessment 16: 277-285. Hewitt, E.A., Crain, D.A., Gunderson, M.P., Gu illette Jr., L.J. (2002) Thyroid status in juvenile alligators ( Alligator mississippiensis ) from contaminated and reference sites on Lake Okeechobee, Florida, USA. Chemosphere 47:1129-1135.

PAGE 104

92 Hulbert, A.J. (2000) Thyroid hormones and thei r effects: a new pers pective. Biol. Rev. 75:519-631. Hulbert, A.J. and Williams, C.A. (1988) Thyr oid function in a lizard, a tortoise, and a crocodile, compared with mammals. Comp. Biochem. Physiol. 90A(1): 41-48. Jallageas, M. and Assemabcher, I. (1979) Fu rther evidence for reciprocal interactions between the annual sexual and thyroid cy cles in male Peking ducks. General and Comparative Endocrinology 37: 44-51. Jallegeas, M., Tamisier, A., A ssenmacher, I. (1978) A comp arative study of the annual cycles of sexual and thyroid function in male Peking ducks ( Anas platyrhynchos ) and teal ( Anas crecca ). General and Comparativ e Endocrinology 36: 201-210. Jefferies, D.J. and French, M.C. (1972) Ch anges induced in the pigeon thyroid by p,p, DDE and dieldrin. Journal of W ildlife Management 36(1): 24-30. Jefferies, D.J. and Parslow, J.L.F. (1972) Eff ect of one polychlorinated biphenyl on size and activity of the gull thyroid. Bulletin of Environmental Contamination and Toxicology 8(5): 306-310. Joanen, T. and McNease, L. (1979) Culture of the American alligato r. International Zoo Yearbook 19: 61-66. Kar, A. and Chandola-Saklani, A. (1984) Circ ulating thyroid hormone concentrations in relation to seasonal events in the male Indian Garden lizard, Calotes versicolor General and Comparativ e Endocrinology 60: 14-19. Kato, Y., Ikushiro, S., Haraguchi, K., Yam azaki, T., Ito, Y., Suzuki, H. Kimura, R., Yamada, S., Inoue, T., Degawa, M. (2004) A possible mechanism for decrease in serum thyroxine level by polychlorinated biphenyls in Winstar and Gunn rats. Toxicological Sciences 81: 309-315. Kato, Y., Haraguchi, K., Yamazaki, T., Ito, Y ., Shoji, M., Nemoto, K., Koga, N., Kimura, R., Degawa, M. (2003) Effects of poly chlorinated biphenyls, Kaechlor-500, on serum thyroid hormone levels in rats and mice. Toxicological Science 72: 235-241. Kobuke, L., Specker, J.L., Bern, H.A. (1987) Th yroxine content of eggs and larvae of coho salmon, Oncorhynchus kisutch The Journal of Experimental Zoology 242: 89-94. Kohel, K.A., MacKenzie, D.S., Rostal, D. C., Grumbles, J.S., Lance, V.A. (2001) Seasonality in plasma thyroxi ne in the desert tortoise, Gopherus agassizii General and Comparative Endocrinology 121: 214-222. Lamai, S.L., Warner, G.F., Walker, C.H. (1999) Effects of dieldrin on life stages of the African catfish ( Clarias gariepinus ). Ecotoxicology and Environmental Safety 42: 22-29.

PAGE 105

93 Licht, P., Denver, R.J., Herrera, B.E. ( 1991) Comparative survey of blood thyroxine binding in turtles. The Journal of Experimental Zoology 259:43-52. Licht, P., Denver, R.J., Pavgi, S. (1989) Te mperature dependence of in vitro pituitary, testis, and thyroid se cretion in a turtle, Pseudemys scripta General and Comparative Endocrinology 76: 274-285. Licht, P., Breitenbach, G.L., Congdon, J.D. (1985) Seasonal cycles in testicular activity, gonadotropin, and thyroxine in the Painted Turtle, Chrysemys picta under natural conditions. General and Compar ative Endocrinology 59: 130-139. Licht, P., Wood, J.F., Wood, F.E. (1985) Annual and diurnal cycles in plasma testosterone and t hyroxine in the male Green Sea Turtle, Chelonia mydas General and Comparative Endocrinology 57: 335-344. Masson, G.R. (1995) Environmental influences on reproductive potential, clutch viability and embryonic mortality of the American Alligator in Florida. PhD Dissertation, University of Florida. McNabb, F.M.A., Jang, D.A., Larsen, C.T. (2004) Does thyroid function in developing birds adapt to sustained ammonium perc hlorate exposure?. T oxicological Science 82: 106-113. McNabb, M.A. and Fox, G.A. (2003) Avian thyroid development in chemically contaminated environments: is there eviden ce of alterations in thyroid function and development?. Evolution and Development 5(1): 76-82. McNabb, F.M.A. (2000) Sturkie’s Avian Physiology Chapter 17: Thyroids pp: 461-469. Ed. G. Causey Whittow. Academic Press. San Diego, CA. McNabb, F.M. (1995) Thyroid hormones, thei r activation, degradation and effects on metabolism. The Journal of Nutrition 125(6 Suppl.): 1773S-1776S. Moccia, R.D., Fox, G.A., Britton, A. (1986) A quantitative assessment of thyroid histopathology of Herring gulls ( Larus argentatus ) from the great lakes and a hypothesis on the causal role of contaminan ts. Journal of Wildlife Diseases 22(1): 60-70. Moccia, R.D., Leatherland, J.F., Sonstegar d, R.A. (1981) Quantitative comparison of thyroid pathology in Great Lakes Coho ( Oncorhynchus kisutch ) and Chinook ( Oncorhynchus tschawytscha ) salmon. Cancer Research 41: 2200-2210. Mommsen, T.P. (1998) The Physiology of Fishes 2nd Edition. Ed. David H. Evans. CRC Press. Boca Raton, FL.. Mondou, P.M. and Kaltenbach, J.C. (1979) Thyr oxine concentrations in blood serum and pericardial fluid of metamorphosing ta dpoles and of adult frogs. General and Comparative Endocrinology 39: 343-349.

PAGE 106

94 Monteverdi, G.H. and DiGuilio, R.T. (2000) Vitellogenin association and oocytic accumulation of thyroxine and 3,5,3’triiodothyronine in gravid Fundulus heteroclitus General and Comparativ e Endocrinology 120: 198-211. Nishimura, N., Miyabara, Y., Sato, M., Yonemoto, J., Tohyama, C. (2002) Immunohistochemical localization of thyr oid stimulating hormone induced by a low oral dose of 2,3,7,8-tetrachlorobenzo-pdioxin in female Sprague-Dawley rats. Toxicology 171(2-3): 73-82. Norman, A.W. and Litwack, G. (1997) Hormones 2nd Edition. Academic Press. San Diego, CA. Olson, J.M., McNabb, F.M.A., Jablonski, M.S ., Ferris, D.V. (1999) Thyroid development in relation to the development of en dothermy in the red Winged Blackbird ( Agelaius phoeniceus ). General and Comparative Endocrinology 116: 204-212. O’Steen, S. and Janzen, F.J. (1999) Em bryonic temperature affects metabolic compensation and thyroid hormones in ha tchling snapping tu rtles. Physiol. Biochem. Zool. 72: 520-533. Parkinson, A. (2001) Cassarett and Doull’s Toxicology The Basic Science of Poisons 6th Edition. Chapter 6: Biotransformation of Xenobiotics. Ed. Curtis D. Klaassen. McGraw-Hill Medical Publishing Division. New York, NY.. Plowman, M.M. and Lynn, W.G. (1973) The role of thyroid in testicular function in the gecko, Coleonyx variegatus General and Comparative Endocrinology 20: 342-346. Porterfield, S.P. (1994) Vulnerability of th e developing brain to thyroid abnormalities: environmental insults to the thyroid syst em. Environmental Health Perspectives 102 (Suppl. 2): 125-130. Prapunpoj, P., Richardson, S.J., Schreiber, G. (2002) Crocodile transt hyretin: structure, function, and evolution. Am. J. Physiol. Inter. Comp. Physiol. 283: R885-R896. Prati, M., Calvo, R., Morreale de Esc obar, G. (1992) L-thyroxine and 3,5,3’triiodothyronine concentrati ons in the chicken egg and in the embryo before and after the onset of thyroid func tion. Endocrinology 130(5): 2651-2659. Rauschenberger, R.H. (2004) Developmenta l mortality in American alligators ( Alligator mississippiensis ) exposed to organochlorine pesticides. PhD dissertation. University of Florida. Rauschenberger, H.R., Seplveda, M.S., Wieb e, J.J., Szabo, N.J., Gross, T.S. (2004) Predicting maternal body burdens of orga nochlorine pesticides from eggs and evidence of maternal transfer in Alligator mississippiensis Environmental Toxicology and Chemistry 23(12): 2906-2915.

PAGE 107

95 Rauschenberger, R.H., Wiebe, J.J., Buckland, J.E., Smith, J.T., Seplveda, M.S., Gross, T.S. (2004) Achieving environmentally relevant organochlorine pesticide concentrations in eggs through maternal exposure in Alligator mississippiensis Marine Environmental Research 58: 851-856. Rice, K.G., Percival, H.F., Woodward, A.R ., Abercrombie, C.L. (1998) Population dynamics of Lake Apopka’s alligators. pp. 191-205 in Crocodiles. Proceedings of the 14th Working Meeting of the Crocodile Specialist Group, IUCN – The World Conservation Union, Gland, Sw itzerland and Cambridge UK. Richardson, S.J., Monk, J.A., Shepherdley, C.A ., Ebbesson, L.O.E., Sin, F., Power, D.M., Frappell, P.B., Khrle, J., Renfree, M.B. (2005) Developmentally regulated thyroid hormone distributor in marsupials, a rept ile and fishes. Am. J. Physiol. Comp. Physiol. 288(5): R1264-R1272. Rickenbacher, U., McKinney, J.D., Oatley, S. J., Blake, C.C.F. (1986) Structurally specific binding of halogenated biphenyls to thyroxine transport protein. Journal of Medicinal Chemistry 29: 641-648. Rolland, R.M. (2000) A review of chemically-i nduced alterations in thyroid and vitamin a status from field studies of wildlife a nd fish. Journal of Wi ldlife Diseases 36(4): 615-635. Rotstein, D.S., Schoeb, T.R., Davis, L.M., Gl enn, T.C., Arnold, B.S., Gross, T.S. (2002) Detection by microsatellite analysis of early embryonic mortality in an alligator population in Florida. J. Wildl. Dis. 38:160-165. Rousset, B.A., Dunn, J.T. (2004) The Thyroi d and Its Diseases. Chapter 2: Thyroid Hormone Synthesis and S ecretion from www.thyroidm anager.org/Chapter2/2text.htm on 5/18/04. Schelske, C.L., Lowe, E.F., Battoe, L.E., Brenner, M., Coveney, M.F., Kenney, W.F. (2005) Abrupt biological response to hydrologic and land-use changes in lake Apopka Florida, USA. Ambio 34(3): 192-198. Schelske, C. and Brezonik, P. (1992) Can Lake Apopka be restored? from: Restoration of Aquatic Ecosystems: Science, Technology, and Public Policy pp. 393-398. National Academy Press. Washington, D.C.. Schew, W.A., McNabb, F.M.A., Scanes, C.G. (1996) Comparison of the ontogenesis of thyroid hormones, growth hormone, a nd insulin-like growth factor-I in ad libitum and food-restricted (altrici al) European starlings and (precocial) Japanese quail. General and Comparative Endocrinology 101: 304-316. Scollon, E.J., Carr, J.A., Cobb, G.P. (2004) Th e effect of flight, fasting and p,p’-DDT on thyroid hormones and corticosterone in Gambel’s white-crowned sparrow Zonotrichia leucophrys gambelli. Comparative Biochemistry and Physiology Part C 137: 179-189.

PAGE 108

96 Sechman, A. and Bobeck, S. (1988) Presence of iodothyronines in th e yolk of a hen’s egg. General and Comparative Endocrinology 69: 99-105. Sellers, J.C., Wit, L.C., Ganjam, V.K., Ethr idge, K.A., Ragland, I.M. (1982) Seasonal plasma T4 titers in the hibernating lizard Cnemidophorus sexlineatus General and Comparative Endocrinology 46: 24-28. Shepherdley, C.A., Richardson, S.J., Evan s, B.K., Kuhn, E.R., Darras, V.M. (2002) Characterization of outer ri ng iodothyronine deiodinases in tissues of saltwater crocodiles. General and Compar ative Endocrinology 125:387-398. Sonstegard, R. and Leatherland, J.F. (1976) The epizootiology and pathogenesis of thyroid hyperplasia in Coho salmon ( Oncorhynchus kisutch ) in Lake Ontario. Cancer Research 36: 4467-4475. Sullivan, C.G., Iwamoto, R.N., Dickhoff, W.W. (1987) Thyroid hormones in blood plasma of developing salmon embryos. General and Comparative Endocrinology 65: 337-345. Suzuki, S. and Suzuki, M. (1981) Changes in thyroidal and plasma iodine compounds during and after metamorphosis of the bullfrog, Rana catesbeiana General and Comparative Endocrinology 45: 74-81. Tagawa, M. and Hirano, T. (1987) Presence of thyroxine in eggs and changes in its content during early development of chum salmon, Oncorhynchus keta General and Comparative Endocrinology 68: 129-135. Van den Berg, K.J., van Raaij, J.A.G.M., Brag t, P.C., Notten, W.R.F. (1991) Interactions of halogenated industrial chemicals with transthyretin and effects on thyroid hormone levels in vivo. Arch ives of Toxicology 65: 15-19. Verreault, J., Skaare, J.U., Jenssen, B.M ., Gabrielsen, G.W. (2004) Effects of organochlorine contaminants on thyroid horm one levels in artic breeding Glaucous gulls, Larus hyperboreus Environmental Health Pe rspectives 112(5): 532-537. Very, N.M. and Sheridan, M.A. (2002) The ro le of somatostatins in the regulation of growth in fish. Fish Physiol ogy and Biochemistry 27: 217-226. Wade, M.G., Parent, S., Finnson, K.W., Fo ster, W., Younglai, E., McMahon, A., Cyr, D.G., Hughes, C. (2002) Thyroid toxic ity due to subchronic mixture of 16 organochlorines, lead, and cadmium. Toxicological Sciences 67: 207-218. Waritz, R.S., Steinberg, M., Ki noshita, F.K., Kelly, C.M., Richter, W.R. (1996) Thyroid function and thyroid tumors in toxaphenetreated rats. Regulatory Toxicology and Pharmacology 24: 184-192. Webb, G.J.W. and Manolis, S.C. (1989) Crocodiles of Australia Reed Books Pty. NSW, Australia.

PAGE 109

97 Wiebe, J.J., Seplveda, M.S., Buckland, J.S., Harvey, A., Rauschenberger, H.R., Gross, T.S. (2002) Alligator embryo and hatchling growth from contaminated and clean lakes in central Florida. 16th Working Meeting of the Crocodile Specialist Group, Gainesville, FL, Oct 7-10. Wiebe, J.J., Rotstein, D. Percival, H.F., Woodward, A., Schoeb, T., Gross, T.S. (2001) Evidence of developmental toxicity in the American alligator from central Florida lakes. 20th Annual Meeting of the Society of Environmental Toxicology and Chemistry Abstract Book. November 15-19. San Francisco, CA. pp. 231. Willingham, E. (2001) Embryonic exposure to lowdose pesticides: effects of growth rate in the hatchling red-eared slider turtle Journal of Toxicology and Environmental Health, Part A 64: 257-272. Wilson, F.E. and Reinert, B.D. (1999) Long da ys and thyroxine program American tree sparrows for seasonality: evidence for temporal flexibility for the breeding of euthyroid females. General and Co mparative Endocrinology 113: 136-145. Woodward, A.R., Percival, H.F., Jennings, M.L. Moore, C.T (1993) Low clutch viability of American alligators on Lake Apopka. Florida Scientist 56(1): 52-63. Wu, T.H., Rainwater, T.R., Platt, S. G., McMurry, S.T., Anderson, T.A. (2000) Organochlorine contaminants in Morelet’s crocodile ( Crocodylus moreletii ) from Belize. Chemosphere 40: 671-678. Yntema, C.L. (1968) A series of stages in embryonic development of Chelydra serpetina Journal of Morphology 125: 219-252. Zhou, T., John-Alder, H., Weis, P., Weis, J. S. (1999) Thyroidal st atus of mummichogs ( Fundulus heteroclitus ) from polluted versus a refe rence habitat. Environmental Toxicology and Chemistry 18(12): 2817-2823.

PAGE 110

98 BIOGRAPHICAL SKETCH Jonathan James Wiebe was born on Decembe r 15, 1969, in Pensacola, Florida, and is the son of Ralph and Linda Wiebe. Jon graduated from Gaines ville High School in 1986 and received a BS in wildlife management from the University of Florida in 2000. Jon has spent an extensive amount of his prof essional career in th e care of large and diverse animal collections among various zoological and private collections. The majority of Jon’s professional career has been spent in the laboratory of Dr. Tim Gross. This laboratory specializes in examining the effects of environmental stressors on reproductive and growth parameters in a variet y of different species. Jon is particularly proud of the collaborative work that he ha s achieved with Dr. Tim Gross, Dr. Heath Rauschenberger and Janet Scarborough in the area of alligator ecotoxicology.


Permanent Link: http://ufdc.ufl.edu/UFE0013331/00001

Material Information

Title: Effects of Organochlorine Contaminants on Hatchling American Alligator (Alligator mississippiensis) Growth
Physical Description: Mixed Material
Copyright Date: 2008

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: UFE0013331:00001

Permanent Link: http://ufdc.ufl.edu/UFE0013331/00001

Material Information

Title: Effects of Organochlorine Contaminants on Hatchling American Alligator (Alligator mississippiensis) Growth
Physical Description: Mixed Material
Copyright Date: 2008

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: UFE0013331:00001


This item has the following downloads:


Full Text












EFFECTS OF ORGANOCHLORINE CONTAMINANTS ON HATCHLING
AMERICAN ALLIGATOR (Alligator mississippiensis) GROWTH












By

JONATHAN J. WIEBE


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


2005

































Copyright 2005

by

Jonathan J Wiebe

































This document is dedicated to Ralph Peter "Joey" Wiebe. Though I have not been able to
see your face, your words, thoughts, and style live on forever.















ACKNOWLEDGMENTS

I would like to thank my committee members, Dr. Tim Gross, Dr. Dave Barber,

and Dr. Franklin Percival, for their patience, understanding, and most importantly their

interest in my project. Tim, I will never be able to truly express my thanks for all the

opportunities that he has given me. I thank him for his counsel, beer making skills, and

ability to know "almost" everything before it happens but, most of all I thank you for

being my friend. Mom, I can't say enough about all of the love, support and

understanding that she has provided. I thank her for being a great friend except for the

following: Jon the Mexican baby, Stretch Marks the Spot references, and Bulgur Wheat

care packages. Cheryl, who is my all-time, favorite chick on this rock. I thank her for

having a great attitude, closet neuroses, and removing that fishing hook. Janet, I cannot

thank her enough for all of her help, guidance, support, understanding and great food.

Thanks for making me laugh at myself when I get... well the way that I get. Ruth, thanks

for her supportive words of encouragement and wonderful sense of humor. Thanks to the

many families that I call my own Smiths, Duncans, Greenans, Scarboroughs, Loverns,

and Mitchells. All of you folks have showed tremendous support and kept me alive with

your amazing hospitality and friendship. Heath, I thank him for his time, assistance as

well as classic Arkansas stories. Phil Wilkinson, Franklin Percival and Woody

Woodward, I thank them for instilling in me an appreciation of alligators, southern jokes,

and appreciation of fine BBQ cuisine. Dwayne Carboneau, I thank him for social

commentary on not only alligator season but, life in general. Drs. Dan Sharp and Alan









Ealy, I thank them for providing time and assistance with my project. Finally, I thank all

of my former and current lab mates: Travis "Smitty" Smith, Carla "CW" Wieser, Jim

"Roll Tide" Williams, Sherry "Lionheart" Bostick, Howard "Howie" Jelks, Nikki

"Nicooola" Kernaghan, Shane "Prarie Boy" Ruessler, Alfred "Fredo" Harvey, Jessica

"Gambusia Girl" Noggle, Kevin "The Stick" Johnson, Jessie "Piggy Girl" Grosso, Adro

"Tweety Bird" Fazio, and James "The Tape Man" Basto. Your friendship, patience, and

understanding throughout this MS experience are greatly appreciated.
















TABLE OF CONTENTS



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

L IST O F T A B L E S .................... ..................................... .... .................... viii

LIST OF FIGURES ......... ........................................... ............ ix

ABSTRACT .............. .................. .......... .............. xi

CHAPTER

1 LITER A TU RE REV IEW ......... ................... ............................. .....................1

Overview ........................................... ..................................................1
Organochlorine Contaminant Exposure and Endocrine Disruption in Alligators........2
Alligator Growth and Mortality in Relation to Organochlorine Contaminants............4
T hy roid Stru ctu re.......................... .............. .............. .. ........ ...................... .. 7
Thyroid Hormone Synthesis and Systemic Availability ......................................7
Thyroid H orm one B finding Proteins ........................................... .....................9
Deiodination of Thyroid Horm ones .......................... ..... ................... 10
Thyroid Hormone Availability and Synthesis among Oviparous Species .................12
Species-Differences in Thyroid Hormone Utilization and Regulation....................... 13
F is h ........................................................................................................ 1 3
A m phibians ............................................. 13
A vian ............................................ .........................14
Physiological and Environmental Influences on Thyroid Regulation........................15
O overview ............... .... .......... .......................... .... .... ............ 15
Reproductive and Thyroidal Seasonal Cycles...............................................16
Nutritional Availability and Hibernation................ .......... ... ...... ......... 18
Physiological and Environment Parameters Influence Growth...........................19
Effects of Organochlorine Contaminant Exposure on Thyroid Regulation ..............20
O v erv iew .................................................................... ... ..... ............... 2 0
Effects of Organochlorine Contaminant Exposure on Alligator Thyroid
R regulation ............... .............. ..... .............. ....... .... ........... 21
Thyroid Histology Alterations in Relation to Organochlorine Contaminant
E x p o su re ....................................... ............. ....... .......... ............... 2 3
Influence of Organochlorine Contaminant Exposure on Integrated Levels of
Thyroid H orm one Regulation....................................................................... 25
Thyroid H orm one Synthesis.......................................... ........... ............... 25









Thyroid Horm one Binding Proteins ....................................... ............... 26
Deiodination of Thyroid Horm ones ....................................... ............... 27
Thyroid Hormone Excretion.................... .... ........ ............... 28
Growth in Relation to p,p'-DDE, dieldrin, chlordane and toxaphene exposure.........30
O v erv iew ....................................................... 3 0
Experim ental D ata ............... .... .............. .... .. .... .... .......... ... ......... ... 31
Organochlorine Contaminant Exposure and Hatchling Alligator Growth .................34

2 M AN U SCRIPT .......... ... ............ .... ... ..... ........ ... .. ........ ............ 37

Intro du action .....................................................................................................3 7
M materials and M methods ......................................................................... ............... 42
Egg Collection, Evaluation and Incubation.............................. ............... 42
Clutch Selection .......... ... .................................. ...... ......... 43
Anim al M maintenance ................... ........ ... ......... ........... ............... 44
Hatchling Morphometrics and Tissue Sampling ...........................................44
Plasma Thyroid Hormone Validation Procedures (Total and Free Thyroxine) ..45
Free T4 (FT4) A ssay Procedures..................................... ......... ............... 46
Total T4 (TT4) A ssay Procedures ................................................ .....................46
Analysis of Chlorinated Analytes from Alligator Egg Yolks ...........................47
S statistic s ................................................................... ....................4 9
R e su lts ...................... ... .. ... ............... ... ....................... ................ 4 9
Clutch and Organochlorine Contaminant Parameters ......................................49
H atchling G row th R ates .............................. ................... ............... ... 50
Thyroid Hormones, Growth and Organochlorine Contaminants ......................51
Discussion .............. ............ .... ...... ........ .......................... 52

LIST OF REFERENCES ......... ........ .................................. ...... ... ............... 89

BIOGRAPH ICAL SKETCH ...................................................... 98
















LIST OF TABLES


Table page

2-1. Total length growth rates among and within sites...................................................81

2-2. Snout-vent length growth rates among and within sites..........................................82

2-3. Head length growth rates among and within sites...................................................83

2-4. Body weight growth rates among and within sites................. ................ ...........84

2-5. Hatchling alligator thyroid (TSI) and liver (LSI) somatic indices among sites. ........85

2-6. Hatchling alligator thyroid somatic indices (TSI) within sites over time ................. 86

2-7. Hatchling alligator liver somatic indices (LSI) within sites over time .......................87

2-8. Multiple linear regression analysis of hatchling alligator growth rates,.....................88















LIST OF FIGURES


Figure page

2-1. Graphical interpretation of thyroid hormone biosynthesis.................................61

2-2. Clutch fecundity and clutch viability (site means)................................................62

2-3. Clutch fecundity and clutch viability (current study)....................................63

2-4. Yolk OC concentrations. site means (a) and current study (b).. ............................64

2-5. Hatchling alligator growth parameters among sites over time ..............................65

2-6. Hatchling alligator total length (mm) within sites over time.. ..............................66

2-7. Hatchling alligator snout-vent length (mm) within sites over time.....................67

2-8. Hatchling alligator head length (mm) within sites over time .............................68

2-9. Hatchling alligator body weight (g) within sites over time............................... 69

2-10.Hatchling alligator growth parameters (necropsy animals) among sites over
tim e ......... .... .............. ....................................... .. .............................. 7 0

2-11 .Hatchling alligator total length (mm)(necropsy animals) within sites over time.....71

2-12.Hatchling alligator snout-vent length (mm)(necropsy animals) within sites over
tim e ............. ......... .................................. .......................... 72

2-13.Hatchling alligator head length (mm)(necropsy animals) within sites over time.. ..73

2-14.Hatchling alligator body weight (g) (necropsy animals) within sites over time......74

2-15.Hatchling alligator thyroid weight (g)(necropsy animals) within sites over time....75

2-16.Hatchling alligator liver weight (g) (necropsy animals) within sites over time.......76

2-17.Hatchling alligator total thyroxine(ng/ml)and free thyroxine (pg/ml) plasma
concentrations am ong sites over tim e.. ........................................ ...............77

2-18.Hatchling alligator total thyroxine (ng/ml) plasma concentrations within sites
over tim e.............................................................................................. 78









2-19.Hatchling alligator free thyroxine (pg/ml) plasma concentrations within sites
ov er tim e ...................................... ................................. ......... ...... 7 9

2-20. Graphical interpretation of factors that control the release of growth hormone.. ...80















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

EFFECTS OF ORGANOCHLORINE CONTAMINANTS ON HATCHLING
AMERICAN ALLIGATOR (Alligator mississippiensis) GROWTH


By

Jonathan J Wiebe

December 2005

Chair: Timothy S. Gross
Major Department: Veterinary Medicine

Alterations in alligator reproductive and growth parameters have been reported in

association with organochlorine (OC) contaminated sites in central Florida. These data

indicate reductions in egg and embryo quality as well as reductions in hatchling growth

and survivability. Thyroid, a growth-regulating tissue, has been suggested as a key bio-

indicator of growth among several species. In addition, several researchers have reported

alterations in thyroid regulation in relation to OC contaminant exposure. Previous field

studies have reported alterations in alligator plasma thyroid hormone concentrations as

well as several thyroid histological parameters. However, these data were unable to relate

plasma thyroid hormone (TH) concentrations to alligator growth. Under captive

conditions, preliminary data demonstrated that hatchlings from high OC environments

had hyperthyroid secretary patterns and accelerated growth. The current study examined

the same relationship; however an additional site with high OC contaminant

concentrations was added in order to evaluate the effects of OC contaminant exposure









versus site as it relates to the observed alterations in hatchling growth and thyroid

regulation. In addition, a subset of hatchlings were sacrificed bi-monthly to compare

thyroid and liver weight (indicators of growth) with both hatchling external

morphometrics and plasma TH concentrations over time. Though TH were shown to be

bio-indicators of hatchling growth, no relationship was observed between OC

contaminant exposure and hatchling alligator growth or plasma TH concentrations. These

data suggest that hatchling alligator growth may be influenced by several key factors

including an integrated endocrine network (GH, IGF-I, TH, corticoids), habitat

degradation, as well as OC contaminant exposure.




















CHAPTER 1
LITERATURE REVIEW

Overview

During the 1980's, significant reductions in American Alligator (Alligator

mississippiensis) egg viability were observed on Lake Apopka (a site positioned at the

headwaters of the Ocklawaha river basin with high organochlorine (OC) pesticide

concentrations) in comparison with lake Woodruff, a national wildlife refuge with

reduced concentrations of OC (Woodward, 1993; Rice et al., 1998). In addition, a severe

(- 90%) reduction in the juvenile alligator population was observed on Lake Apopka

(1981-1986) that was likely attributed to reproductive failure (Woodward, 1993). These

observed reductions in juvenile survivability and adult reproductive success have been

attributed in part to the influence of agriculture and anthropogenic alterations

specifically: extensive utilization of organochlorine pesticides by muck farming

operations (i.e., (z 6,000 ha) of the lake's northern wetland was converted for vegetable

production), citrus crops, and effluent discharges from both the citrus processing plant

and sewage treatment facility located at the city of Winter Garden (Woodward et al.,

1993; Schelske and Brezonik, 1992). These environmental alterations were compounded

by the overflow of a wastewater pond located at the Tower Chemical facility which is

adjacent to the Gourd Neck region of Lake Apopka (1980) consisting of high









concentrations of sulfuric acid, DDT, dicofol and several unidentified OC compounds in

which by 1983, the EPA designated this facility's property as a superfund site

(Rauschenberger, 2004). Though several of these OC compounds were identified in yolk

from alligator eggs, no direct association with reduced clutch viability was observed

suggesting other cofactors (i.e., diet, population dynamics, specific OCP mixtures) might

be involved and/or the developmental effects resulted from altered maternal physiology

(caused by OC exposure) as opposed to direct embryotoxicity (Rauschenberger et al.,

2004; Heinz et al. 1991). Therefore, sites that have been historically impacted by varying

degree of OC contamination (lakes Griffin and Apopka as well as the Emeralda Marsh

Conservation Area) continue to demonstrate coincident alterations in reproductive

function and success as measured by sex steroid biomarkers, sexual differentiation, clutch

viability, embryonic mortality, post hatch survivability, and growth (Rauschenberger,

2004; Wiebe et al., 2002; Gross et al. 1994).

Organochlorine Contaminant Exposure and Endocrine Disruption in Alligators

Reductions in alligator reproductive success as well as egg and embryo qualities

have been observed in relation to sites with intermediate to high concentrations of OC

contaminants (Rauschenberger, 2004; Masson, 1995). These chemicals have often been

referred to as "endocrine disruptors" or exogenous agents that interfere with the

production, release, transport, metabolism, binding, action, or elimination of natural

hormones in the body responsible for the maintenance of homeostasis and regulation of

developmental processes (Rolland, 2000; Brucker-Davis, 1998). As some of these OC

contaminants (i.e., p,p'-DDE) have been suggested to have positive and/or negative

estrogenic or androgenic activity, plasma sex steroid concentrations have been one of the

principal biomarkers utilized to examine the relationship between exposure to OC









contaminants and alterations in reproductive productivity. Gross et al. (1994) noted

alterations in plasma sex steroids among juvenile alligators from lakes Apopka (high OC

concentrations) and Woodruff (reference). Specifically, female juvenile alligators had

significantly higher plasma estradiol concentrations versus females from the reference

site (Gross et al., 1994) In contrast, juvenile male alligators from lake Woodruff exhibited

plasma testosterone concentrations that were almost four times higher than males on lake

Apopka (Gross et al., 1994). A similar incidence of altered plasma testosterone

concentrations in juvenile male alligators was reported by Guillette et al. (1999) among

seven Florida lakes. In addition, the author's suggested a relationship between phallus

size (a sex steroid-dependent tissue) as a bio-indicator of anti-androgenic or estrogenic

contaminant exposure (Guillettte et al., 1999).

Masson (1995) reported significant reductions in alligator clutch viability (i.e.

embryonic mortality) on lake Apopka (3.9%) versus conservation sites with low OC

concentrations (71%). The author suggested that lake Apopka's extremely variable, low

clutch viability and hatch percentages confirmed the suggestion that a severe

environmental problem exists at this lake site (Masson, 1995). Rice et al. (1998) observed

that the majority of lake Apopka's embryonic mortality occurred during pre-egg

deposition or in early incubation with the next largest proportion of mortality occurring

very late in incubation. These data continue to support several hypotheses: 1) maternal

OC exposure alters reproductive regulation (as demonstrated by alterations in plasma

estrogen and testosterone concentrations) and, 2) the reported alterations in adult

reproductive fitness as well as maternal-transfer of OC contaminants among yolk

constituents appears to be related to the observed increase in embryonic mortality.









Alligator Growth and Mortality in Relation to Organochlorine Contaminants

It has been suggested that many of the observed embryonic and post-natal

alterations in offspring viability are the result in part of parental exposure to

environmental contaminants (Guillette, 1995). This exposure is primarily associated with

maternal transfer of lipophilic compounds (i.e., OCs) among yolk constituents to

developing offspring (Rauschenberger et al., 2004; Wu et al., 2000). OC exposure has

been suggested to alter hormones that control the course of development and growth and

may have the potential to alter differentiation of major organ systems resulting in

physiological and morphological changes (Rauschenberger et al., 2004; Wu et al., 2000;

Guillette et al., 1995). Wiebe et al. (2001) reported significant alterations in alligator

clutch viability and embryonic and post hatch survivabilities among sites of intermediate

(Griffin) to high (Apopka and Emeralda Marsh) OC concentrations. These data were

strengthened by Rauschenberger's (2004) examination of the relationship between OC

exposure and subsequent reductions in egg and embryo qualities under field and

laboratory conditions. During 2000-2002 field collections, eggs collected from OC

contaminated sites had higher fecundity, lower average clutch mass and reduced clutch

viability in comparison with lake Lochloosa, a site with determined low OC

concentrations (Rauschenberger, 2004). Through the utilization of a captive adult

alligator treatment study, populations (treated and control) were orally dosed with eco-

relevant doses of the four principal OC contaminants identified from the previous field

egg collection: DDT and metabolites (principally p,p'-DDE), dieldrin, chlordanes and

toxaphene or vehicle control (Rauschenberger, 2004). Though reduced clutch viability

was observed in the treated versus control clutches, the majority of the observed mortality

was in the form of unbanded eggs which may represent either early embryonic mortality









or lack of conception (Rotstein et al., 2002). These data, from both field and laboratory,

continue to suggest that overall clutch survival appears to be related to total OC yolk or

maternal burdens (Rauschenberger, 2004).

Alterations in embryonic and hatchling growth as well as reduced post-hatch

survivability in relation to OC exposure has been reported in the American alligator

(Rauschenberger et al., 2004, Wiebe et al., 2002, Wiebe et al., 2001). It seems empirical

that alterations in growth and survivability among animals in these OC contaminated

environments would have ramifications at both site and population levels.

Rauschenberger (2004) examined the incidence of embryonic growth retardation and

survivability in relation to OC exposure utilizing an established embryo staging

methodology (Ferguson, 1985). This evaluation not only examined embryonic

morphological differences among sites over specific developmental time points but, also

evaluated the histopathology of live and dead embryos from "best-case" (clutches with

low mortality rates and low OC egg yolk concentrations) and "worst-case" (clutches with

high mortality rates and high OC egg yolk concentrations) clutches independent of site

(Rauschenberger, 2004). These data demonstrated several key points: 1) the youngest

embryos sampled (calendar day 14 of artificial incubation) showed the strongest

relationship between OC egg concentrations and morphometric parameters, 2)

morphology of live embryos was not consistently different among sites, except during

calendar day 25 (timeframe signifies the middle of organogenesis and may be a more

sensitive time period to OC exposure), 3) morphometry of live embryos was not

significantly related to variation in clutch mortality (i.e.., live embryos from clutches with

high mortality rates develop similarly to those of low mortality rates) 4), cyclodienes









(i.e., chlordane analytes) accounted for an average of 70% of the morphometric variation

that could be attributed to OC variables which is surprising considering DDT and its

metabolites compose an average of 66% of the total OC burden among all sites, 5)

concurrent decreases in maturational age and mass of dead embryos in comparison with

live embryos may have represented normal development up to a point at which the

development stalled and the embryo eventually perished, or embryos could have

developed at a much slower overall rate until the point at which they perished, and 6) no

significant differences in histopathology were observed among "best-case" and "worst-

case" clutches. (Rauschenberger 2004).

The principal mode of alligator embryonic exposure to OC contaminants has been

suggested to occur via maternal transfer among yolk constituents. Several examples have

demonstrated increased incidence of embryonic mortality in relation to exposure to high

concentrations of OC contaminants under both field and laboratory conditions. In

addition, Rauschenberger (2004) detailed significant relationships between OC exposure

and subsequent reductions in embryonic growth and development. Therefore, OC

contaminants are suggested to interfere with the regulation of critical growth and

developmental time periods which may ultimately contribute to the observed increase in

embryonic mortality on OC contaminated sites. These data demonstrate a critical need to

better understand the physiological role in regulating growth and development among

species exposed to OC contaminants.

The thyroid is one of the principal regulatory tissues of growth and development

among multiple taxonomic groups which has been demonstrated to regulate diverse

physiological endpoints including: metabolic rate, tissue differentiation and subsequent









growth and development (Rousset and Dunn, 2004). The two principal physiological

actions of thyroid hormones consist of 1) regulation of cellular differentiation and

development and, 2) regulation of metabolic pathways (Rousset and Dunn, 2004). These

general actions share a common integration in that changes in development and growth

are due to both hormone modulation of metabolism. In addition, cellular differentiation

changes inherently alter changes in gene expression, resulting in modulation of metabolic

pathways (Rousset and Dunn, 2004). A detailed working knowledge of thyroid regulation

is critical in understanding the complex and integrated roles the thyroid plays in growth

and development. Therefore, a literature review is provided which summarizes the

principal factors that regulate thyroid function including tissue structure, thyroid hormone

synthesis, availability, distribution, and deiodination in both embryonic and post-natal

life stages among several poikilothermic as well as homeothermic species.

Thyroid Structure

The thyroid gland is a bilobular tissue that is organized into spherical follicles

whose walls are composed of follicle cells that surround a central lumen filled with

colloid (McNabb, 2000). Colloid is primarily composed of thyroglobulin, a large protein

which is constructed in the rough endoplasmic reticulum, glycosylated in the reticular

lumen, and further post-translationally modified in the golgi apparatus of the follicle cell

(Norman and Litwack, 1997). Thyroglobulin with its tyrosine residues provides the

polypeptide backbone for the synthesis and storage of thyroid hormones as well as an

interim iodine storage area (McNabb, 2000; Norman and Litwack, 1997).

Thyroid Hormone Synthesis and Systemic Availability

The biosynthesis and secretion of thyroid hormones requires four principal

components including: thyroglobulin, thyroperoxidase, hydrogen peroxide and iodide.









Initially, dietary iodide is absorbed from the intestine and transferred from systemic

circulation across the basal lateral membrane of the follicle cells utilizing an ATP-driven

Na+ I active transport (Norman and Litwack, 1997). The sequestered iodide is oxidized

to iodine via thyroperoxidase enzymatic activity in the presence of hydrogen peroxide

(principal electron acceptor) at the cell/colloid interface (McNabb, 2000). Concurrently,

follicle cells synthesize thyroglobulin which contains select tyrosyl residues that will

ultimately be iodinated and coupled to form either monoiodotyrosyls (MIT) or

diiodotyrosyls (DIT) residues and stored as colloid (Norman and Litwack, 1997). In total,

the catalyzing action of thyroperoxidase is required for the oxidation of iodide, iodination

of the thyroglobulin tyrosyl residues and the coupling of the MIT and DIT tyrosyls (i.e.,

thyronines) which based on the coupling combination produces either triiodothyronine

(T3) or thyroxine (T4) (Norman and Litwack, 1997).

Systemic TH availability is regulated utilizing a classic negative feedback

mechanism among the hypothalamic-pituitary-thyroid (HPT) axis (Norman and Litwack,

1997). As thyroid hormones occupy their nuclear receptors in the anterior pituitary, it

suppresses the transcriptional synthesis of preproTSH in the thyrotropes of the anterior

pituitary (Norman and Litwack, 1997). Under conditions of reduced T4, negative

feedback is reduced on thyrotropes of the anterior pituitary (McNabb 2000; Norman and

Litwack, 1997) Thyroid-releasing hormone (TRH) is secreted from the hypothalamus via

the hypophyseal portal vessels interacting with the anterior pituitary which results in the

release of thyroid-stimulating hormone (TSH). TSH interacts with its 7 transmembrane,

G coupled protein receptor on the thyroid follicle cells (Norman and Litwack, 1997,

Eales, 1984). As TSH is the most important controlling factor in iodine availability, the









thyroid follicle will proceed to generate free hormones from the stored hormones

sequestered among thyroglobulin (Norman and Litwack, 1997). This is accomplished as

the apical cell membrane engulfs the colloid by endocytosis and resulting cytoplasmic

colloid droplets fuse with lysosomes to form phagolysosomes (Norman and Litwack,

1997). Thus, the internalized thyroglobulin molecules are subject to a variety of

hydrolytic reactions leading to generation of free thyroid hormones and the complete

degradation of the protein (Rousset and Dunn, 2004; Brown et al., 2004; McNabb, 2000;

Norman and Litwack, 1997).

Thyroid Hormone Binding Proteins

Upon the release of TH from degraded thyroglobulin, a system of plasma proteins

that bind and distribute thyroid hormones is critical to counteract their loss from the

vascular and interstitial compartments by permeation into cell membranes (Prapunpoj et

al., 2002). These binding proteins are integral for systemic circulation due to THs high

lipid solubility (Richardson et al., 2005; Prapunpoj et al., 2002). Albumin (ALB) and

prealbumin or transthyretin (TBPA / TTR) are generally regarded as the two major T4

binding proteins throughout vertebrates; these having low binding affinity and high

capacity (Licht et al., 1991). In addition, many mammals possess thyroxine binding

globulin (TBG), a separate high binding affinity, low capacity binding protein that is

responsible for the principal portion of thyroid hormone binding (Licht et al., 1991).

Thyroid hormone binding protein(s) among vertebrate taxa demonstrate an evolutionary

progression towards increasing thyroid hormone distribution capacity during both

developmental and adult life stages (Richardson et al., 2005). An example of this can be

observed in the binding protein, transthyretin (TTR). TTR is transiently synthesized by

the liver during the time of increased thyroid hormone concentrations (i.e., smoltification,









metamorphosis and development) in fish, amphibians, reptiles whereas it is synthesized

by the liver during development and adult life stages in eutherians and birds (Richardson

et al., 2005). In crocodilians, TTR immunoreactivity has been detected in saltwater

crocodile (Crocodylusporosus) serum on days 60, 68, 75 of egg incubation, and day 1

post-hatch, but not detected in serum at 6 months of age or a 3 year old animal. In

addition, serum albumin was observed at all C. porosus age classes examined

(Richardson et al., 2005). Prapunpoj et al. (2002) demonstrated that C. porosus TTR has

higher binding affinity for T3 versus T4 suggesting that TTR was the principal

transporter of T3 to the crocodilian brain. These data in conjunction with an observed

higher percentage of amino acid sequence identity of C. porosus TTR to chicken TTR

versus lizard TTR and, Chang et al. (1999) observation of avian TTRs having higher

binding affinity for T3 versus eutherian TTRs suggest that the binding properties of C.

porosus TTR are more evolutionarily similar to those of avian TTRs versus eutherian

TTRs (Prapunpoj et al., 2002). Indeed, the separation in evolutionary functionality

between eutherian, avian and poikilotherm thyroid hormone regulation appears to be the

eutherian's ability to generate and regulate thyroid hormones in a tissue-specific manner

(i.e., the evolution of 5' deiodinases) and the utilization of additional binding proteins

(i.e., TBG) which enhances thyroid hormone regulation and distribution (Prapunpoj et al.,

2002).

Deiodination of Thyroid Hormones

The delivery of the predominant circulating TH (T4) to specific target tissues (i.e.,

liver, choroid plexus) is critical for the subsequent conversion of T4 to T3; which is

considered the principal, biologically-active form of TH. The majority of systemic T3

availability for multiple taxa is generated via extrathyroidal mechanisms in these target









tissues utilizing a process known as deiodination (Brown et al., 2004; McNabb, 2000).

The process of deiodination is catalyzed by a family of selonoenzymes called

deiodinases. These membrane-bound enzymes are located primarily in the microsomal

fraction of tissue homogenates suggesting an endoplasmic reticulum and/or plasma

membrane location (Hulbert, 2000). T4 is deiodinated by removal of iodine from the

outer ring of the molecule (ORD) to produce T3 or the inner ring of the molecule (IRD)

producing reverse T3 (rT3). ORD and IRD are catalyzed by three distinct deiodinases.

Type I catalyzes both ORD and IRD by preferentially removing phenolic and tyrosyl

iodide. This type of deiodinase is probably located in all tissues but has especially high

activity in the liver, kidney, thyroid tissue, and the central nervous system. Type II,

catalyzes only ORD by removing only phenolic iodide and has been found in the central

nervous system, brown adipose tissue, anterior pituitary and placenta. Type III catalyzes

exclusively IRD by removing only tyrosyl iodide and is found in the central nervous

system and the placenta (Shepherdley et al., 2002; Hulbert, 2000; Eales, 1984).

The integrated nature of thyroid regulation reflects a system principally regulated

by classic endocrine feedback mechanisms. In oviparous embryos, thyroid hormone

synthesis and availability are governed by a developmentally-regulated system utilizing

two sources: 1) maternal deposition in yolk (utilized during early stages of embryonic

development) and, 2) embryonic endogenous synthesis (utilized during later stages of

embryonic development). The next section details the principal mechanisms) that

regulate oviparous embryo TH availability. In addition, a brief summary is provided to

demonstrate species-differences in TH utilization and regulation.









Thyroid Hormone Availability and Synthesis among Oviparous Species

Thyroid hormone availability during embryonic and early post-natal development

in oviparous species has been principally investigated through the examination of TH

synthesis, availability, compartmentalization, functionality, and utilization during several

lifestages (Prati et al., 1992; Greenblatt et al., 1989; Tagawa and Hirano, 1987; Sullivan

et al., 1987). The principal sources of thyroid hormones for developing oviparous

embryos have been identified as maternal deposition in yolk and endogenous synthesis by

the embryo (Greenblatt et al., 1989). In salmonids, high-density lipoproteins (HDL) and

vitellogenin (VTG), a yolk precursor protein, have been identified as the major carriers of

thyroid and other hormones, vitamins, ions, and minerals from maternal circulation and

subsequent sequestering in the yolk for the developing oocyte (Monteverdi and Di Giulio,

2000; Conley et al., 1997). In addition, Prati et al. (1992) suggested that TTR from

chicken extra embryonic membranes may bind iodothyronines of maternal origin

constituting the mechanism by which THs become available to the fetus before the onset

of thyroid function. In an examination of the relationship between TH content and yolk

mass, Sechman and Bobeck (1988) observed that a linear increase in both T4 and T3

concentrations in oocytes was proportional to the weight of the yolk without changes in

the iodothyronines content per 100 mg of yolk which indicated transfer of iodothyronines

together with other yolk constituents as a principal source of TH for developing oocytes.

Greenblatt et al. (1989) examined the compartmentalization of both T4 and T3 in yolk

and larvae in coho (Oncorhynchus kisutsch) and chinook (0. tschawytscha) salmon.

These data demonstrated an asynchronous species difference in thyroid hormone

utilization versus time between yolk reserves and endogenous TH production (Sullivan et

al., 1989). However, both species demonstrated a decreasing reliance on TH yolk









reserves in step with an increase in endogenous TH production in relation to increasing

larvae development (Sullivan et al., 1989).

Species-Differences in Thyroid Hormone Utilization and Regulation

Fish

In teleosts, T4 has been reported as the primary hormone released by the thyroid

(Eales, 1985). Under TSH stimulation, Eales (1985) reported a surge in both

endogenously labeled and stable plasma T4 concentrations with no corresponding

changes in plasma T3 concentrations. Kinetic studies have shown that about 80% of T3

in salmonids may reside in a slowly exchanging reserve pool, mainly represented by

skeletal muscle (Brown et al., 2005). This constancy in plasma T3 concentrations is due

at least in part to a rapid decrease in the proportion of available plasma T4 peripherally

monodeiodinated to plasma T3 (Eales, 1985). Though total thyroxine (TT4) and total

triiodothyronine (TT3) plasma hormone concentrations have been shown to be highly

correlated with their respective free plasma hormone concentrations, both percent free

thyroxine (%FT4) and free triodothyronine (%FT3) plasma hormone concentrations

demonstrated a negative correlation with TT4 and TT3 indicating that a smaller

proportion of total hormone is free at higher total hormone concentrations (Eales and

Shostak, 1985). In general, poikilotherm plasma TH concentrations contrast with those of

both Japanese Quail and humans where %FT3 exceeds %FT4, and are 3-5x higher than

those reported in both trout and charr (Eales and Shostak, 1985).

Amphibians

Amphibian utilization of TH has been primarily reported during several critical

stages of metamorphosis (Galton and Cohen, 1980; Suzuki and Suzuki, 1980; Mondou

and Kaltenbach, 1979). At stages V-XVIII (limb differentiation), plasma T4









concentrations were undetectable suggesting that bullfrog (Rana catesbeiana) tadpoles

were responsive to very low concentrations of thyroid hormones (Mondou and

Kaltenbach, 1979). During stage XIX (forelimb emergence) through stage XXI (tail

resorption), a rapid increase was observed in both circulating plasma T4 and T3

concentrations (Suzuki and Susuki, 1981). In addition, the T3/T4 ratio of plasma TH

concentrations suggested extrathyroidal deiodination during these stages of amphibian

metamorphosis (Suzuki and Susuki, 1981). At the conclusion of metamorphosis (stages:

XXIV XXV), a rapid decline was observed in both plasma T3 and T4 concentrations in

froglets of four months of age (Suzuki and Susuki, 1981). In adult frogs, low but

detectable plasma T4 concentrations were observed (Mondou and Kaltenbach, 1979).

Avian

Birds possess the ability through the actions of thyroid hormones to regulate and

maintain thermal independence (i.e., homeothermy) (Schew et al., 1996; McNabb, 1995).

The initiation of avian thyroidal function is discriminatively observed among two

separate modes of hatchling development: precocial and altricial. Chicks of precocial

species have dramatic peaks of plasma T3 and T4 concentrations at hatching, which is

marked by the initiation of thermoregulation. By contrast, altricial chick plasma TH

concentrations are very low at hatching which is followed by a progressive increase by

the time of the greatest endothermic improvements during nestling life (McNabb 2000;

Olson et al., 1999). McNabb et al. (1991) noted in Japanese quail (Coturnix c. japonica),

a precocial species, that both plasma T4 and T3 concentrations as well as T3/T4 ratio

increased following the chick's penetration of the air cell. Thus, both TH release and

utilization in quail increase concurrently with the beginning of pulmonary respiration and

increased metabolic rate (McNabb et al., 1991). The proposed functionality of this rapid









increase in TH release and utilization during the perinatal period probably institute a level

of metabolic readiness and final maturation of the nervous system (McNabb et al., 1991).

In altricial species, a significant increase in plasma T4 concentrations have been observed

in the red-winged blackbird (Agelaiusphoeniceus) from hatching to day 8 by which

nestlings can achieve significantly large factorial increases in both instantaneous and

steady state rates of oxygen consumption in response to cold challenge (i.e., gradual

cooling) versus their younger counterparts (Olson et al., 1999). In addition, early nestling

blackbirds demonstrated increased plasma T3 concentrations which have been suggested

to be important in the organization and maturation of skeletal muscle essential for

shivering thermogenesis (Olson et al., 1999).

These data demonstrate the diverse and multifaceted roles that THs play in the

areas of growth and development among several species. In addition, thyroid regulation

as well as growth and development have been reported to be significantly influenced by

several physiological and environmental parameters. Therefore, a review of the principal

physiological and environmental effectors that have been reported to influence thyroid

regulation is provided.

Physiological and Environmental Influences on Thyroid Regulation

Overview

Several studies have reported an inter-relation between physiological and

environmental parameters and subsequent alterations in thyroid hormone regulation

among a number of species (Kohel et al., 2001; Denver and Licht, 1991; Eales, 1985).

Primarily, a seasonal, counter-regulatory system involving plasma T4 and testosterone

(T) concentrations has been suggested among several poikilothermic species. In this

system, plasma T4 generally increases in conjunction with and beyond testis growth and









subsequently regresses reproductive tissues (Bona-Gallo et al., 1980). In addition,

physiological and environmental factors such as: ambient and water temperatures,

photoperiod, nutritional availability and hibernation have been reported to play critical

roles in TH regulation among several poikilothermic and homeothermic species (Kohel et

al., 2001; Schew et al., 1996; Denver and Licht, 1991; Jallageas and Assenmacher, 1979).

Reproductive and Thyroidal Seasonal Cycles

Gonadal and thyroid seasonal cycles have been described for numerous reptile

and avian species (Hulbert, 2000; Kar and Chandola-Sakalani, 1984; Licht et al., 1984;

Bona-Gallo et al., 1980; Jallageas et al., 1978). Bona-Gallo et al. (1980) examined both

male and female cobra (Naja naja). In female N. naja, plasma T4 concentrations were

reported low in pre-vitellogenic animals, rose significantly in vitellogenic and pre-

ovulatory animals and showed only a slight decline after ovulation (Bona-Gallo et al.,

1980). Females demonstrated their greatest rise in plasma T4 concentrations during the

peak of vitellogenesis but, these were observed to be much more variable than males

(some values ranged up to 70 ng/ml) (Bona-Gallo et al., 1980). Male N. naja plasma T4

concentrations increased significantly in March-April, coincident with rapid increase in

testis weight however, plasma T4 concentrations demonstrated their greatest rise a full

month after the peak in testis weight and plasma T concentrations (Bona-Gallo et al.,

1980).These data suggest a distinct seasonality for plasma T4 concentrations in the male

cobra as plasma T4 concentrations generally increased in conjunction with and beyond

testis growth and subsequent regression (Bona-Gallo et al., 1980). Jallageas et al. (1978)

reported a strong inhibitory effect of elevated plasma T4 concentrations on sex steroid

synthesis and secretion in male Peking ducks (Anasplatyrhynchos) rather than LH

concentrations suggesting that plasma T4 concentrations may be responsible for a









seasonal state of reduced sensitivity of the endocrine testis toward circulating LH. This

suggestion, observed both in male Peking ducks and male teal (Anas creeca), was based

on the observation that the highest concentration of plasma T4 coincided with a

substantial decrease in circulating plasma T concentrations, whereas a transient rebound

of plasma testosterone concentrations (August/September) was associated with a decline

in plasma T4 concentrations (Jallageas and Assenmacher, 1979; Jallageas et al., 1978).

Licht et al. (1985) noted a seasonal peak in plasma T4 concentrations in comparison with

plasma T concentrations and follicle-stimulating hormone (FSH) concentrations in the

painted turtle (Chrysemyspicta). Following emergence in mid-March to April, C. picta

plasma T and FSH concentrations demonstrated a transient peak for about 2 weeks

followed by a decline. In contrast, plasma T4 concentrations continued to progressively

increase and did not peak until late May (i.e., the conclusion of reproductive activity).

Licht et al. (1985) noted that plasma T4 concentrations tended to fall more slowly or even

remain relatively stable in spite of the observed decline in plasma T concentrations.

Though a coincident regulatory pattern has been observed between plasma T4 and T

concentrations, Licht et al. (1985) suggests that these separate androgen and thyroid

cycles may simply reflect independent or differential responsiveness of the gonads and

thyroid to changing environmental stimuli in the temperate-zone reptiles.

Several authors have experimentally demonstrated the influence of both ambient

temperature and photoperiod as it relates to testosterone and thyroid hormone synthesis

and regulation (Jallageas and Assenmacher, 1979; Jallageas et al., 1978). In ducks and

teal, cold environments have been shown to induce increased plasma T4 concentrations

as well as moderate but, significant inhibition of plasma T concentrations (Jallageas et al.,









1978). However, these observed effects have not been determined to be a clear inhibition

of photogonadal response or merely an example of cold-induced hyperthyroidism

increasing metabolic rate and subsequent inhibition of sex steroid secretion (Jallageas et

al., 1978). Under artificial lighting conditions (20D: 4N), Wilson and Reinert (1999)

noted that female tree sparrows (Spizella arborea) demonstrated both thyroid-dependent

and thyroid-independent components that were coincident with reproductive activity.

Animals that received thyroidectomy (THX) demonstrated an inhibition of ovarian

growth by 81 to 84% in comparison to (THX) supplemented with T4 and controls.

Interestingly, ovarian growth in THX animals was still progressing whereas both THX

supplemented with T4 and controls had completed 40-50% of their postnuptial molt and

significant ovarian reduction had occurred by day 84 of treatment (Wilson and Reinert,

1999). These data suggest that both temperature and delayed expression of absolute

photorefractoriness (i.e., state of unresponsiveness to previously gonadostimulatory

daylength which terminates breeding in many photoperiodic bird species) are associated

with alterations in both reproductive and thyroid function (Wilson and Reinert, 1999;

Jallageas and Assenmacher, 1979).

Nutritional Availability and Hibernation

Schew et al. (1996) examined the relationship between food availability and TH

regulation among precocial and altricial species. Initially, birds were placed on a

maintenance diet (i.e., a limited ration of food was provided). Plasma T3 concentrations

among both species were significantly decreased not only compared to controls, but also

compared to each species' own values at the beginning of the restriction period (Schew et

al., 1996). Realignmentation (i.e., birds returned to ad libitum feeding), resulted in a

rebound of plasma T3 concentrations among both species in comparison to controls









(Schew et al., 1996). Upon emergence from their burrows, both male and female Desert

Tortoise's (Gopherus agassizi) demonstrated elevated plasma T4 concentrations with

increased feeding, activity (i.e., mating, locomotion), and warmer temperatures (Kobel et

al., 2001). Female tortoises exhibited a single, dramatic increase in plasma T4

concentrations during the spring (i.e., warmer ambient temperatures and peak

reproductive period) while males exhibited a longer plateau in plasma T4 concentrations

throughout the summer (Kohel et al., 2001). Sellers et al. (1982) noted in the lizard

(Cnemidophorus sexlineatus) significant increases in plasma T4 concentrations coincided

with the entrance and emergence of hibernation. The author's suggested that the observed

increase in plasma T4 concentrations were the result of decreased peripheral utilization of

TH.

Physiological and Environment Parameters Influence Growth

Denver and Licht (1991) examined the inter-relationship between thyroid

hormones, photoperiod, ambient temperature and growth utilizing slider turtles

(Pseudemys scripta. Animals were treated by either sham, partial (PTX) or complete

(TX) thyroidectomy (Denver and Licht, 1991). Significant reductions in plasma T4

concentrations and increased plasma TSH concentrations were observed in TX treatment

versus sham. By 8 weeks (post-treatment), TX treatment had a significant reduction in

both body mass and carapace length in comparison to sham treatment (Denver and Licht,

1991). Interestingly, partial groupings of sham PTX and TX treatment were maintained

under either constant (30C ambient temperature, 27+ 1C water temperature and constant

light) or variable (40C to 24 C ambient temperature, 190C to 24 C water temperature

and a 12L:12D photoperiod) environmental conditions (Denver and Licht, 1991). Under

constant environmental conditions, growth rates in the sham and TX treatments exhibited









a significant decline whereas growth rates of sham and TX animals under variable

conditions declined only slightly by week 14 (Denver and Licht, 1991). These data

demonstrate the profound influence of both physiological and environment parameters on

brain-pituitary-thyroid axis regulation (Denver and Licht, 1991).

OC contaminants have been reported to alter thyroid regulation producing

deleterious effects in the areas of growth and development. As alligators have exhibited

alterations in growth and survivability in relation to OC exposure, a review is provided

demonstrating reported alterations in TH synthesis, deiodination, delivery, activity,

metabolism and excretion in relation to OC exposure.

Effects of Organochlorine Contaminant Exposure on Thyroid Regulation

Overview

Thyroid hormones are one of the principal regulators of diverse physiological

endpoints including: metabolic rate, oxygen consumption, tissue differentiation, and

subsequent embryonic and post-natal growth and development. However, these endpoints

have been shown to be highly influenced by a variety of physiological and/or

environmental influences including but not limited to nutritional state, ambient

temperature, photoperiod, and potentially coincident counter-regulation by hypothalamic-

pituitary cascades involved in reproductive tissue development and subsequent

reproductive quiescence. Currently, environmental research has been examining the

influence of introduced chemical compounds (i.e., environmental contaminants) which

have been suggested to alter thyroid function, a growth-regulating endocrine tissue

(Brouwer et al., 1998). Many of the observed actions of environmental contaminants

have been reported to occur during embryonic development and sensitive early life stages

resulting in impaired reproduction and developmental abnormalities in the offspring









(Guillette, 1995). These chemicals have been referred to as "endocrine disrupters" or

exogenous agents that interfere with the production, release, transport, metabolism,

binding, action, or elimination of natural hormones in the body responsible for the

maintenance of homeostasis and regulation of developmental processes (Rolland, 2000;

Brucker-Davis, 1998). Due to the reported structural similarity among THs and

chlorinated hydrocarbons (i.e., DDT, PCBs and dioxins), it has been hypothesized that

these chemicals may elicit alterations in several areas of TH regulation including: TH

synthesis, deiodination, delivery, activity, metabolism and excretion (Brucker-Davis,

1998; Porterfield, 1994). Therefore, OC contaminant exposure may contribute to the

observed alterations in alligator embryonic and hatchling growth, development and

survivability. In order to examine this relationship in greater detail, a detailed review is

provided which 1) provides the current information on alligator thyroid regulation and

growth in relation to OC exposure, 2) presents reported alterations in both thyroid

histology and regulation among several species in OC contaminated environments, 3)

demonstrates the potential disruptive influence OC contaminants may have at all levels of

thyroid regulation, and 4) provides experimental data that demonstrate alterations in

growth in relation to exposure by the four primary OC compounds identified among OC

contaminated sites in central Florida: p,p'-DDE, dieldrin, chlordane and toxaphene.

Effects of Organochlorine Contaminant Exposure on Alligator Thyroid Regulation

American alligators (Alligator mississippiensis) have been considered a

particularly suitable indicator species as they have been shown to bioaccumulate and

biomagnify contaminants to levels equal to or greater than reported in birds and

mammals (Crain and Guillette, 1998). However, an understanding of alligator thyroid

function is limited as the principal data available is in relation to OC exposure









(Gunderson et al., 2002; Crain et al., 1998). Crain et al. (1998) noted a negative

relationship with both plasma T3 and T4 concentrations and body size among male and

female animals from lake Woodruff (low OC). However, a general lack of correlation

between plasma TH concentrations, sex and body size was observed in sub-adult

alligators from both lakes Apopka and Okeechobee (Crain et al., 1998). These data may

potentially reflect altered reproductive potential in these animals, as THs cooperatively

regulate the reproductive activities of vertebrates (Crain et al., 1998). Gunderson et al.

(2002) and Hewitt et al. (2002) reported on sub-adult (0.9 to 1.5 m) alligator plasma T4

concentrations and quantitatively assessed sub-adult alligator thyroid function in sites

with varying degrees of OC contamination in south Florida (Belle Glade > WCA3A >

Moonshine Bay). No obvious relationship was observed between body size and plasma

T4 concentrations (Gunderson et al., 2002). Data generated from combined sampling

years demonstrated that WCA3A had significantly higher plasma T4 concentrations than

either Belle Glade or Moonshine Bay (Gunderson et al., 2002). In addition, no

differences in plasma T4 concentrations were observed between Belle Glade and

Moonshine Bay (Gunderson et al., 2002). However, significant differences were observed

between Belle Glade versus Moonshine Bay in epithelial width and colloid content

(Hewitt et al., 2002). The author's suggest an interrelation between the observed

reduction in colloid content and reduced plasma T4 concentration observed in Belle

Glade animals. Therefore, reductions in the observed plasma T4 concentrations may be

related to OC competition with TH for binding proteins as well as elevation of UDP-GT

enzymatic activity which induces T4 glucuronidation and subsequent biliary hormone

excretion. The inter-regulatory actions of both OC contaminant affinity for TH binding









proteins and biliary TH excretion may have led to the equivalent plasma T4

concentrations observed between Belle Glade and Moonshine Bay (Hewitt et al., 2002).

Thyroid Histology Alterations in Relation to Organochlorine Contaminant
Exposure

Several field-oriented studies utilizing both qualitative and quantitative

methodologies have provided insight as to the potential interrelation between

environmental contaminant exposure and observed pathological thyroidal alterations

among several species (Zhou et al., 1999; Moccia et al., 1986; Moccia et al., 1981;

Sonstegard and Leatherland, 1976). Sonstegard and Leatherland (1976) noted that coho

salmon (Oncorhynchus kisutch) from several Great Lakes had increased incidence of

goiter (distinct growths located on the gill arches) and diffuse swelling at the base of the

gill arches which is indicative of thyroid neoplasia. Oblate or extremely elongated thyroid

follicles with thickened, columnar shaped epithelial and extensive colloid vacuolation

were observed among spawning coho (0. kisutch) and chinook ( 0. tschawytscha)

salmon among the Great Lakes in comparison with the Fraser River (control site)

(Moccia et al., 1981). In addition, dense aggregations of thyroid microfollicles were

observed in many of the Great Lakes salmon (Moccia et al., 1981). In order to assess the

degree of observed thyroid hyperplasia, Moccia et al. (1981) developed a thyroid index

for inter-lake and inter-species comparisons among the two salmon species. These data

demonstrated a significant correlation between the thyroid index and observed goiter

frequencies in the coho salmon (Moccia et al., 1981). The author's reported a 13-fold

difference in goiter frequency among Great Lake coho salmon populations (Moccia et al.,

1981). Though the Great Lakes region has previously been documented with reduced

iodine availability, the documented incidence of goiter has been reported to fluctuate over









several years demonstrating goiters are not solely due to low iodine availability but, may

be attributed to the presence of organochlorine contaminants (principally: PCB

congeners) in the environment (Moccia et al., 1981; Sonstegard and Leatherland, 1978).

In order to examine the direct effects of environmental contaminants and

subsequent thyroid hyperplasia, Zhou et al. (1999) quantitatively evaluated mummichogs

(Fundulus heteroclitus) exposed to high sediment concentrations of PCBs, PAHs, DDT

and various metals (Mercury, Lead, Copper, Zinc, Chromium, Cadmium) under both

field and captive conditions. The author's reported greater epithelial height, larger

follicles, and partially depleted colloid in fish from the contaminated site (PC) in

comparison with the control site (TK) (Zhou et al., 1999). Both male and female fish

from (PC) demonstrated a greater liver somatic index (LSI) in comparison with animals

from (TK) (Zhou et al., 1999). The author's suggested that LSI may be utilized as a

biomarker of extrathyroidal T4 conversion (Zhou et al., 1999). Fish (male and female)

from PC demonstrated significantly higher plasma T4 concentrations versus TK, which is

different than what would typically be observed in goiterous fish (Zhou et al., 1999). No

significant differences were observed in plasma T3 concentrations among PC and TK fish

(Zhou et al., 1999). A captive reciprocal environment experiment was conducted

utilizing animals and sediment from both contaminated and control environments (Zhou

et al., 1999). These data suggest that the simulated PC environment could elevate plasma

T4 concentrations in TK fish, whereas an unpolluted environment could reduce plasma

T4 and T3 concentrations in PC fish (Zhou et al., 1999). However, conditions of goiter

as noted by Sonstegard and Leatherland (1978) were not observed in fish under field or

experimental conditions (Zhou et al., 1999).









Accumulation and biomagnification of high concentrations of lipophilic,

polyhalogenated hydrocarbons has been suggested as an additive cause for the observed

thyroid hyperplasia in several salmonid species among the Great Lakes region

(Sonstegard and Leatherland, 1978). Adult herring gull (Larus argentatus), a non-

migratory, piscivorous bird of the Great Lakes region were utilized to quantitatively

examine the incidence of thyroidal hyperplasia in relation to dietary environmental

contaminant exposure (Moccia et al., 1986). Great Lakes herring gulls demonstrated

predominantly microfollicular follicles, enlarged epithelial height, limited/no colloid

versus established controls (Bay of Fundy) which displayed normal thyroid morphology

(Moccia et al., 1986). Many of the microfollicular thyroids from Great Lakes herring

gulls also had a severely hyperplastic epithelial component (Moccia et al., 1986). These

data in conjunction with Moccia et al. (1981) demonstrated diffuse, microfollicular

hyperplasia in both herring gulls and salmon in the Great Lakes region (Moccia et al.,

1986). The author's noted the increased prevalence of diffuse, microfollicular hyperplasia

in most of the Great Lake collections and its absence in similar collections from the Bay

of Fundy (control site) which are relatively free of environmental contaminants (i.e.,

lipophilic organohalogens) is consistent with the existence of thyrotoxic factors in the

Great Lakes food chain (Moccia et al., 1986).

Influence of Organochlorine Contaminant Exposure on Integrated Levels of
Thyroid Hormone Regulation

Thyroid Hormone Synthesis

A wide variety of chemicals, drugs and other xenobiotics have been determined to

affect thyroid hormone biosynthesis. A number of anions act as competitive inhibitors of

iodide transport in the thyroid, including perchlorate, thiocyanate, and pertechnetate









(McNabb et al., 2004; Capen, 2001). In addition, several classes of chemicals have been

identified that inhibit the organification of thyroglobulin including: 1) thionamides

(thiourea, thiouracil, PTU), 2) alanine derivatives (sulfonamides), 3) substituted phenols,

4) and miscellaneous inhibitors (aminotriazole) (Capen, 2001). Many of these chemicals

have been reported to exert their action by inhibiting thyroperoxidase, responsible for

iodide oxidation to iodine, which results in the disruption of both iodination of tyrosyl

residues in thyroglobulin and also the coupling reaction of iodotyrosines (i.e., MIT and

DIT which form iodothyronines: T3 and T4) (Capen, 2001; McNabb, 2000).

Thyroid Hormone Binding Proteins

Concomitant reduction in plasma T4 concentrations has been reported in some

cases to be an indication of compromised plasma transport system for both ligands and of

the presence of hydroxylated PHAHs on the TTR protein (Brouwer et al., 1998). Cheek

et al. (1999) noted that hydroxylated PCBs are potent ligands for TTR, having affinities

in the 1 nM range, 50-fold greater than that of T4. TTR is a major T4 binding protein in

the blood, and it shows in addition to the thyroxine binding sites a site that is

complimentary to the DNA double helix, indicating a possible relationship to the

thyroxine nuclear receptor (Rickenbacher et al., 1986). The TTR molecule has two-fold

symmetry, and the binding site is lined primarily with hydrophobic amino acid side

chains that form polarizable pockets for halogen interactions (Rickenbacher et al., 1986).

In view of the highly hydrophobic/polarizable nature of the TTR binding site, the

author's suggest that van der Waals / hydrophobic interactions would be dominant in

controlling the binding strength of biphenol compounds (Rickenbacher et al., 1986).

Contaminants with the highest TTR binding efficiencies were shown to have apara

hydroxyl substituent flanked by two meta chlorines which is analogous to the









diiodophenolic ring system in T4 (Rickenbacher et al., 1986). Van den Berg et al.

(1991) noted that chlorophenols demonstrated the highest level of competition for TTR

binding utilizing a competition assay (i.e., radiolabelled T4, TTR versus individual

contaminant). These data suggest that 1) interaction with the T4 binding site is dependent

on the degree of chlorination, 2) the combination of hydroxyl and chlorine groups is more

competitive than either group separately, and 3) displacement of T4 from the binding site

is by a competitive type of interaction (Van den Berg et al., 1991). The author's noted

that DDT isoforms such as p, p'-DDD, o, p'-DDD as well as dicofol, in particular, were

found to interact with TTR (Van den Berg et al., 1991). A large proportion of the

chemicals with affinity for TTR appear to have neurotoxic properties (Van den Berg et

al., 1991). In addition, transthyretin has been reported as one of the few proteins

identified in the cerebrospinal fluid (CSF) that is synthesized by the choroids plexus and

may function in the transport of T4 through the blood-CSF barrier (Van den Berg et al.,

1991). Therefore, chemicals interacting with TTR may affect the transportation function

of the choroids plexus with possible consequences on brain function (Van den Berg et al.,

1991).

Deiodination of Thyroid Hormones

iodothyronine deiodinase activity is principally responsible for TH conversion in

extrathyroidal tissues has been suggested as a more sensitive thyroidal index of

contaminant exposure (Adams et al. 2000). Male plaice dosed (ip) with 5 ng PCB 77 / g

body mass demonstrated reduced plasma T4 and T3 concentrations as well as increased

hepatic T4 ORD activity during week one versus week four post-exposure (Adams et al.,

2000). Coimbra et al. (2005) noted that Nile Tilapia receiving dietary treatments (0.1 mg

Endosulfan / g -1 of food (EL), 0.5 ig Endosulfan / g -1 of food (EH), or 0.5 ig Arochlor









1254 / g -1 of food (A)) demonstrated alterations in both plasma T4 and ORD activity

(time points: days 21 and 35). Tilapia exposed to EL21 demonstrated lower plasma T4

concentrations than either EH (days 21 and 35), A (days 21 and 35), and control

treatments (Coimbra et al., 2005). Plasma T3 concentrations were not significantly

altered in any treatments (Coimbra et al., 2005). Liver DI ORD activity was found to be

depressed by both EL treatments while liver D3 activity was found to be enhanced by the

EL treatment in relation to time of exposure (Coimbra et al., 2005). The observed

changes in the activity of several deiodinases could result in decreased plasma T3

availability (Coimbra et al., 2005). The fact that plasma T3 concentrations remained

unaltered, is probably indicative of the prominent role of hepatic D2 activity and renal Dl

activity, both of which remained stable (Coimbra et al., 2005).

Thyroid Hormone Excretion

Hepatic microsomal enzymes (specifically: uridine diphosphate

glucuronsyltransferase UDP-GTs) play an important role in thyroid hormone

economy/availability which is accomplished in part through glucuronidation (a rate-

limiting step in the biliary excretion of T4) and sulfation (which utilizes phenol

sulfotransferase for the excretion of T3) (Capen, 2001). Glucuronidation and sulfation are

responsible, in part, for the conversion/mobilization of aglycones (parent compounds or

phase I metabolites) into water-soluble conjugates that can be subsequently excreted from

the body (Parkinson, 2001). Sulfation and desulfation appear to be very important

pathways to regulate free TH concentrations in the fetal compartment (Brouwer et al.,

1998). Since hydroxylated PCBs tend to accumulate in the fetal department, where

sulfation is a major regulation pathway, it is hypothesized that the fetal regulation of free









TH concentrations may be compromised by PHAHs which may have serious negative

consequences for fetal and neonatal development (Brouwer et al., 1998).

Several xenobiotics have been reported to induce microsomal enzymes and

disrupt function in rats including: CNS-acting drugs (phenobarbital) and chlorinated

hydrocarbons (i.e., chlordane, DDT, and TCDD) and polyhalogenated biphenyls (PCB,

PBB) (Capen, 2001). McClain et al. (1989) provided a detailed assessment of hepatic T4-

UDP-glucuronyl transferase activity in phenobarbital-treated rats. A significantly higher

cumulative biliary excretion of 125I-labeled T4 was observed in rats orally treated with

phenobarbital versus controls bile (McClain et al., 1989). The observed increase in biliary

excretion was accounted for by an increase in T4-glucoronide resulting from increased

T4 metabolism (McClain et al., 1989). This was consistent with enzymatic activity

measurements which resulted in increased hepatic T4-UDP-glucuronyl transferase

activity (McClain et al., 1989). In addition, histological alterations including: follicular

cell hypertrophy followed by hyperplasia in association with both a marked increase in

biliary T4 excretion and sustained increases in TSH (McClain et al., 1989). These data

are consistent with the hypothesis that the promotion of observed thyroid tumors in rats is

not a direct effect of phenobarbital treatment on the thyroid gland but rather an indirect

effect mediated by plasma TSH concentrations secreted from the pituitary secondary to

the hepatic microsomal enzyme -induced increase of T4 excretion in the bile (McClain et

al., 1989). In addition, significant species differences in UDP-GT expression have been

observed between rats and mice exposed to the PCB, Kanechlor-500 (K-500) (Kato et al.,

2003). Though K-500 treatment resulted in a significant decrease in plasma T4

concentrations in both rats and mice, a significant increase in UDP-GT activity was









observed only in the rat (Kato et al., 2003). These data were further supported following

K-500 treatment as gene expression of hepatic UDP-GT isoforms UGT1A1 and

UGT1AG in the rat liver were enhanced prior to the decrease in plasma T4

concentrations as opposed to the mouse liver (Kato et al. 2003). Utilizing Gunn rats

(UGT1A deficient) and Winstar rats (normal), Kato et al. (2004) dosed both species with

KC-500 and 2,2',4,5,5'-Pentachlorobiphenyl (PentaCB) examining deiodinase activity

and additional mechanisms of biliary excretion of thyroid hormones. Plasma total T4 and

free T4 concentrations were significantly decreased in both PCB treated species (Kato et

al., 2004). In addition, type I deiodinase activity (converts T4 to T3) in Winstar rats was

significantly decreased by KC-500 but not by PentaCB, although in Gunn rats, it was

significantly decreased by both PCB isoforms (Kato et al., 2004). These data led the

author's to suggest that PCB-mediated decrease in plasma T4 concentrations does not

occur through the induction of hepatic T4 glucuronidation enzymes (Kato et al., 2004).

These conflicting reports regarding UDP-GT activity prompted several authors to suggest

potential mechanisms/factors that may individually/collectively reduce plasma T4

concentrations including: displacement of T4 from transthyretin (TTR) binding by PCBs

facilitating free T4 excretion in urine or bile, alteration in the HPT axis, and/or increase

in estrogen sulfotransferase, which efficiently catalyzes the sulfation of iodothyronines

(Kato et al., 2004, McNabb and Fox, 2003).

Growth in Relation to p,p'-DDE, dieldrin, chlordane and toxaphene exposure

Overview

Several PCBs and organochlorine pesticides (i.e., DDE, dieldrin, chlordanes, and

toxaphene) have been suggested to alter thyroid regulation in several species under

experimental (in-ovo and in-vivo) conditions (Scollon et al., 2004; Nishimura et al., 2002,









Willingham, 2001; Waritz et al., 1996; Jefferies and French, 1972). These OC pesticides

have been previously identified in both alligator maternal tissues and egg yolk which

have been associated with alterations in alligator egg and embryo qualities as well as

hatchling growth among several contaminated lakes and reclaimed agricultural properties

in central Florida (Rauschenberger, 2004; Wiebe et al., 2002). TH regulation and

alterations in thyroid histology in relation to OC exposure have been primarily examined

utilizing pharmacological dosing methodologies. A consistent observation among several

controlled treatment studies was thyroid gland histological alterations consisting of

increases in overall thyroid weight, epithelial hyperplasia and colloid depletion in relation

to exposure by several PCBs and/or OC pesticides among several species (Fowles et al.,

1996; Jefferies and French, 1972; Jefferies and Parslow, 1972; Fregly et al., 1967).

Experimental Data

As thyroid hormones are an integral component in embryonic and hatchling

growth, the observed thyroidal alterations in relation to chlorinated hydrocarbon exposure

suggest the potential for subsequent growth alterations. O'Steen and Janzen (1999)

reported that plasma TH concentrations and resting metabolic rate in hatchling snapping

turtles (Chelydra serpentina) correlated with incubation temperature. As incubation

temperature is strongly linked with sex determination in many reptile species, compounds

that mimic or antagonize steroid hormones may affect metabolism, TH concentrations, or

growth rate (Willingham, 2001). Red Eared Slider (Trachemys elegans) eggs were

topically treated prior to the temperature-sensitive window of sex determination (Stage

17, from Yntema, 1968) with low, intermediate, and high concentrations of either trans-

Nonachlor and chlordane or p,p'-DDE (Willingham, 2001). Upon hatching, hatchling

turtles were fasted for 28 days and subsequently re-fed ad-libitum for 14 days









(Willingham, 2001). At the conclusion of a 28 day fast, the intermediate trans-Nonachlor

group significantly lost mass in comparison with controls (Willingham, 2001). Following

re-feeding, both the intermediate and high trans-Nonachlor groups significantly increased

in mass (Willingham, 2001). The author suggests that the reduction in mass observed in

several OC treatments may have elicited a temporal, hyperthyroid state in which yolk

reserves were utilized more quickly, thus reducing overall mass. As was observed in

Schew et al. (1996) following fasting and ad-libitum re-feeding, compensatory increases

in mass were observed in several trans-Nonaclor and p, p'-DDE treatment groups

(Willingham, 2001). Janz and Bellward (1996) examined in-ovo exposure of 2,3,7,8-

tetrachlorodibenzo-p-dioxin (TCDD) in precocial (chicken), semi-altricial (great blue

heron) and altricial (pigeon) species and subsequent alterations in growth and

development (Janz and Bellward, 1996). In both chickens and great blue herons, no effect

in plasma TH concentrations or hatchling growth and development was observed in

relation to TCDD exposure (Janz and Bellward, 1996). However, pigeons exposed to

TCDD demonstrated significant reductions in both plasma TH concentrations and

hatchling growth and development decreases including: crown-rump length, wing length,

and tibia length (Janz and Bellward, 1996). These data are reaffirmed by the established

temporal differences in TH maturation among precocial and altricial species (McNabb

2000, Olson et al., 1999). Hatchling Artic Glaucous Gulls (Larus hyperboreous) growth

was assessed in relation to parental bird serum OC concentrations over a three year

period (Bustnes et al., 2005). Adult female gulls with high OC burdens spent significantly

longer time periods in search of nutritional resources for their chicks (Bustnes et al.,

2005). In addition, a significant negative relationship was reported between chick growth









and increasing adult OC serum concentrations of HCB, oxychlordane, p,p'-DDE, and

several PCBs (Bustnes et al., 2005).. The author's suggest that there may be interactions

between energy expenditure and different OC concentrations, and females with high OC

concentrations may have fewer resources available to provide for their chicks (Bustnes et

al., 2005). In addition, significant reductions in weight were observed in juvenile (- 37

day old) Nile Tilapia (Oreochromis niloticus) exposed to aqueous dieldrin (1.0 to 2.4

.g/liter-) for 30 days in comparison with controls (Lamai et al., 1999). Finally, Blanar et

al. (2005) noted that juvenile Artic Charr (Salvelinus alpinus) orally dosed (lx) with

toxaphene (10 pg/g) demonstrated decreased growth and overall body condition (k) as

well as decreased muscle lipid and protein content. These reports suggest the potential

direct (i.e., feeding, injection, aqueous OC exposure) and indirect (i.e., reduced parental

fitness due to OC exposure) influences that OC contaminants may have to influence

growth among several oviparous species.

These data suggest that OC exposure can elicit alterations in both thyroid function

and subsequent growth. Several field studies have reported severe alterations in both

plasma T4 concentrations and thyroid histology in relation to OC contaminated

environments among avian and several fish species (Rolland, 2000). In addition,

controlled treatment studies utilizing either p, p' DDE, dieldrin, chlordane, or toxaphene

reported altered thyroid regulation and growth reduction. These data suggest that OC

exposure may be related to the observed reductions in alligator embryo and hatchling

growth from OC contaminated sites in central Florida (Rauschenberger, 2004; Wiebe et

al., 2001; Gross et al., 1994). In addition, several authors have reported modified alligator

thyroid function in relation to OC exposure (Hewitt et al., 2002; Crain et al., 1998). These









reported modifications have taken the forms of reductions in plasma T4 concentrations

and changes in thyroid histology compared with controls. However, researchers must be

keenly aware of both physiological (i.e., sex, age, nutritional availability, reproduction,

hibernation) and environmental factors (i.e., ambient and water temperatures and

photoperiod). These factors have been reported to vary thyroid regulation and may

complicate data interpretation regarding OC exposure and subsequent alterations in

thyroid function. Therefore, a captive study providing a controlled, structured

environment presents a more applicable means to test the relationship between OC

exposure and subsequent differences in hatchling thyroid function and growth.

Organochlorine Contaminant Exposure and Hatchling Alligator Growth

Wiebe et al. (2002) evaluated hatchling alligator thyroid regulation and growth

from several lakes in central Florida: lakes Apopka (high OC concentrations), Griffin

(Intermediate OC concentrations), and lake Lochloosa (Low OC concentrations) under

captive conditions for a period of 6 months. These experimental conditions included: a

restricted photoperiod (12D:12N), controlled ambient and water temperatures, ad-libitum

feeding twice a week, and restricted number of animals per enclosure to limit stressful

overcrowding. Though egg viability rates did not differ among sites, lake Apopka

hatchlings demonstrated a significantly higher growth rate and plasma T3 and T4

concentrations in comparison with lakes Griffin and Lochlooosa. These data suggest that

lake Apopka hatchlings demonstrated a hyperthyroid secretary pattern resulting in an

enhancement of hatchling growth in relation to exposure to high OC concentrations.

However, OC contaminants, due to their structural similarity with THs, have been

predominantly suggested to reduce TH systemic availability by competing for binding

proteins. These conflicting data suggest the need for further examination of thyroid









regulation among hatchling alligators exposed to OC contaminants. Specifically,

hatchlings from a site of similar OC contaminants and concentrations (i.e., Emerelda

Marsh Conservation Area) to lake Apopka should be utilized in a comparative growth

study. A comparison of hatchling thyroid regulation and growth among several sites with

high OC concentrations may provide further insight (i.e., OC exposure versus site-

specific variables) into the observed hyperthyroid secretary pattern and accelerated

growth rate observed in lake Apopka hatchlings. Therefore, a captive hatchling growth

study was undertaken utilizing animals from lakes Apopka, Griffin as well as Orange (a

site with low OC concentrations) and Emeralda Marsh Conservation Area (Area #7) to

assess if in-ovo exposure to high concentrations of OC contaminants elicits a

hyperthyroid secretary pattern that accelerates hatchling alligator growth. The following

hypotheses were tested by this study.

Hypothesis #1

Ho: No change in hatchling growth rates will be observed among all sites in relation to

high in-ovo OC contaminant exposure.

Ha: In-ovo exposure to high concentrations of OC pesticides will accelerate hatchling

alligator growth rates in comparison with animals exposed to intermediate to low in-ovo

OC concentrations.

Hypothesis # 2

Ho: No change in hatchling TH secretary pattern will be observed among all sites in

relation to high in-ovo OC contaminant exposure.






36


Ha: In-ovo exposure to high concentrations of OC contaminants will elicit a hyperthyroid

secretary pattern in hatchling alligators that will result in an accelerated growth rate in

comparison with animals exposed to intermediate to low in-ovo OC concentration














CHAPTER 2
MANUSCRIPT

Introduction

During the 1980's, significant reductions in American Alligator (Alligator

mississippiensis) egg viability were observed on Lake Apopka (high OC concentrations)

in comparison with lake Woodruff, a national wildlife refuge (low OC concentrations)

(Woodward, 1993; Rice et al., 1998). In addition, a severe (- 90%) reduction in the

juvenile alligator population was observed on Lake Apopka (1981-1986) that was likely

attributed to reproductive failure (Woodward, 1993). The observed reductions in juvenile

survivability and adult reproductive success have been attributed in part to the influence

of agriculture and anthropogenic alterations specifically: extensive utilization of

organochlorine pesticides by muck farming, citrus crops, and effluent discharges from

both the citrus processing plant and sewage treatment facility located at the city of Winter

Garden (Woodward et al., 1993; Schelske and Brezonik, 1992). These environmental

alterations were compounded by the overflow of a wastewater pond located at the Tower

Chemical facility, adjacent to the Gourd Neck region of Lake Apopka (1980), consisting

of high concentrations of sulfuric acid, DDT, dicofol and several unidentified OC

compounds. This event resulted in the EPA designation of this property as a superfund

site in 1983 (Rauschenberger, 2004). Though several of these OC compounds were

identified in yolk from alligator eggs, no clear association with reduced clutch viability

was observed for specific OC contaminants (Rauschenberger et al., 2004, Heinz et al.

1991). Therefore, sites that have been historically impacted by varying degrees of OC









contamination continue to demonstrate coincident alterations in reproductive function as

measured by sex steroid biomarkers, sexual differentiation, clutch viability, embryonic

mortality, post-hatch growth and survivability (Rauschenberger, 2004; Wiebe et al.,

2002; Guillettte et al., 1999; Gross et al. 1994).

Guillette (1995) suggested that many of the observed embryonic and post-natal

alterations in offspring viability are the result, in part, of parental exposure to

environmental contaminants. OC exposure has been reported to alter hormones that

control the course of growth and development and may have the potential to alter

differentiation of major organ systems resulting in physiological and morphological

changes (Rauschenberger et al., 2004; Wu et al., 2000; Guillette et al., 1995). Significant

alterations in alligator clutch viability and embryonic and post-hatch survivability have

been reported among sites of intermediate to high OC concentrations, suggesting an inter-

relationship between in-ovo OC exposure and subsequent reductions in embryonic and

hatchling survivability (Wiebe et al., 2001). The predominant exposure route for

developing offspring would be maternal transfer of OC contaminants among yolk

constituents (Rauschenberger et al., 2004; Wu et al., 2000). Rauschenberger (2004) noted

that eggs collected from OC contaminated sites had higher fecundity, lower average

clutch mass and reduced clutch viability in comparison with sites with low OC

contamination (Rauschenberger, 2004). The observed alterations in embryo morphology

appear to be in association with variation in OC contaminant burdens of eggs. In addition,

OC analyte composition was determined to be equally as important as concentration,

suggesting the importance of mixture composition (Rauschenberger, 2004). These data

demonstrate a need to better understand the physiological and/or chemically-induced









mechanisms that may effect alligator growth and development from OC contaminated

sites (Rauschenberger, 2004; Wiebe, 2001).

One of the principal regulators of growth and development among multiple

taxonomic groups are thyroid hormones (TH) which have been demonstrated to regulate

diverse physiological endpoints including: metabolic rate, tissue differentiation and

subsequent growth and development (Rousset and Dunn, 2004). Several literature

reviews have suggested that alterations in thyroid function may be in relation to exposure

to a variety of compounds including OC contaminants (Rolland, 2000; Brucker-Davis,

1998). Due to the structural similarities between THs and DDT, PCB's and dioxins,

these chemicals may act as weak agonists that have the potential to reduce/block thyroid

hormone activity (Brucker-Davis, 1998; Porterfield, 1994). French and Jefferies (1972)

however, noted that pigeons fed low concentrations of p,p'-DDE and dieldrin induced

hyperthyroidism whereas higher doses of both OC contaminants caused hypothyroidism.

Therefore, thyroid regulation alterations due to OC contaminant exposure may have

contributed to the observed variation in alligator embryo and hatchling growth and

development.

American alligators (Alligator mississippiensis) have been considered a

particularly suitable indicator species as they have been shown to bio-accumulate and

biomagnify contaminants to levels equal to or greater than reported in birds and

mammals (Crain and Guillette, 1998). However, an understanding of alligator thyroid

function is limited as the principal data available is in relation to OC exposure

(Gunderson et al., 2002; Crain et al., 1998). Crain et al. (1998) noted a negative

relationship with both plasma T3 and T4 concentrations and body size among male and









female lake Woodruff (low OC) animals. However, a general lack of correlation between

plasma TH concentrations, sex and body size was observed in sub-adult alligators from

both lakes Apopka and Okeechobee (Crain et al., 1998). The author's suggested that

these data may potentially reflect altered reproductive potential in these animals, as THs

cooperatively regulate the reproductive activities of vertebrates (Crain et al., 1998).

Gunderson et al. (2002) and Hewitt et al. (2002) reported significantly higher plasma T4

concentrations among sub-adult alligators exposed to intermediate OC contaminant

concentrations versus animals from either high and low OC environments. In addition,

the author's observed no relationship between body size and plasma TH concentrations.

The author's suggested that the observed reductions in plasma T4 concentrations from

animals located at the site of high OC contamination may be related to OC competition

with TH for binding proteins as well as elevation of UDP-GT enzymatic activity which

induces T4 glucuronidation and subsequent biliary TH excretion.

Alterations in thyroid regulation in relation to OC exposure have been reported to

cause reductions in growth. Red Eared Slider (Trachemys elegans) eggs topically treated

with trans-Nonachlor significantly lost mass in comparison with controls (Willingham,

2001). The author suggested that the reduction in mass may be the result of a temporal,

hyperthyroid state in which yolk reserves were utilized more quickly, thus reducing

overall mass. A significant negative relationship was reported between Artic Glaucous

Gull (Larus hyperboreous) hatchling growth chick and increasing adult OC serum

concentrations of HCB, oxychlordane, p,p'-DDE, and several PCBs (Bustnes et al.,

2005). Juvenile Nile Tilapia (Oreochromis niloticus) exposed to aqueous dieldrin for 30

days demonstrated significant reductions in weight in comparison with controls (Lamai et









al., 1999). In addition, Blanar et al. (2005) reported that juvenile Artic Charr (Salvelinus

alpinus) orally dosed with toxaphene demonstrated decreased growth and overall body

condition (k). These reports suggest the potential direct (i.e., feeding, injection, aqueous

OC exposure) and indirect (i.e., reduced parental fitness due to OC exposure) influences

that OC contaminants may have to influence growth among several oviparous species.

Several authors have reported altered alligator thyroid function in relation to OC

exposure (Hewitt et al., 2002; Crain et al., 1998). In addition, controlled treatment studies

utilizing several OC contaminants reported both modified thyroid regulation and

subsequent growth reductions. These data suggest that OC exposure may be related to the

observed reductions in alligator embryo and hatchling growth from OC contaminated

sites in central Florida (Rauschenberger, 2004; Wiebe et al., 2001; Gross et al., 1994).

However, researchers must be keenly aware of both physiological (i.e., sex, age,

nutritional availability, reproduction, hibernation) and environmental factors (i.e.,

ambient and water temperatures and photoperiod) which have been reported to alter

thyroid regulation and may complicate data interpretation regarding OC exposure and

subsequent alterations in thyroid function. Therefore, a captive study providing a

controlled, structured environment presents a more applicable means to test the

relationship between OC exposure and subsequent alterations in hatchling thyroid

function and growth.

Wiebe et al. (2002) evaluated hatchling alligator thyroid regulation and growth

from lakes Apopka, Griffin, and Lochloosa under captive conditions for a period of 6

months. These experimental conditions included: a restricted photoperiod (12D:12N),

controlled ambient and water temperatures, ad-libitum feeding twice a week, and









restricted number of animals per enclosure. Egg viability rates did not differ among sites.

However, lake Apopka hatchlings demonstrated a significantly higher growth rate and

plasma TH concentrations in comparison with lakes Griffin and Lochlooosa. These data

suggest that lake Apopka hatchlings demonstrated a hyperthyroid secretary pattern

resulting in enhanced hatchling growth in relation to exposure to high OC concentrations.

OC contaminants, due to their structural similarity with THs, have been predominantly

suggested to reduce TH systemic availability by competing for binding proteins.

Therefore, these conflicting data suggest a need to compare hatchling thyroid regulation

and growth among several sites with high OC concentrations to provide further insight

(i.e., OC exposure versus site-specific variables) into the observed hyperthyroid secretary

pattern and accelerated growth rate observed in lake Apopka hatchlings. Therefore, a

captive hatchling growth study was undertaken utilizing animals from lakes Apopka and

Griffin as well as lake Orange (a site of low OC concentrations) and Emerelda Marsh

Conservation Area (Area #7) (a site of high OC concentrations) to assess if in-ovo

exposure to high concentrations of OC contaminants elicits a hyperthyroid secretary

pattern that accelerates hatchling alligator growth.

Materials and Methods

Egg Collection, Evaluation and Incubation

Clutches (n=10/site) were collected from lakes Apopka (N 280 35', W 81 39'),

Griffin (N 280 53', W 810 46'), and Orange (N 290 30', W 820 13') as well as Emerelda

Marsh Conservation Area (Area # 7)(N 280 55', W 81 47'). Nests were located by

aerial (helicopter) and ground airboatt) surveys. Clutches were collected and transported

in their original nesting substrate. To provide proper positioning for subsequent artificial

incubation, a black mark was placed on top of each egg to indicate the original egg









orientation in the nest. Eggs were evaluated utilizing a bright light candling procedure

(Lyon Electric, Chula Vista, CA, USA) in order to observe the presence/absence of a

calcium rich band (an indicator of developing embryos) encircling the midsection of each

egg. Each clutch was evaluated by the following measures: 1) clutch weight (Kg), 2)

fecundity (total number of eggs in clutch), 3) number of banded eggs (number of

currently viable eggs in clutch), 4) number of unbanded eggs (number of eggs with no

band which represents early embryonic mortality or lack of fertilization), and 5) number

of damaged eggs (eggs that were cracked and leaking due to nest predators or collection

error). Yolk was collected from one viable egg per clutch to assess clutch age (Ferguson,

1985) as well as identify and quantify lipophilic OC pesticide concentrations. Following

the initial clutch evaluation, the remaining banded, viable eggs from each clutch were

transferred to an incubation pan (18.5" x 14" x 7") containing moist sphagnum moss

substrate. Clutches were maintained in an artificial incubation building (13' x 11' x 7.5')

at ambient temperatures of 31.50 C + 1 C and =95% relative humidity. Individual clutch

viability (total number of hatchlings / total number of eggs collected) was assessed at the

completion of hatching. Upon hatching, external morphometrics including: total length

(mm), snout-vent length (mm), and head length (mm) (Wildlife Supply Co., Saginaw,

MI, USA; Mitutoya Calipers, Japan) and weight (g) (Ohaus, Inc., Pine Brook, NJ, USA)

were collected on each animal. In addition, a unique Monel web tag (National Band

and Tag Co., Newport, KY, USA) was provided to allow for individual animal

identification.

Clutch Selection

Three to five clutches per site were selected, based upon specific selection criteria

for this study. Clutch selection criteria included: 1) The clutch must have at least 15









hatchlings (as per the sample numbers required to satisfy the goals of the study), and 2)

Clutches were selected based on site mean yolk OC pesticide concentrations among the

four principal OCs (p,p'-DDE, dieldrin, chlordane, toxaphene) (as variance in OC

concentrations among sites limits the ability to test the direct effects of OCs on hatchling

alligator growth). Hatchlings (n=15) were randomly selected from each study-related

clutch. Prior to the studies onset, hatchlings (n=3/clutch) from all sites were sacrificed in

order to establish baseline values of free and total T4 plasma concentrations, thyroid

weight (g) (a suggested indicator of thyroid activity), and liver weight (g) (a suggested

indicator of extrathyroidal conversions of THs) (McNabb, 2004; Zhou et al., 1999). All

remaining animals selected for this study received a corresponding microchip (Biomark,

Inc., Boise, ID, USA) at the base of the tail utilizing a trocar delivery system.

Animal Maintenance

Hatchlings (n=12) per clutch were maintained for a period of eight months. Each

clutch was housed in a fiberglass tank (4' x 2' x 2') (Rowland Fiberglass, Ingleside, TX,

USA) with an aquarium heater and heat lamp to maintain uniform ambient and water

temperatures. All clutches were fed a commercial alligator diet (Burris Mill and Feed,

Franklinton, LA, USA) ad libitum twice a week.

Hatchling Morphometrics and Tissue Sampling

Hatchlings were measured once a month for a period of eight months. These

measurements include: total length (mm), snout-vent length (mm), head length (mm)

SVL, and weight (g). In addition, a 1.5 mL blood sample was taken from the cranial

sinus. Whole blood was centrifuged at 1000 x g for 10 minutes. Plasma was aliquoted

into several cryogenic vials (2 mL) and frozen at -80o C. Following the initial sampling

date (Oct 2004), a subset of hatchlings (n=3 / clutch) from all sites were sacrificed on a









quarterly schedule (Nov, Jan, Apr) to allow for a time series evaluation of thyroid and

liver activity as it relates to hatchling morphometrics and circulating free and total plasma

T4 concentrations. Sacrificed animals were selected by random number generation to

avoid researcher bias.

Plasma Thyroid Hormone Validation Procedures (Total and Free Thyroxine)

Plasma samples from alligator hatchlings were analyzed for total thyroxine (TT4)

and free thyroxine (FT4) using commercially available radioimmunoassay (RIA)

procedures. The TT4 and FT4 analyses each utilized a monoclonal solid phase

radioimmunoassay component system (MP Biomedicals Costa Mesa, CA). For the TT4

analysis, samples (50 ipl) were assayed directly as per the component system instructions.

For the FT4 analysis, sample (25 ul) were analyzed as per the provide instructions. RIA

analyses utilized iodinated (1251) ligand (L-thyroxine) and antibody coated tubes. Each

sample was analyzed in duplicate for both TT4, and FT4. Standard curves were prepared

in buffer with known amounts of radioinert T4 (0, 2, 4, 8, 12, and 20 ug/dl) or FT4 (0,

0.34, 0.64, 133, 3.27, 10.18 ng/dl). The minimum concentration distinguishable from

zero was 0.81 ug/dl for TT4 and 0.025 ng/dl for FT4 and results were listed as ng/ml for

TT4 and pg/ml for FT4. Cross-reactivities of the TT4 antiserum were; 30.9 % for D-

thyroxine; 1.0% for 3,3,5 triiodo-thyronine; and <0.1% for3,5-diodo-thyronine, 3,5-

diodo-tyrosine, 3-ido-tyrosine and phenytoin. Cross-reactivities of the FT4 antiserum

were, 91.05 % for (D-thyroxine), 7.92% for 3,3,5-triiodo-rev-thyronine, 3.05 % for 3,3,5-

triiodo-thyronine, <1.0% for 3,3-diodo-thyronine, and <0.1% for 3,5-diodo-tyrosine, 3-

iodo-tryosine, 5,5-diphenylhydantoin, sodium salicylate, acetylsalicylic acid and

phenylbutazone. A pooled sample (approximately 550 ng/ml TT4 and 480 pg/ml FT4

was assayed serially in 10, 20, 30, 40, and 50 [l volumes for Free-T4 and in 5, 10, 15, 20









and 25 ul volumes for TT4. The resulting inhibition curves were parallel to the respective

standard curve, with the tests for homogeneity of regression indicating that the curves did

not differ. Further characterization of the assays involved measurement of known

amounts (0, 2, 4, 8, 12, and 20 ug/dl) of TT4 in 25ul plasma or (0, 0.34, 0.64, 133, 3.27,

10.18 ng/dl) of TT4 in 50 of plasma. For TT4, mass recoveries were estimated as:

Y=0.16 + 0.97X, R2=0.9018; and for free-T4: Y= 0.015 + 0.96X, R2=0.8814 (Y=

amount of TT4, FT4 measured; X= amount ofTT4, FT4 added). Interassay and intrassay

coefficients of variation were 9.2 and 8.7 % respectively for plasma TT4, and 10.3 and

8.7% respectively for plasma FT4.

Free T4 (FT4) Assay Procedures

FT4 (representing <1% of available T4) is considered the most biologically

available form of thyroxine for cellular interaction. Plasma (50pL in duplicate) was added

to solid-phase coated count tubes (MP Biomedicals, Costa Mesa, CA). 1.0 mL of 125I free

T4 tracer was added to each tube. Tubes were vortexed (< 10 seconds) and incubated in

an IR-Autoflow C02 water-jacketed incubator (Nuare, Plymouth, MA, USA) at 37 1 C

for a period of 90 minutes. Contents of tubes were decanted and 1 mL of distilled water

was added / decanted to rinse each tube. Tubes were counted on a LKB-Wallac 1282

CompuGamma gamma counter (PerkinElmer, Boston, MA, USA).

Total T4 (TT4) Assay Procedures

TT4 (representing >99% of available T4) is reversibly associated with several

binding proteins including transthyretin, thyroglobulin, and albumen. Plasma (25p L in

duplicate) was added to solid-phase coated count tubes (MP Biomedicals, Costa Mesa,

CA). 1.0 mL of 125I total T4 tracer was added to each tube. Tubes were vortexed (< 10

seconds) and incubated at room temperature (18 to 250 C) for a period of 60 minutes.









Contents of tubes were decanted and counted on a LKB-Wallac 1282 CompuGamma

gamma counter (PerkinElmer, Boston, MA, USA).

Analysis of Chlorinated Analytes from Alligator Egg Yolks

Analytical grade standards for the following compounds were purchased from the

sources indicated: aldrin, a-BHC, P-BHC, lindane, 6-BHC, p,p '-DDD, p,p '-DDE, p,p '-

DDT, dieldrin, endosulfan, endosulfan II, endosulfan sulfate, endrin, endrin aldehyde,

endrin ketone, heptachlor, heptachlor epoxide, hexachlorobenzene, kepone,

methoxychlor, mirex, cis-nonachlor, and trans-nonachlor from Ultra Scientific

(Kingstown, RI); cis-chlordane, and trans-chlordane from Supelco (Bellefonte, PA);

oxychlordane from Chem Service, Inc. (West Chester, PA); o,p '-DDD, o,p '-DDE, o,p '-

DDT from Accustandard (New Haven, CT); and toxaphene from Restek Corp.

(Bellefonte, PA). All reagents were analytical grade unless otherwise indicated. Water

was doubly distilled and deionized.

Alligator egg yolk samples were analyzed for OCP content using methods modified

from Holstege et al. [1] and Schenck et al. [2]. For extraction, a 2-g tissue sample was

homogenized with -1 g of sodium sulfate and 8 mL of ethyl acetate. The supernatant

was decanted and filtered though a Bichner funnel lined with Whatman #4 filter paper

and filled to a depth of 1.25 cm with sodium sulfate. The homogenate was extracted

twice with the filtrates collected together. The combined filtrate was first concentrated to

a volume of ~2 mL by rotary evaporation, then further concentrated until solvent-free

under a stream of dry nitrogen. The residue was reconstituted in 2 mL of acetonitrile.

After vortexing (30 s) the supernatant was applied to a C18 SPE cartridge (pre-

conditioned with 3 mL of acetonitrile; Agilent Technologies, Wilmington, DE) and was

allowed to pass under gravity. This procedure was repeated twice with the combined









eluent collected in a culture tube. After the last addition, the cartridge was rinsed with 1

mL of acetonitrile which was also collected. The sample was then applied to a 0.5 g NH2

SPE cartridge (Varian, Inc., Harbor City, CA), was allowed to pass under gravity, and

was collected in a graduated conical tube. The cartridge was rinsed with an additional 1

mL portion of acetonitrile which was also collected. The combined eluents were

concentrated under a stream of dry nitrogen to a volume of 300 ptL and transferred to a

GC vial for analysis.

Analysis of the samples was performed using a Hewlett Packard HP-6890 gas

chromatograph (Wilmington, DE) with split/splitless inlet operated in splitless mode. The

analytes were introduced in a 1 [iL injection and separated across the HP-5MS column

(30 m x 0.25 mm; 0.25 [m film thickness; J & W Scientific, Inc., Folsom, CA) under a

temperature program that began at 600 C, increased at 100 C/min to 2700 C, was held for

5 min, then increased at 250 C/min to 3000 C and was held for 5 min. Detection utilized

an HP 5973 mass spectrometer in electron impact mode. Identification for all analytes

and quantitation for toxaphene, was conducted in full scan mode, where all ions are

monitored. To improve sensitivity, selected ion monitoring was used for the quantitation

for all other analytes, except kepone. The above program was used as a screening tool

for kepone which does not optimally extract with most organochlorines. Samples found

to contain kepone would be reextracted and analyzed specifically for this compound.

For quantitation, a five-point standard curve was prepared for each analyte (R2 >

0.995). Fresh curves were analyzed with each set of twenty samples. Each standard and

sample was fortified to contain a deuterated internal standard, 5 [iL of US-108 (120

[ig/mL; Ultra Scientific), added just prior to analysis. All samples also contained a









surrogate, 2 gg/mL oftetrachloroxylene (Ultra Scientific) added at homogenization.

Duplicate quality control samples were prepared and analyzed with every twenty samples

(typically at a level of 1.00 or 2.50 gg/mL ofy-BHC, heptachlor, aldrin, dieldrin, endrin,

andp,p '-DDT) with an acceptable recovery ranging from 70 130%. Repeated analyses

were conducted as allowed by matrix interference and sample availability.

Statistics

Initial RIA data was analyzed and fit four parameters logistic curve utilizing

Beckman EIA/RIA ImmunoFit software (Fullerton, CA). All statistics were performed on

SAS version 9.1 for windows (SAS Institute, Inc., Cary, NC, USA). PROC GLM

procedures including Tukey multiple comparison analysis was utilized to detect

differences (p < .05) among hatchling external morphometrics, plasma thyroid hormone

concentrations and OC contaminant concentrations between and within sites. Correlative

analysis among growth rates, plasma thyroid hormone concentration rates and OC

contaminated concentrations was performed with PROC REG procedures (p < .05).

Differences in thyroid and liver somatic indices were analyzed by the Wilcoxon Rank

Sum Test in which the Kruskal-Wallis Test was utilized to determine significant

differences among and within sites (p< .05).

Results

Clutch and Organochlorine Contaminant Parameters

Clutches (n=40) were collected from lakes Apopka (n=10), Griffin (n=10),

Orange (n=12) and Emerelda Marsh Conservation Area (n=8). Two principal clutch

parameters were utilized to select clutches for the current study: fecundity and viability.

A summary of all clutches collected demonstrated site differences among both clutch

fecundity and viability (p < .05) (Fig. 2-2). In specific, lake Apopka clutches had









significantly reduced clutch viability in comparison with the remaining sites. Selected

clutches for the current study demonstrated similar trends in clutch fecundity (p < .05)

(Fig. 2-3). However, no differences were observed in clutch viability between sites for

these select clutches (Fig. 2-4). Hatchling OC (specifically: total chlordane, total DDE,

dieldrin, and toxaphene) exposure was determined from a representative yolk sample per

clutch. Total OC concentrations per site (i.e., all clutches and growth clutches)

concentrations were distributed as follows: (EM>AP>GR>OR) (Fig. 2-4).

Hatchling Growth Rates

Hatchling growth morphometrics were monitored monthly for a period of eight

months. Multiple comparative analyses among sites demonstrated that lake Griffin

hatchlings grew significantly larger in total length, snout-vent length, head length and

weight (Fig. 2-5) (p < .05). Clutches within each site demonstrated similar trends in total

length (Fig. 2-6), snout-vent length (Fig. 2-7), head length (Fig. 2-8), and weight (Fig. 2-

9). Hatchlings (n=3/clutch/site) sacrificed on a quarterly schedule to compare thyroid and

liver weights to growth over time demonstrated similar trends among (Figs. 2-10) and

within sites in total length (Fig. 2-11), snout-vent length (Fig. 2-12), head length (Fig. 2-

13), weight (Fig. 2-14), thyroid weight (Fig. 2-15), and liver weight (Fig. 2-16) (p < .05).

No differences were observed in thyroid somatic indices among sites (Table 2-5).

However, significant temporal differences were observed in thyroid somatic indices

within sites (p< .05) (Table 2-6) Liver somatic indices demonstrated several temporal

significant differences among and within sites (p < .05) (Tables 2-5 and 2-7).

Mean growth rates were tabulated among and within sites to examine hatchling

growth per day including: total length/day (Table 2-1), snout-vent length/day (Table 2-2),

head length/day (Table 2-3) and weight/day (Table 2-4). Multiple comparative analyses









of growth rates among clutches demonstrated several significant differences among total

length, snout-vent length and head length rates (p < .05). However, analysis of growth

rates among sites again demonstrated that lake Griffin hatchlings grew larger than the

other sites (p < .05). A correlative analysis in which all clutches were independent of site

demonstrated no differences in growth rates among all sampling dates (p < .05). These

data present several isolated differences in growth rates within clutches which can

potentially be attributed to inter-clutch variability. The dominant inference taken from

both correlative and multiple comparative analyses continues to indicate no significant

differences in hatchling growth among sites.

Thyroid Hormones, Growth and Organochlorine Contaminants

Thyroid hormones (specifically: total (TT4) and free (FT4) thyroxine) were

utilized as bio-indicators of hatchling alligator growth. Multiple comparison analysis of

plasma TT4 concentrations over time demonstrated an asynchronous secretary pattern

among (Fig. 17) and within sites (Figs. 18). Similarly, plasma FT4 concentrations over

time demonstrated an asynchronous secretary pattern among (Fig. 17) and within sites

(Fig. 19). No significant alterations in either TT4 or FT4 plasma concentrations were

observed over the eight month sampling period. In addition, no paired relationship was

observed among either growth rates or any growth parameter during specific sampling

dates and plasma thyroid hormone concentrations. However, a review of monthly mean

hatchling growth parameters and plasma thyroid hormone concentration distributions

demonstrate a temporal relationship between TH secretion and subsequent hatchling

growth.

Modifications in growth and plasma thyroid hormone concentrations have been

reported in association with OC contaminant exposure among several species (Bustnes et









al., 2005, Willingham, 2001). To examine the potential interactive nature of these

experimental variables, a correlative analysis was performed utilizing hatchling growth

rates, thyroid hormone rates and the four principal OC contaminants (i.e., total chlordane,

total DDE, dieldrin and toxaphene). No significant correlative relationships were

observed (Table 2-8).

Discussion

The objective of the current study was to determine if in-ovo exposure to high

concentrations of OC contaminants elicits a hyperthyroid secretary pattern that

accelerates hatchling alligator growth. This assessment was based on several reports

indicating both alterations in thyroid function and/or subsequent growth in relation to OC

exposure under field and experimental conditions. Rauschenberger (2004) reported

alterations in embryonic alligator growth and development in relation to maternal OC

exposure. In addition, several reports have related OC exposure to modified alligator

thyroid histological parameters and regulation (Gunderson et al., 2002; Hewitt et al.,

2002; Crain et al., 1998). Wiebe et al., (2002) reported both hyperthyroid secretary

patterns of THs and subsequent accelerated growth among hatchlings from high OC

contaminated environments. In addition, controlled treatment studies utilizing OC

contaminants (i.e., total chlordane, total DDE, dieldrin, and toxaphene) demonstrated

altered growth in relation to OC contaminant exposure (Blanar et al., 2005; Bustnes et al.,

2005; Willingham, 2001; Lamai et al., 1999). These combined data suggest an inter-

relation between OC exposure and modification of growth and growth-regulating factors

such as thyroid hormones. However, the current study demonstrated no significant

differences in hatchling alligator thyroid regulation or growth rates in relation to in-ovo

OC exposure. Results of the current study may be attributed to the 1) inability to utilize









clutches with low viability, 2) additional growth-regulating products) other than or

integrated with THs that regulate hatchling alligator growth, or 3) non-OC contaminant

influences including: maternal size and/or habitat and nutritional quality.

Reduced alligator clutch viability has been reported within sites of intermediate to

high concentrations of OC contaminants in central Florida (Rauschenberger, 2004; Gross,

1994). These data demonstrate growth retardation and subsequent mortality during both

early and late embryonic development, and among hatchlings from high OC

environments. However, low viability clutches were excluded from the current study due

to clutch selection requirements: study clutches required at least fifteen hatchlings in

order to test the current studies hypothesis. The elimination of these clutches from the

current study likely removed hatchlings with an increased potential to demonstrate

irregularities in growth and developmental regulation in relation to OCs and/or additional

environmental stressors. Though differences in OC contaminant concentrations were

observed (p<.05), clutches utilized in the current study demonstrated no significant

differences in viability between sites.

Several authors have reported alterations in alligator thyroid regulation in relation

to OC contaminant exposure (Gunderson et al., 2002; Hewitt et al., 2002; Crain, 1998). In

addition, these reports stated that alligators from OC contaminated environments

demonstrated a general lack of correlation between plasma TH concentrations, sex and

body size. However, these studies were not able to eliminate several physiological and

environmental factors (i.e., age, sex, photoperiod, water and ambient temperatures and

food availability) reported to influence thyroid regulation. In addition, these studies

examined the relationship between OC contaminant exposure and hatchling growth as









well as plasma TH concentrations utilizing a single point in time sampling procedure. As

thyroid hormones have been reported to have a pulsatile secretary pattern, multiple

sampling over time would appear to be pertinent when trying to relate plasma TH

concentrations and growth to the hyper-variable influences of environmental contaminant

exposure. Wiebe et al. (2002) correlated both plasma T3 and T4 concentrations with

growth over time among hatchling alligators from sites of varying OC contamination

under captive conditions. To more directly examine the relationship between OC

exposure and alterations in hatchling growth and thyroid regulation, captive conditions

were designed to limit the influence of physiological and environmental influences on

thyroid regulation. These conditions included: 12L:12D photoperiod; constant ambient

and water temperatures; restricted pod size to avoid stressful overcrowding; documented

hatchling age; and ad-libitum food availability. Results from the 2002 study demonstrated

that hatchlings from high OC environments demonstrated a hyperthyroid TH secretary

pattern and accelerated growth. Utilizing the comparable captive conditions, a temporal

relationship between plasma TH concentrations and hatchling alligator growth was

observed over time in the current study. Additionally, thyroid and liver weights as well as

liver somatic indices were found to be representative biomarkers of hatchling growth

among and within sites over time. However, no relationship was observed between OC

exposure and hatchling alligator growth rates or plasma thyroid hormone concentrations

among or within sites over time. Therefore, future research may require examination of

additional growth regulating endocrine pathways when assessing the potential influence

of OC contaminant exposure on hatchling alligator growth regulation.









These conflicting data suggest that THs may not be the principal growth

regulating hormone influenced by OC contaminant exposure. McNabb (2000) noted that

THs act permissively or indirectly, in concert with the principal growth regulators:

growth hormone (GH) and insulin-like growth factor I (IGF-I). In addition, THs have

been reported to participate in highly integrated growth regulation among both

somatotropic and corticotropic axis' (Kuihn et al., 2005, Kobel et al., 2001, Elsey et al.,

1990).

Growth hormone (GH) is an essential regulator of growth with complex

metabolic functions (Bjoornsson et al., 2002). Pituitary GH secretion reported to be under

the dual control by two neuropeptides from the hypothalamus: GH releasing hormone

(GHRH) which stimulates GH release and somatostatin (SRIH) which has an inhibitory

action (Renaville et al., 2002). However, plasma GH concentrations have been

demonstrated to be influenced by a variety of hormones, growth factors, and

environmental influences (Fig. 2-20). The anabolic and growth-promoting effects of GH

are to a large extent mediated by the stimulation and expression of insulin-like growth

factor I (IGF-I) in the liver and peripheral tissues (Sjogren et al., 1999). The interactive

(i.e., local and systemic) regulation demonstrated between GH and IGF-I is known as the

"dual effector theory of action." (Bjiornsson et al., 2002). Several reports have examined

plasma IGF-I concentrations among reptilian models (Guillette et al., 1996, Crain et al.,

1995, Crain et al., 1995). These reports indicated that increased plasma IGF-I

concentrations were coincident with egg formation as oviparous species must

compartmentalize growth-promoting substances and nutrients into the yolk and albumen

of eggs (Guillette et al., 1996). In addition to IGF-I, maternal transfer of growth-









regulating substances (i.e., GH, TH) appears to be critical for embryo development with

implications on future hatchling growth and survival (Greenblatt et al., 1989). Therefore,

maternal quality, which encapsulates animal health in relation to exposure to

environmental stressors including OCs, continues to appear to be a dominant regulatory

factor in clutch growth and survival.

OC contaminant exposure has been reported to influence reproductive and

developmental parameters among adult and juvenile alligators (Rauschenberger et al.,

2004, Gross et al., 1994). These data suggest an integrated relationship between adult

alligator exposure to multiple environmental stressors (i.e., OCs, water quality,

nutritional quality) and subsequent alterations in clutch and hatchling quality. Under

captive conditions, the current study demonstrated that THs may serve as indicators of

hatchling alligator growth utilizing multiple sampling procedures over time. However, no

relationship was observed between OC exposure and hatchling alligator growth and

thyroid regulation in the current study. These data suggest that THs may not represent the

principal endocrine pathway affected by OC contaminant exposure. Therefore, the null

hypothesis which states that there will be no effects of in-ovo OC exposure on hatchling

alligator growth or thyroid regulation must be accepted. Future research efforts

examining the relationship between hatchling alligator growth and OC exposure should

incorporate an integrated evaluation of multiple endocrine pathways (i.e., GH, IGF-I, TH,

corticoids), utilize multiple sampling techniques over time, and, when possible, limit the

influence of reported physiological and environmental parameters on growth regulation

(Scollon et al., 2004).









In addition to OC exposure, anthropogenic habitat modifications have been

suggested as potential influential factors in the observed modifications in alligator

reproductive and growth parameters among OC contaminated sites. Schelske et al. (2005)

provides details a chronology of habitat modification in the upper Ocklawaha river basin

including: construction of the Beauclair canal and extensive levee systems, extensive

citrus and muck farming operations, as well as municipal sewage discharge. These habitat

modifications and the subsequent "back pumping" of phosphorous from muck farming

operations is credited with creating marginal habitat with an extensive changes of both

flora and fauna among this river system (Schelske et al., 2005). Several reports have

investigated the influences) of habitat modification and other non-OC related influences

on alligator clutch viability among OC contaminated sites. Rauschenberger (2004)

examined the incidence of alligator nutritional deficiencies specifically: thiamine

(Vitamin B1) deficiency, which has been suggested to increase embryonic mortality in

relation to OC contaminant exposure. Results from clutches collected from OC

contaminated environments demonstrated that thiamine deficiency may be involved in

decreased clutch viability. In addition, Mason (1995) suggested that changes in available

nesting vegetation had the potential to reduce alligator clutch viability through reduction

in insulation as well as inappropriate moisture content. Though alterations in alligator

reproductive and growth quality have been associated with OC contaminant exposure, the

tremendous influence of habitat modifications (i.e., habitat quality, water quality, non-

indigenous species) on alligator growth, reproduction and survival should not be

discounted.









Though no significant differences were observed in hatchling alligator growth or

thyroid regulation in relation to in-ovo OC exposure, these data do not discount the

potential for growth alterations among wild alligator populations in OC contaminated

environments. Significant variability in alligator reproductive and growth regulation

continues to be observed in relation to OC contaminated environments (Rauschenberger,

2004, Guillette et al., 1999, Gross et al., 1994). These data include: 1) gonadal

modifications such as: altered plasma sex steroid concentrations and histological

abnormalities, 2) increased fecundity 3) increased incidence of early and late embryonic

mortality, as well as 4) growth disparities between and within OC contaminated sites

versus control sites. In order to better relate the observed reductions in alligator

reproductive and clutch qualities to OC exposure, Rauschenberger et al. (2004) orally

dosed captive adult alligators in reproductive groups (1 male: female) with an eco-

relevant OC contaminant mixture. This experimental mixture was representative of OC

isoforms concentrations analytically identified among yolks from OC environments in

central Florida. Experimental clutches demonstrated comparable reductions in clutch

viability specifically: increased incidence of early embryonic mortality, which has been

observed in wild clutches from OC environments.

Data from the current study demonstrates that THs can be utilized as a

bioindicator of hatchling alligator growth under captive conditions. Therefore,

experimental control of established physiological and environmental influences on

thyroid regulation allowed for a more through examination of not only hatchling alligator

growth but, the potential inter-relation between OC contaminant exposure and subsequent

growth and thyroid regulation. These data represent a more direct examination of the









inter-relationship between OC exposure and altered hatchling alligator growth and

thyroid regulation. Though previous work reported a hyperthyroid secretary pattern and

accelerated growth in hatchling alligators from high OC environments, the current study

demonstrated no relationship between OC exposure and subsequent alterations in growth

or thyroid regulation(Table 2-8). These data suggest that hatchling alligator growth is

regulated by an integrated endocrine network (i.e., GH, IGF-I, corticoids) in which THs

may not be the principal regulatory agent. In addition, the inability to utilize clutches of

lower viability from OC contaminated sites may have restricted the incidence of

observing growth and developmental alterations.

Alligator reproductive and growth alterations continue to be reported in

association with OC contaminated sites. Previous data reported hyperthyroid secretary

patterns and accelerated hatchling alligator growth in association with high OC

contaminants. However, no relationship was observed between OC contaminant exposure

and hatchling growth or thyroid regulation in the current study. Though THs were

deemed useful for monitoring hatchling alligator growth, they do not appear to be the

principal growth regulatory factor. Future examination of both individual as well as

integrated regulatory relationships between growth-regulating hormones / growth factors

(i.e., GH, IGF-I, TH, corticoids) may prove more useful when trying to relate OC

contaminant exposure to observed alterations in alligator growth. In conclusion, there has

been a singular focus in associating the observed reductions in alligator reproductive and

growth parameters with OC contaminant exposure. However, the significant influences)

of environmental factors (i.e., habitat modification as well as water and nutritional






60


quality) should not be discounted when evaluating alligator physiology in relation to OC

contaminant exposure.


























Biosynthesis of Thyroid Hormowns
pical Follicle
Metmbrane
I [I


I-


Deiodinae


PTU, Tapazal

...'inbibit
Thyirid
Perondase
H20 as


Protease


o dotyrosifies


CI

I I

. :- -







I I -


Free roas Colloid
T4, T3 I Lysosom roplet


Figure 2-1.Graphical interpretation of thyroid hormone biosynthesis. Taken from
www.neurosci.pharm.otoledo.edu/MBC3320/thyroid.htm. (11/04/05).







62



















Fecundity Viability

60 100

50 a 80 a a

40
b 60
30F

20

10 20

0 0
GR OR EM AP GR OR EM AP


Figure 2-2. Clutch fecundity and clutch viability (site means). Significant differences
among sites were determined by Tukey Multiple Comparison Analysis (p
<.05).























Fecundity


GR OR EM AP


GR OR EM AP


Figure 2-3. Clutch fecundity and clutch viability (current study). Significant differences
among sites were determined by Tukey Multiple Comparison Analysis (p
<.05).


Viability





























YOLK OCP Concentrations
(Site Means)


10000



1000


Yolk OCP Concentrations
(Current Study)

m Total Chlordane lb
I DDTx
[ Dieldrin
S Toxaphene













OR GR AP EM


Figure 2-4.Yolk OC concentrations. site means (a) and current study (b). Significant
differences among sites were determined by Tukey Multiple Comparison
Analysis (p <.05).


10000



1000


STotal Chlordane a
SDDTx La
C Dieldrin
M Toxaphene b
a

b b

b a b a
b
aa I aa

h a
i_
























Total Length


Snout-Vent Length


600 300







I aI I I a a a b
S00 50
b bbb E a b bb




0 0
Sept Oct Nov Dec Jan Feb Mar Apr May Sept Oct Nov Dec Jan Feb Mar Apr May


Head


Weight


aa a
840
0 a bbbbb




20


Sept Oct Nov Dec Jan Feb Mar Apr May


600
500
400
S300
200
100


Sept Oct Nov Dec Jan Feb Mar Apr May


Figure 2-5.Hatchling alligator growth parameters among sites over time. Significant
differences among sites were determined by Tukey Multiple Comparison
Analysis(p < .05).








































Orange


OR 04-12
600 OR-04-B a
OR-04-3 ab
500 OR-04-W1

E 400 a babb
S300

200

100

0 Sa
Sept Oct Nov Dec Jan Feb Mar Apr May


Apopka


600

500

400

300

200

100


Sept Oct Nov Dec Jan Feb Mar Apr May


700 o-.-A o4 ----b-

600 ^HAP-04W10 a

E a a


6300 0
500




100





Sept Oct Nov Dec Jan Feb Mar Apr May


Figure 2-6.Hatchling alligator total length (mm) within sites over time. Significant

differences among sites were determined by Tukey Multiple Comparison

Analysis (p < .05).


Griffin


700

600

500

E 400
E
_ 300

200

100


Emerelda







67















Griffin Orange
350 350
300 GR4-o a b 300 o 18
250 a 250bb
ae bab a aId Ab








300 i350 -----
250b b 25300 a bb
200 a a a0 b b abb
150 b b a 00 0 b a b










Illll Illlll
100 100
50 50
0 _A 0





Sept Oct Nov Dec Jan Feb Mar Apr May Sept Oct Nov Dec Jan Feb Mar Apr May

EmereFigure 2-7.Hatchling alligator snout-vent length (mm) within sites over time. SignificantApopka
300 40 a 350 A01

250
200 a b b a I

100 10

50 50

Sept Oct Nov Dec Jan Feb Mar Apr May Sept Oct Nov Dec Jan Feb Mar Apr May

Figure 2-7.Hatchling alligator snout-vent length (mm) within sites over time. Significant
differences among sites were determined by Tukey Multiple Comparison
Analysis (p < .05).









68
























Griffin Griffin

100 o 100

80 GR-04-C 80 GR-04C b
S GR-04-D b b"

E 60 60 a
E. 11 iti E1 i- |
40 : 40

20 20

0 0
Sept Oct Nov Dec Jan Feb Mar Apr May Sept Oct Nov Dec Jan Feb Mar Apr May


Emerelda Apopka

8 EM-04 -01 a 1 AP-04-10
80 100



6O 60 b
E bb bb Eb

40

20
20


Sept Oct Nov Dec Jan Feb Mar Apr May Sept Oct Nov Dec Jan Feb Mar Apr May



Figure 2-8.Hatchling alligator head length (mm) within sites over time. Significant

differences among sites were determined by Tukey Multiple Comparison

Analysis (p < .05).









69

























Griffin Orange

800 700
800 GR-4- 700 OR4-12
700 GR4 a O 4-00
600R4-
50000
500 .. I,400
400 00






Sept Oct Nov Dec Jan Feb Mar Apr May Sept Oct Nov Dec Jan Feb Mar Apr May


Emerelda Apopka

600 700


40500
400
300 b
200200 b
bb I,
100 100 b b








-0 0

Sept Oct Nov Dec Jan Feb Mar Apr May Sept Oct Nov Dec Jan Feb Mar Apr May















Figure 2-9.Hatchling alligator body weight (g) within sites over time. Significant
differences among sites were determined by Tukey Multiple Comparison

Analysis (p < .05).
Analysis (p < .05).












































Snout-Vent Length


600

500

400 a

300 a a

200

100

0--
Sept Nov Jan March May



Head Length

Sn .


- OR
-- I
R O
E'E


Sept Nov Jan March May


b bR














Sept Nov Jan March May



Weight


G R
o O


Sept Nov


Jan March May


Thyroid Weight


Liver Weight


14

12

10
-
8 -
S-l
6

4

2

0-


Sept Nov Jan March May


S GR







a





Sept Nov


Jan March May


Figure 2-10. Hatchling alligator growth parameters (necropsy animals) among sites

over time. Significant differences among sites were determined by Tukey

Multiple Comparison Analysis (p < .05).


Total Length


60



40-



20 -


0.05


0.04


, 0.03


0.02



0.01
0.00


i a


0 -





































Griffin


- Sept
- No
- Jn
SMny


GR-0451 GR-04- G R-04-C GR-04-


Emerelda


700

600

500

400 b

300

200 b b

100


EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11


Orange


- Sept
- No
- Jn
SMny


OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5


Apopka


- Sept
- No
M Jy


AP-04-10 AP-04-W2 AP-04-W10


Figure 2-11. Hatchling alligator total length (mm)(necropsy animals) within sites over

time. Significant differences among sites were determined by Tukey Multiple

Comparison Analysis (p < .05).


b






























Griffin


Sa a a
b M b


b b




GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D

Eme reld a



May a o b


EM0401 EM0402 EM0403 EM0404 EM04


Orange


- Sept
-s t
I Jan
I Mar
I May
= .:t
M41


OR-04-1 OR-04-13 O04 R-04-W1 OR-04-W


Apopka


- Sept
I Jan
I Mar
I May


AP-04-10 AP-04-W2 AP-04-W10


Figure 2-12. Hatchling alligator snout-vent length (mm)(necropsy animals) within sites
over time. Significant differences among sites were determined by Tukey
Multiple Comparison Analysis (p < .05).









73
























Griffin Orange

100 100



M m m aa


0 0




OR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5


Emerelda Apopka




0 My ab b May



40 b i 40 b
20 20

0 0








EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 AP-04-10 AP-04-W2 AP-04-W10



Figure 2-13. Hatchling alligator head length (mm)(necropsy animals) within sites over



Comparison Analysis (p < .05).
Comparison Analysis (p < .05).










74


























Griffin Orange

800 800
700 700

Emerelda Apopka
500 b 500
400 00

300 n300
200 200 b b b
100 100
0 0
GR-04-51 GR-04-A GR-04-B GR-04-C R-04- OR-04-12 OR-04-13 OR-04-B OR-04-W 1 OR-04-W5


Emerelda Apopka

700 700










SSptime. Significant differences among sites were determined by Tukey Multiple
600 ~,Anl 600p < .

500 5 500
400 b 400

300 300

200 b 200
b b
b b
100 L100
b jbb b b Of lb -b b

EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 0AP-04-10 AP-04-W2 AP-04-WI0



Figure 2-14. Hatchling alligator body weight (g) (necropsy animals) within sites over

time. Significant differences among sites were determined by Tukey Multiple

Comparison Analysis (p < .05).





































Griffin Orange

0.07 0.05

0.05
S Mar 0.04 m M




S0.03
0.05 2



0.01
0b 0.







0.00 0.00
GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D OR-04-12 OR-04-13 OR-04-B OR-04-WI OR-04-W5


Emerelda Apopka

0.06 0.06

0.05 a 0.05 a

S0.04 b 0.04

0.03 b 0.03 b




0.01 .b0.01

0.00 0.00
EM 04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11 AP-04-10 AP-04-W2 AP-04-WI0



Figure 2-15. Hatchling alligator thyroid weight (g)(necropsy animals) within sites over

time. Significant differences among sites were determined by Tukey Multiple

Comparison Analysis (p < .05).









76

























Griffin Orange

18 18
16 16
14 14






4 4
2 2
0 0
EM-04-01 GR-04-A GR-04-B GR-04-C GR-04-D OR-04-12 OR-04-13 OR-04-B OR- 04-WI OR-04-W5

Emerelda Apopka

1 time. Significant 14differences among sites were determined by Tukey Multiple


12 a ( <
' 10
8 b 6
.j 6
b b
4 4 b
2 2
2a b bbb b2
0 0
EM 04-01 EM.-04-02 E M04-03 EM.404 EM.-04-.11 AP-04-10 AP-04-W2 AP-04-W1 0



Figure 2-16. Hatchling alligator liver weight (g) (necropsy animals) within sites over
time. Significant differences among sites were determined by Tukey Multiple

Comparison Analysis (p < .05).



























Total Thyroxine


Free Thyroxine


Oct Nov Dec Jan Feb Mar Apr May


Oct Nov Dec Jan Feb Mar Apr May


Figure 2-17. Hatchling alligator total thyroxine(ng/ml)and free thyroxine (pg/ml)
plasma concentrations among sites over time. Significant differences among
sites were determined by Tukey Multiple Comparison Analysis (p < .05).





































Orange


Oct Nov Dec Jan Feb Mar Apr May


Emerelda


S EM-04-01
m EM-04-02












Oct Nov Dec Jan Feb


OR-04-B
OR-04-W1

a N








Oct Nov Dec Jan Feb Mar Apr May


Mar Apr May


Apopka


Figure 2-18. Hatchling alligator total thyroxine (ng/ml) plasma concentrations within

sites over time. Significant differences among sites were determined by Tukey

Multiple Comparison Analysis (p < .05).


Oct Nov Dec Jan Feb Mar Apr May


Griffin








79






















Griffin Orange

7 8













Emerelda Apopka





IL2 IM aPa41
G aR a b.A -04-2
D. 0
















Oct Nov Dec Jan Feb Mar Apr May Oct Nov Dec Jan Feb Mar Apr



Figure 2-19. Hatchling alligator free thyroxine (pg/ml) plasma concentrations within
sites over time. Significant differences among sites were determined by Tukey

Multiple Comparison Analysis (p < .05).
Ec o e a e a p a c o e a e a p

Fiur E-9 5acln liao rety oie(g m)pa m o cnrtoswti
site ovr tme.Sigifiantdiferecesamog steswer deermnedby uke
Mutil 3oprsnAayss( 0)


lay


lay























Somatostatin-28 GH Estradiol
IGF-I
Somatostatin-25 SRIF TRH

GnRH
5-Hydroxytryptamine SRIF >
Norepinephrine Estradiol GLP-I
Dopamine I /

DOPA SRIF
Neuropeptide Y

adiol Somatotroph cells
NMA of anterior pituitary Estradiol
tosterone Norepinephrine


T4 Galanin
T3

TSH Ration Size Exercise
Protein Intake Ovulation
Starvation Temperature Bombesin
Acute Stress Daylength
Chronic Stress Seawater Adaptation CCK GHRH



Growth Hormone Release





Figure 2-20.Graphical interpretation of factors that control the release of growth
hormone. Adapted from Mommsen, 1998.











Table 2-1. Total length growth rates among and within sites.

Apopka Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
AP-04-10 0.92719 1.2397 0.8521 0.8392 1.061 1.0336 1.0406 0.9725 1.0382 1.0005
AP-04-W2 0.65300 0.8118 0.9669 0.9649 0.929 0.9695 0.9993 1.0239 1.0647 0.9314
AP-04-W10 0.87014 1.2942 1.3956 1.3069 1.2078 1.1868 0.9884 1.1727 1.2023 1.1805
MEAN 0.81678 1.1152 1.0715 1.037 1.0659 1.0633 1.0094 1.0564 1.1017 1.0375


Emeralda Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
EM-04-01 0.77604 1.1813 1.3056 1.2425 1.1461 1.0249 1.1103 1.0215 1.0639 1.0969
EM-04-02 1.35882 1.4240 1.5496 1.5557 1.4495 1.3829 1.3603 1.1997 1.2105 1.3879
EM-04-03 0.60000 0.8813 0.9617 0.8891 0.9475 0.9306 0.9593 0.9855 1.0178 0.9081
EM-04-04 0.71874 0.8774 1.0094 1.0137 0.9838 1.0014 0.981 0.9953 1.0308 0.9568
EM-04-11 1.30798 1.2455 1.2505 1.1403 1.062 1.0473 1.024 1.0319 1.0678 1.1308
MEAN 0.95232 1.1219 1.2154 1.1683 1.1178 1.0774 1.087 1.0468 1.0781 1.0961


Griffin Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
GR-04-51 1.05110 1.1677 1.3124 1.2542 1.1867 1.183 1.1959 1.1849 1.195 1.1923
GR-04-A 1.17706 1.1154 1.2462 1.1698 1.1075 1.0523 1.016 0.9912 1.0139 1.0988
GR-04-B 1.25966 1.4946 1.5755 1.4899 1.3532 1.4082 1.3217 1.2943 1.3036 1.389
GR-04-C 1.01199 1.1343 1.3977 1.362 1.2779 1.2944 1.2933 1.2191 1.2655 1.2507
GR-04-D 0.95602 1.3219 1.4598 1.3821 1.2988 1.3109 1.147 1.3192 1.3961 1.288
MEAN 1.09117 1.2468 1.3983 1.3316 1.2448 1.2498 1.1948 1.2017 1.2348 1.2438


Orange Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
OR-04-12 0.78744 1.0712 1.165 1.105 1.0405 0.999 0.9785 0.985 1.026 1.0175
OR-04-13 0.51497 0.8102 1.0353 1.0101 1.0077 1.0107 1.0663 0.9947 1.009 0.9399
OR-04-B 2.94167 1.4562 1.4706 1.4429 1.346 1.3432 1.3341 1.2683 1.2874 1.5434
OR-04-W1 0.77147 1.0547 1.1855 1.1606 1.1091 1.1101 1.062 1.0742 1.1135 1.0713
OR-04-W5 1.20455 1.3867 1.4210 1.3716 1.2823 1.3337 1.3012 1.2719 1.2736 1.3163
MEAN 1.24402 1.1558 1.2555 1.218 1.1571 1.1593 1.1484 1.1188 1.1419 1.1777











Table 2-2. Snout-vent length growth rates among and within sites.

Apopka Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
AP-04-10 0.41725 0.5400 0.4251 0.4391 0.513 0.5115 0.5123 0.4711 0.4985 0.4809
AP-04-W2 0.29293 0.3679 0.4742 0.4943 0.4633 0.4845 0.4915 0.5021 0.5209 0.4546
AP-04-W10 0.35694 0.6228 0.6875 0.5936 0.594 0.5738 0.5838 0.562 0.579 0.5726
MEAN 0.35571 0.5102 0.529 0.509 0.5234 0.5233 0.5292 0.5117 0.5328 0.5027


Emeralda Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
EM-04-01 0.35677 0.5489 0.6454 0.612 0.5553 0.5518 0.5424 0.4982 0.5136 0.536
EM-04-02 0.0239 0.4779 0.6486 0.6905 0.645 0.6256 0.6075 0.5411 0.5434 0.5337
EM-04-03 0.30013 0.4205 0.4855 0.4862 0.4772 0.4622 0.4723 0.4821 0.4935 0.4533
EM-04-04 0.31928 0.4267 0.5154 0.525 0.4989 0.5042 0.4818 0.4978 0.5114 0.4756
EM-04-11 0.70714 0.6343 0.6458 0.6089 0.543 0.5394 0.5204 0.5207 0.5309 0.5834
MEAN 0.34145 0.5017 0.5881 0.5845 0.5439 0.5366 0.5249 0.508 0.5185 0.5164


Griffin Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
GR-04-51 0.45901 0.5561 0.6835 0.6408 0.592 0.5932 0.5952 0.5848 0.5841 0.5876
GR-04-A 0.49338 0.5452 0.6354 0.5891 0.5478 0.5264 0.6536 0.4852 0.4899 0.5518
GR-04-B 0.51534 0.684 0.7696 0.7288 0.6537 0.6817 0.6582 0.6192 0.6277 0.6598
GR-04-C 0.41335 0.5216 0.6842 0.6815 0.6404 0.6308 0.6271 0.5895 0.6091 0.5997
GR-04-D 0.36947 0.5993 0.7288 0.6933 0.6338 0.6469 0.7693 0.6421 0.6752 0.6398
MEAN 0.45011 0.5813 0.7003 0.6667 0.6135 0.6158 0.6607 0.5842 0.5972 0.6078


Orange Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
OR-04-12 0.31765 0.4799 0.5794 0.5471 0.5018 0.4862 0.4729 0.4669 0.4829 0.4816
OR-04-13 0.31012 0.3886 0.5214 0.5114 0.4996 0.4954 0.5174 0.4926 0.4879 0.4694
OR-04-B 0.52337 0.6927 0.7193 0.7192 0.6548 0.658 0.6483 0.6109 0.6211 0.6497
OR-04-W1 0.36559 0.5127 0.5961 0.5916 0.5435 0.5632 0.5356 0.5445 0.5561 0.5343
OR-04-W5 0.51056 0.6264 0.6669 0.6606 0.6066 0.6115 0.6203 0.5984 0.5876 0.6099
MEAN 0.40546 0.54 0.6166 0.606 0.5613 0.5629 0.5589 0.5427 0.5471 0.549











Table 2-3. Head length growth rates among and within sites.


Apopka
AP-04-10
AP-04-W2
AP-04-W10
MEAN


Emeralda
EM-04-01
EM-04-02
EM-04-03
EM-04-04
EM-04-11
MEAN


Griffin
GR-04-51
GR-04-A
GR-04-B
GR-04-C
GR-04-D
MEAN


Orange
OR-04-12
OR-04-13
OR-04-B
OR-04-W1
OR-04-W5
MEAN


Sept Oct Nov
0.10576 0.1288 0.0943
0.08664 0.0973 0.1133
0.12184 0.1536 0.1593
0.10475 0.1266 0.1223


Sept Oct Nov
0.11993 0.1372 0.1445
0.11278 0.1434 0.1605
0.07988 0.11 0.1101
0.15792 0.1075 0.118
0.1484 0.1422 0.1386
0.12378 0.1281 0.1344


Sept Oct Nov
0.13007 0.1454 0.1532
0.05335 0.1392 0.1373
0.12613 0.159 0.1683
0.12601 0.138 0.1556
0.08718 0.1407 0.1511
0.10455 0.1445 0.1531


Sept Oct Nov
0.10978 0.1278 0.1300
0.06736 0.0980 0.1125
0.12861 0.1578 0.1572
0.03139 0.0911 0.1105
0.14353 0.1569 0.1543
0.09613 0.1263 0.1329


Dec
0.0990
0.1126
0.1474
0.1197


Dec
0.1366
0.1619
0.1104
0.1175
0.1300
0.1313


Dec
0.1466
0.1371
0.1600
0.1511
0.1448
0.1479


Dec
0.1223
0.1144
0.1567
0.1136
0.1482
0.1311


Jan Feb Mar Apr May MEAN
0.1227 0.1183 0.1178 0.11 0.1175 0.1127
0.1096 0.1122 0.1137 0.116 0.1232 0.1094
0.1394 0.1374 0.1388 0.134 0.1385 0.1411
0.1239 0.1227 0.1235 0.12 0.1264 0.1211


Jan Feb Mar Apr May MEAN
0.1283 0.1237 0.1216 0.1127 0.1196 0.1271
0.1549 0.1477 0.1471 0.1319 0.1348 0.1439
0.1112 0.108 0.1109 0.1115 0.1172 0.1077
0.1152 0.1158 0.1142 0.1167 0.1202 0.1203
0.1212 0.1175 0.1167 0.117 0.1219 0.1282
0.1262 0.1225 0.1221 0.118 0.1228 0.1254


Jan Feb Mar Apr May MEAN
0.1412 0.1375 0.1373 0.1374 0.1397 0.1409
0.1288 0.1223 0.1181 0.1149 0.1178 0.1188
0.1502 0.1551 0.1523 0.1439 0.1459 0.1512
0.1448 0.1325 0.1445 0.1368 0.1458 0.1417
0.1402 0.1392 0.1375 0.1404 0.1501 0.1368
0.141 0.1373 0.138 0.1347 0.1399 0.1379


Jan Feb Mar Apr May MEAN
0.1168 0.1126 0.1095 0.1085 0.1139 0.1168
0.1125 0.1105 0.1169 0.1099 0.1132 0.1061
0.1499 0.1492 0.1472 0.14 0.1441 0.1479
0.1119 0.116 0.1141 0.1154 0.1215 0.1028
0.1409 0.1396 0.1398 0.1388 0.142 0.1449
0.1264 0.1256 0.1255 0.1225 0.1269 0.1237











Table 2-4. Body weight growth rates among and within sites.

Apopka Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
AP-04-10 -0.142 0.3339 0.4028 0.5629 0.664 0.8235 0.9383 0.9884 1.2239 0.6439
AP-04-W2 -0.186 0.2751 0.5495 0.7342 0.6933 0.9298 1.1494 1.2807 1.483 0.7677
AP-04-W10 -0.192 0.4847 0.8331 1.0648 0.9643 1.157 1.4169 1.5286 1.7323 0.9989
MEAN -0.173 0.3646 0.5951 0.7873 0.7739 0.9701 1.1682 1.2659 1.4797 0.8035


Emeralda Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
EM-04-01 -0.096 0.4238 0.6937 0.8715 0.814 0.9782 1.1183 1.1024 1.3155 0.8024
EM-04-02 -0.053 0.565 0.9901 1.3877 1.2348 1.4202 1.7494 1.4623 1.6692 1.1584
EM-04-03 -0.060 0.2368 0.3948 0.5283 0.5246 0.6456 0.8348 0.9462 1.1025 0.5725
EM-04-04 -0.244 0.2415 0.5107 0.7018 0.6939 0.8735 1.0231 1.1602 1.3324 0.6992
EM-04-11 -0.125 0.3893 0.6335 0.7705 0.7107 0.8668 0.9828 1.1521 1.2867 0.7408
MEAN -0.116 0.3713 0.6445 0.8519 0.7956 0.9569 1.1417 1.1646 1.3412 0.7947


Griffin Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
GR-04-51 -0.085 0.4041 0.7652 1.0108 0.9695 1.2064 1.3834 1.5226 1.5575 0.9705
GR-04-A -0.196 0.3000 0.6267 0.8584 0.7982 0.9064 1.0218 1.0647 1.2059 0.7318
GR-04-B -0.048 0.5038 0.8912 1.1153 1.0485 1.3593 1.5635 1.6191 1.8298 1.0981
GR-04-C -0.097 0.4397 0.8715 1.1457 1.0766 1.3123 1.5465 1.5398 1.8769 1.0791
GR-04-D -0.034 0.5374 0.8623 1.1058 1.0394 1.2682 1.5675 1.7796 2.18 1.1451
MEAN -0.092 0.4370 0.8034 1.0472 0.9864 1.2106 1.4165 1.5051 1.73 1.0049


Orange Sept Oct Nov Dec Jan Feb Mar Apr May MEAN
OR-04-12 -0.118 0.3779 0.6099 0.7557 0.7193 0.8099 0.8451 0.9966 1.1556 0.6836
OR-04-13 -0.102 0.2698 0.5399 0.6947 0.6913 0.8461 1.0955 1.0616 1.1855 0.698
OR-04-B -0.113 0.5184 0.8935 1.1375 1.0759 1.3295 1.6161 1.6289 1.8493 1.104
OR-04-W1 -0.127 0.2920 0.5567 0.7012 0.6772 0.8885 1.0224 1.1775 1.348 0.7263
OR-04-W5 0.008 0.4602 0.7101 0.9814 0.9218 1.1863 1.4461 1.5617 1.7317 1.0008
MEAN -0.09 0.3837 0.662 0.8541 0.8171 1.0121 1.2051 1.2853 1.454 0.8426






85


Table 2-5. Hatchling alligator thyroid (TSI) and liver (LSI) somatic indices among sites
over time. No differences were observed in TSI among sites. Temporal
differences were observed in LSI among sites. Significant differences
determined utilizing Wilkoxon analysis with the Kruskal -Wallis Test
(p < .05).
TSI among Sites LSI among Sites

Date Chi Square Pr> Chi Square Date Chi Square Pr > Chi Square
Sept 4.2855 0.2322 Sept 6.8526 0.0767
Nov 5.936 0.1148 Nov 17.0271 0.0007
Jan 1.1091 0.7749 Jan 6.1687 0.1037
Mar 4.8678 0.1817 Mar 9.786 0.0205
May 1.0038 0.8003 May 6.3759 0.0947









Table 2-6. Hatchling alligator thyroid somatic indices (TSI) within sites over time. No
significant differences were observed. Significant differences determined
utilizing Wilkoxon analysis with the Kruskal -Wallis Test (p < .05).
Apopka Chi-Square Pr> Chi Square
Sept 3.4667 0.1767
Nov 3.2000 0.2019
Jan 0.3556 0.8371
Mar 5.0667 0.0794
May 5.0667 0.0794
Emeralda Chi-Square Pr> Chi Square
Sept 5.1434 0.2729
Nov 10.8945 0.0278
Jan 9.5667 0.0484
Mar 5.5667 0.2339
May 4.9333 0.2942
Griffin Chi-Square Pr> Chi Square
Sept 9.1747 0.0569
Nov 6.8706 0.1429
Jan 4.7667 0.3121
Mar 7.2667 0.1224
May 9.7333 0.0452
Orange Chi-Square Pr> Chi Square
Sept 4.2000 0.3796
Nov 8.9667 0.0619
Jan 10.7667 0.0293
Mar 2.5796 0.6304
May 7.0000 0.1359









Table 2-7. Hatchling alligator liver somatic indices (LSI) within sites over time. No
significant differences were observed. Significant differences determined
utilizing Wilkoxon analysis with the Kruskal -Wallis Test (p < .05).
Apopka Chi-Square Pr> Chi Square
Sept 3.2889 0.1931
Nov 3.2000 0.2019
Jan 4.6222 0.0992
Mar 0.8000 0.6703
May 5.9556 0.0509
Emeralda Chi-Square Pr> Chi Square
Sept 10.4333 0.0337
Nov 4.4667 0.3465
Jan 4.6333 0.3270
Mar 10.833 0.0285
May 3.9000 0.4197
Griffin Chi-Square Pr> Chi Square
Sept 12.0333 0.0171
Nov 8.1667 0.0857
Jan 11.7000 0.0197
Mar 7.9667 0.0928
May 3.7000 0.4481
Orange Chi-Square Pr> Chi Square
Sept 7.1711 0.1271
Nov 9.2333 0.0555
Jan 5.3000 0.2579
Mar 2.5000 0.6446
May 2.7667 0.5976









Table 2-8. Multiple linear regression analysis of hatchling alligator growth rates, thyroid
hormone secretary rates and organochlorine contaminant concentrations. No
significant relationships were demonstrated (p < .05).
Total Length Snout-Vent Head Length Body Weight
Rate Length Rate Rate Rate

Total Chlordane 0.1084 0.1108 0.3362 0.3072


Total DDTx 0.1129 0.0959 0.3938 0.1462


Dieldrin 0.1281 0.1246 0.3376 0.4492


Toxaphene 0.4905 0.6954 0.8230 0.5753


TT4 Rate 0.3704 0.7254 0.5308 0.8545


FT4 Rate 0.2137


0.1193 0.4314


0.1983