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.

iv

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

v

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.

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 Relation 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 Hormone Utilization and Regulation.......................13

Fish ......................................................................................................................13 Amphibians..........................................................................................................13 Avian ...................................................................................................................14

Physiological and Environmental Influences 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 Contaminant Exposure on Thyroid Regulation ...............20 Overview .............................................................................................................20 Effects of Organochlorine Contaminant Exposure on Alligator Thyroid

Regulation........................................................................................................21 Thyroid Histology Alterations in Relation to Organochlorine Contaminant

Exposure ..........................................................................................................23 Influence of Organochlorine Contaminant Exposure on Integrated Levels of

Thyroid Hormone Regulation ..........................................................................25 Thyroid Hormone Synthesis................................................................................25

vii

Thyroid Hormone Binding Proteins ....................................................................26 Deiodination of Thyroid Hormones ....................................................................27 Thyroid Hormone Excretion................................................................................28

Growth in Relation to p,p’-DDE, dieldrin, 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 Procedures (Total and Free Thyroxine) ..45 Free T4 (FT4) Assay Procedures.........................................................................46 Total T4 (TT4) Assay Procedures .......................................................................46 Analysis of Chlorinated Analytes 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

viii

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

ix

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 time……...................................................................................................................70

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 time……...................................................................................................................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 among sites over time.. ....................................................................77

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

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

xi

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 secretory 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

xii

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.

1

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., (≈ 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

2

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

3

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.

4

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

5

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

6

(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

7

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.

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+ 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

9

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,

10

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 (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 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

11

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 mechanism(s) that

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

demonstrate species-differences in TH utilization and regulation.

12

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 (O. 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

13

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

14

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

15

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 (Agelaius phoeniceus) 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

16

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 (Anas platyrhynchos) rather than LH

concentrations suggesting that plasma T4 concentrations may be responsible for a

17

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 (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 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.,

18

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

19

(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 (30ºC ambient temperature, 27± 1ºC water temperature and constant

light) or variable (40ºC to 24 ºC ambient temperature, 19ºC to 24 ºC 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

20

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

21

(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

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 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

23

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 (O. kisutch) and chinook ( O. 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

24

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).

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 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

26

(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 a para

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

27

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 (0.1µg

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

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 (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 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 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

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 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

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. (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,

31

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

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 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

33

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

µg/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 µg/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

34

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 secretory 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

35

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 secretory 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 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 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 secretory 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

secretory 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

37

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

38

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

39

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

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 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

41

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

42

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 secretory 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 secretory

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 secretory

pattern that accelerates hatchling 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 (airboat) 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

43

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 31.5º C ± 1º 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

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, 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

45

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 µl) 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 (125I) 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 µl volumes for Free-T4 and in 5, 10, 15, 20

46

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 (50µL 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 ± 1º 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 (25µL 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 25º C) for a period of 60 minutes.

47

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, α-BHC, β-BHC, lindane, δ-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 Büchner 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

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. 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 µL 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 µL injection and separated across the HP-5MS column

(30 m x 0.25 mm; 0.25 µm film thickness; J & W Scientific, Inc., Folsom, CA) under a

temperature program that began at 60º C, increased at 10º C/min to 270º C, was held for

5 min, then increased at 25º C/min to 300º 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 µL of US-108 (120

µg/mL; Ultra Scientific), added just prior to analysis. All samples also contained a

49

surrogate, 2 µg/mL of tetrachloroxylene (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 µg/mL of γ-BHC, heptachlor, aldrin, dieldrin, endrin,

and p,p’-DDT) with an acceptable recovery ranging from 70 – 130%. Repeated analyses

were conducted as allowed by matrix interferences 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

50

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

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 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 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 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

52

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 secretory 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 secretory

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

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 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

54

well as plasma TH concentrations utilizing a single 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 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 secretory

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.

55

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’ (Kühn et al., 2005, Kobel et al., 2001, Elsey et al.,

1990).

Growth hormone (GH) is an essential regulator of growth with complex

metabolic functions (Bjöornsson 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 (Sjögren 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.” (Bjöornsson 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-

56

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).

57

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 influence(s) 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.

58

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:1 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

59

inter-relationship between OC exposure and altered hatchling alligator growth and

thyroid regulation. Though previous work reported a hyperthyroid secretory 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 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 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 influence(s)

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.

61

Figure 2-1.Graphical interpretation of thyroid hormone biosynthesis. Taken from

www.neurosci.pharm.otoledo.edu/MBC3320/thyroid.htm. (11/04/05).

62

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

b

Fecundity

GR OR EM AP0

10

20

30

40

50

60

a

b a a

63

Figure 2-3. Clutch fecundity and clutch viability (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

aa a

a

Fecundity

GR OR EM AP0

10

20

30

40

50

60

a

b a a

64

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).

YOLK OCP Concentrations (Site Means)

OR GR AP EM

(ng

/ g)

1

10

100

1000

10000Total Chlordane DDTx Dieldrin Toxaphene

a

ab

ab

b

a

b

b

b

aa

a

a

a

b

b

Yolk OCP Concentrations (Current Study)

OR GR AP EM

(ng

/ g)

1

10

100

1000

10000Total Chlordane DDTx Dieldrin Toxaphene

a b

65

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

differences among sites were determined by Tukey Multiple Comparison Analysis(p < .05).

Total Length

S ept O ct N ov D ec Jan Feb M ar A pr M ay

TL (m

m)

0

100

200

300

400

500

600G R O RE MAP

a a aaa

ab bb

a b b b

a ab b b

a ab b b

a ab b b

a ab a

b b

a ab b b

a ab

abb

Head

Sept O ct Nov Dec Jan Feb M ar Apr M ay

Hea

d (m

m)

0

20

40

60

80G RO RE MAP

aa a a

a bb ba b bb

a bb b

a bb b

a ab

abb

a ab b b

a ab b

ab

a ab b

ab

W eight

S ept O ct N ov D ec Jan Feb M ar A pr M ay

Wei

ght (

g)

0

100

200

300

400

500

600G R O RE MAP

a bb ba bb

ab

a b b b

abb b

ab b b

ab bb

aab

ab b

aab

ab

b

aab

b

ab

S n o u t-V en t L en g th

S ep t O ct N o v D e c Ja n F e b M ar A p r M a y

SVL

(mm

)

0

5 0

1 00

1 50

2 00

2 50

3 00G RO RE MAP

aab

ab b

a b bb

a b bb

abb b

abbb

a ab

ab b

a

bb b

a ab bb

a ab b

ab

66

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

Sept Oct Nov Dec Jan Feb Mar Apr May

TL (m

m)

0

100

200

300

400

500

600

700GR-04-51 GR-04-A GR-04-B GR-04-C GR-04-D

aab

bc c

bab a a

ab

bab a a

ab b

ab a a

ab a

aa a abc c

aab

abc

aaa

aa

ab

b

ab

ab a

ab

b

aaa

b

Orange

Sept Oct Nov Dec Jan Feb Mar Apr May

TL (m

m)

0

100

200

300

400

500

600

700OR-04-12 OR-04-13 OR-04-B OR-04-W1 OR-04-W5

a a

a

aa

bc b ba

ac

ab

abab b

b b ba

ab

cd

bc

a bd

ab c

bc

ab

abc

ac

abc

a bc

ab c

a ab

aba

bb

aba

ab

ab

Emerelda

Sept Oct Nov Dec Jan Feb Mar Apr May

TL (m

m)

0

100

200

300

400

500

600EM-04-01 EM-04-02 EM-04-03 EM-04-04 EM-04-11

a ab bc

a a a ab

ab a

cb

ab

ab

a

cbb

ba

c

bc

bc

ab

a

b

ab

ab

ab

a

bb

baa

aa

aa a

a a

a

Apopka

Sept Oct Nov Dec Jan Feb Mar Apr May

TL (m

m)

0

100

200

300

400

500

600

700AP-04-10 AP-04-W2 AP-04-W10

a baa a

b b a bab

a bab

a b

ab a

a

aa

a

a aa

aa

c

67

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).

O range

Sept O ct Nov Dec Jan Feb M ar A pr M ay

SVL

(mm

)

0

50

100

150

200

250

300

350O R-04-12O R-04-13O R-04-BO R-04-W 1O R-04-W 5

aab

bc c

bc

abb b b

ab bab

ab

ab

abb b

ab b b

ab

aabc

ab

bc

a ab

c

bc

bc

c

a

b

aba

bab

a

b b

ab

ab

Griffin

Sept Oct Nov Dec Jan Feb Mar Apr May

SVL

(mm

)

0

50

100

150

200

250

300

350GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D

aab

bc c c

aa aa aa

aa a aa

aaa aab b

ab

aba

aab

baa

b

aa

a

a

a ab

b

aba

ba

abc

abc

c

ab

a

Em erelda

Sept O ct Nov D ec Jan Feb M ar A pr M ay

SVL

(mm

)

0

50

100

150

200

250

300EM -04-01EM -04-02EM -04-03EM -04-04EM -04-11

ab bc d

aa a ab

ab a

cb b

abb b

c

ba

cb

bc

abc

a

c

ab

bc

ab

a

bab

b

a

aaaa

a aa

aa

Apopka

Sept O ct N ov D ec Jan Feb M ar A pr M ay

SVL

(mm

)

0

50

100

150

200

250

300

350AP-04-10AP-04-W 2AP-04-W 10

a bc

aab

a ab

aa a

aabb

aaa a

aaa

a

a aa

a

68

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).

Griffin

Sept Oct Nov Dec Jan Feb Mar Apr May

Hea

d (m

m)

0

20

40

60

80

100GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D

a aa aaa a aa a

aa a aaaaa a a

a a a a aaa

aa a

ab b

aaab

a a aa

a

ab

b

ab a a

Griffin

Sept Oct Nov Dec Jan Feb Mar Apr May

Hea

d (m

m)

0

20

40

60

80

100GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D

a aa aaa a aa a

aa a aaaaa a a

a a a a aaa

aa a

ab b

aaab

a a aa

a

ab

b

ab a a

Em erelda

Sept Oct Nov Dec Jan Feb Mar Apr May

Hea

d (m

m)

0

20

40

60

80EM -04-01EM -04-02EM -04-03EM -04-04EM -04-11

aabb bb

aab

ab b

c

ab a

cb b

ba

cb b

ba

cb

bc

bc

a

c

ab

bc b

a

babb

aa

a

a aa a

aa

a

Apopka

Sept Oct Nov Dec Jan Feb Mar Apr May

Hea

d (m

m)

0

20

40

60

80

100AP-04-10AP-04-W 2AP-04-W 10

a bca ab

a aba a

ba

abb

aab

ba a

aa a a a

a a

69

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).

Griffin

Sept Oct Nov Dec Jan Feb M ar Apr M ay

Wei

ght (

g)

0

100

200

300

400

500

600

700

800GR-04-51GR-04-AGR-04-BGR-04-CGR-04-D

a a bb ca a aa a

ab b

ab a a

b

abb

abaa

b

ab b

aba

ab

ab

aa

b

ab

ab

b

aaa

ab

b

ab

ab

aab

b

ab

a

a

Orange

Sept Oct Nov Dec Jan Feb Mar Apr May

Wei

ght (

g)

0

100

200

300

400

500

600

700OR-04-12OR-04-13OR-04-BOR-04-W1OR-04-W5

ab b bba bb bb

ab b bb

a ab

cbc

bc

a abb

cbc c

a abb

c bcc

a ab

bc

bc

c

a

ac

c

ab

bac

a

ab

ab

bb

E m erelda

Sept O ct N ov D ec Jan Feb M ar A pr M ay

Wei

ght (

g)

0

100

200

300

400

500

600EM -04-01EM -04-02EM -04-03EM -04-04EM -04-11

aabbc d

a b bbc

ba

cb b

b

a

c

b bb

a

c

b bc

a

b b

b

b

a

b bb

b

ab

a

b

aba

b

a

a

a

a

a

Apopka

Sept O ct N ov D ec Jan Feb M ar A pr M ay

Wei

ght (

g)

0

100

200

300

400

500

600

700AP-04-10AP-04-W 2AP-04-W 10

a bca ab

a ab

aab

b

aab

b

aab

baab

b

a a

a

a

aa

70

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).

Thyroid W eight

Sept N ov Jan M arch M ay

Thyr

oid

(g)

0 .00

0.01

0.02

0.03

0.04

0.05G RO REMAP

a

ab

bab

a a a a

a a a a

a

a

a

a

a

a a a

L iver W eight

Sept N ov Jan M arch M ay

Live

r (g)

0

2

4

6

8

10

12

14GROREMAP

a a a a

aa a a

aa a a

aa a a

a

b

ab a

b

W e ig h t

S e p t N o v J a n M a r c h M a y

Wei

ght (

g)

0

1 0 0

2 0 0

3 0 0

4 0 0

5 0 0

6 0 0

7 0 0G RO RE MA P

a a a a

a abb b

a aa a

aa a

a

a

b

ab

ab

Snout-V ent Length

Sept N ov Jan M arch M ay

SVL

(mm

)

0

50

100

150

200

250

300

350G RO REMAP

a a a a

a bb b

a aa a

a a aa

a ab

ab

b

H ead Length

Sept N ov Jan M arch M ay

Hea

d (m

m)

0

20

40

60

80G RO REMAP

a aa a

a a a a

a aa a

a a a a

a ab

ab

b

Total Length

Sept Nov Jan M arch M ay

TL (m

m)

0

100

200

300

400

500

600

700G RO REMAP

a bab

ab

a ab b b

a a aa

a aaa

ab

ab

ab

71

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).

O range

O R -04-12 O R -04-13 O R -04-B O R -04-W 1 O R -04-W 5

TL (m

m)

0

100

200

300

400

500

600

700S eptN ovJanM arM ay

a

a

ab

b

b

a

a

ab

a

ab

a

a

a

a

a

a

a

bab

b

a

a

ab

a

ab

E m ere lda

E M -04-01 E M -04-02 E M -04-03 E M -04-04 E M -04-11

TL (m

m)

0

100

200

300

400

500

600

700S eptN o vJanM arM ay

ba a a a

aab

ab a

bb

a

bb

bb

a

ab

bc

bc

c

a

ab

ab

ab b

A popka

A P -04-10 A P -04-W 2 A P -04-W 10

TL (m

m)

0

100

200

300

400

500

600

700SeptN ovJanM arM ay

aabb

aabb

a a

aa

a

a

aa

a

G riffin

G R -04-51 G R -04-A G R -04-B G R -04-C G R -04-D

TL (m

m)

0

100

200

300

400

500

600

700SeptN ovJanM arM ay

aab

abc

bc c

aab

ab

bb

a

a

a

a

aa

a

a

a

a

a

a

a

a

a

72

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).

O ran ge

O R -04-12 O R -04-13 O R -04-B O R -04-W 1 O R -04-W 5

SVL

(mm

)

0

50

100

150

200

250

300

350S ep tN ovJanM arM ay a

ab a

b

bb

aab

ab

b b

a

aa

a

aa

a

a

a

a

aa

a

a

a

G riffin

G R -04-51 G R -04-A G R -04-B G R -04-C G R -04-D

SVL

(mm

)

0

50

100

150

200

250

300

350SeptNovJanM arM ay

aab

abc

bc c

aab

ab

b b

a

aa

a

aa

a

a

a

a

a

a

a

a

a

E m e r e ld a

E M - 0 4 - 0 1 E M - 0 4 - 0 2 E M - 0 4 - 0 3 E M - 0 4 - 0 4 E M - 0 4 - 1 1

SVL

(mm

)

0

5 0

1 0 0

1 5 0

2 0 0

2 5 0

3 0 0

3 5 0S e p tN o vJ a nM a rM a y

aab

bcc d

aab

ab

ab

b

a

b bbb

aab

bc

bc

c

a

a

aa

a

A popka

A P-04-10 A P-04-W 2 A P -04-W 10

SVL

(mm

)

0

50

100

150

200

250

300

350SeptNovJanM arM ay

a bb

aab

b

a aa

a aa

aa

a

73

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

time. Significant differences among sites were determined by Tukey Multiple Comparison Analysis (p < .05).

O ran ge

O R -04-12 O R -04-13 O R -04-B O R -04-W 1 O R -04-W 5

Hea

d (m

m)

0

20

40

60

80

100S ep tN ovJanM arM ay

aab

ab b

ab

aa aa a

ab

ab

ab

ab

a

b b

ab a

ba a

aa

a

E m ere ld a

E M -04-01 E M -04-02 E M -04-03 E M -04-04 E M -04-11

Hea

d (m

m)

0

20

40

60

80

100S eptN o vJanM arM ay

abb bc

a aa

aa

aab b bb

aab

bb b

aab

ab

ab

b

A po pka

A P -04-10 A P -04-W 2 A P -04-W 10

Hea

d (m

m)

0

20

40

60

80

100Sep tN o vJanM arM ay

ab b

aab

b

a aa

aa

a

a

aa

G riffin

G R -04-51 G R -04-A G R -04-B G R -04-C G R -04-D

Hea

d (m

m)

0

20

40

60

80

100SeptN o vJanM arM ay

a a a bb

a a a aaa

aa a

aa a

a a

a

aab

ab

ab

b

74

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).

G riffin

G R -04-51 G R -04-A G R -04-B G R -04-C G R -04-D

Wei

ght (

g)

0

100

200

300

400

500

600

700

800SeptN ovJanM arM ay

baaaa

a

aba

bb

b

a

a

a

a

a

a

a

a

a

a

a

a

a

a

a

O range

O R -04-12 O R -04-13 O R -04-B O R -04-W 1 O R -04-W 5

Wei

ght (

g)

0

100

200

300

400

500

600

700

800SeptN ovJanM arM ay

aab bbb

a aba

bab

b

aab

ab

bc

c

a

bb

b

ab

a a a a a

Em erelda

EM -04-01 EM -04-02 EM -04-03 EM -04-04 EM -04-11

Wei

ght (

g)

0

100

200

300

400

500

600

700SeptN ovJanM arM ay

ab ba a a

a aab

ab

b

a

bb

bb

a

b

b

bb

a

aba

b

ab b

A p opka

A P -04-10 A P -04-W 2 A P -04-W 10

Wei

ght (

g)

0

100

200

300

400

500

600

700S eptN ovJanM arM ay

aab

ba bb

a

a

a

a

a

a

a

a

a

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).

G riffin

G R -04-51 G R -04-A G R -04-B G R -04-C G R -04-D

Thyr

oid

Wt.

(g)

0 .00

0.01

0.02

0.03

0.04

0.05

0.06

0.07SeptN ovJanM arM ay

aab

ab

b b

a

ab

ab

b

b

aa

a

a a

a

a a

a

a

a

a

a a

a

O range

O R -04-12 O R -04-13 O R -04-B O R -04-W 1 O R -04-W 5

Thyr

oid

Wt (

g)

0 .00

0 .01

0 .02

0 .03

0 .04

0 .05Sep tN o vJanM arM ay

a

aa

a

a

a

a

a

a

a

aa

a

a

a

aa

a

a

a

aa

a

a

a

E m ere lda

E M -04-01 E M -04-02 E M -04-03 E M -04-04 E M -04-11

Thyr

oid

Wt (

g)

0 .00

0.01

0.02

0.03

0.04

0.05

0.06S eptN ovJanM arM ay

a

b b b b

a

ab

b

b b

a

b

bb

b

a ab

ab a

b

b

a aa a a

A popka

A P -04-10 A P -04-W 2 A P -04-W 10

Thyr

oid

Wt (

g)

0 .00

0.01

0.02

0.03

0.04

0.05

0.06SeptN ovJanM arM ay

a

ab

b

a a aa

a

a

a

a

a

a

a

a

76

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).

G riffin

G R -04-51 G R -04-A G R -04-B G R -04-C G R -04-D

Live

r Wt (

g)

0

2

4

6

8

10

12

14

16

18SeptN ovJanM arM ay

aab

ab

bc c

a

ab

bb

b

aa

a

a

a

a

a a

a

a

a

a

a

a

a

O rang e

O R -04-12 O R -04-13 O R -04-B O R -04-W 1 O R -04-W 5

Live

r Wt (

g)

0

2

4

6

8

10

12

14

16

18S ep tN o vJanM arM ay

aab a

b

bb

a

ab

b

b

b

a

aa

a

a a

a

a

a

a

aa

a

a

a

E m ere lda

E M -04-01 E M -04-02 E M -04-03 E M -04-04 E M -04-11

Live

r Wt (

g)

0

2

4

6

8

10

12

14

16S ep tN o vJanM arM ay

aab

bc cd

a ab

ab a

bb

aab

bb

b

a

b

b

b

b

a

a

a

a

a

A popka

A P-04-10 A P-04-W 2 A P -04-W 10

Live

r Wt (

g)

0

2

4

6

8

10

12

14SeptN ovJanM arM ay

a ab

a

a

a

a

aa

a

a

a

a

a

a

77

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).

Total Thyroxine

Oct Nov Dec Jan Feb Mar Apr May

TT4

(ng/

mL)

0

2

4

6

8

10

12

14

16

18GROREMAP

b

cd

a

db

ab

ab

a

a aa

a

a

bb

b aa

a

a

bc

ab

a

c

aa aa aa

b

b

Free Thyroxine

Oct Nov Dec Jan Feb Mar Apr May

FT4

(pg/

mL)

0

1

2

3

4

5

6GROREMAP

b b b

a

aa

b bb

ab

ab

a

a

b

b

c

a

b

bb

a

ab a

b

b aaa

a

a

a

a a

78

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).

G riffin

O c t N o v D e c J an F e b M a r A p r M a y

TT4

(ng

/ ml)

0

2

4

6

8

1 0

1 2

1 4

1 6

1 8G R -04-51 G R -04 -A G R -04-B G R -04-C G R -04-D

a

aa

bb

a

a

a

a

a aaa

a

ab

ab

ab

ab

a

aa

a

a

a a

a

a

aa

b

a a

ab

ab

a

aa a

a

A p o p ka

O c t N o v D e c J a n F e b M a r A p r M a y

TT4

(ng/

mL)

0

2

4

6

8

10

12

14

16

18

20AP -04-10 AP -04 -W 2 AP -04 -W 10

a a

ba

a

a

a

a

a

a

aa a

a

a

a

a

a

aa

a

a

bb

O ran ge

O ct N o v D ec J an F eb M a r A p r M ay

TT4

(ng

/ ml)

0

2

4

6

8

10

12

14

16

18O R -04-12 O R -04 -13 O R -04 -B O R -04-W 1 O R -04-W 5

a

a

a

a

aa

a

aa

a

a

a a

aa

b b

ab

a

ab

aa

a

aa

a

aaa

a

aa

aa

a

a

b

a

a

a

E m ere lda

O ct N o v D ec Jan F eb M ar A p r M ay

TT4

(ng

/ ml)

02468

10121416182022

E M -04-01 E M -04-02 E M -04-03 E M -04-04 E M -04-11

d

bc

cd

ab

a

a

aa

a a

aa

a

aa

a

aa

aa

a

aa

a

ab

ab

ab

b

a ab

a

b

ab

b

a

aaa a

a

79

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).

G riffin

O c t N o v D e c J a n F e b M a r A p r M a y

FT4

(pg/

mL)

0

1

2

3

4

5

6

7G R -04-5 1 G R -04 -A G R -04-B G R -04-C G R -04-D

ab

b

a

b b

b

ab

a

ba

ab

a aaa

a

a ab a

b

b

ab

c

ab

aabc b

c

a

a

a a

a

a

a

a

a

a

a

aa

a

a

O ra n g e

O c t N o v D e c J a n F e b M a r A p r M a y

FT4

(pg/

mL)

0

1

2

3

4

5

6

7

8O R -0 4-1 2 O R -0 4 -1 3 O R -0 4 -B O R -0 4-W 1 O R -0 4-W 5

a

a

aaa

a

a

a

aa a

bb

b ba a

a a a

abc

ab

c

a

bc

a

a

a

a

a

a

aa a a

a

c

b

b b

E m e re ld a

O c t N o v D e c J a n F e b M a r A p r M a y

FT4

(pg/

mL)

0

1

2

3

4

5

6E M -0 4 -0 1 E M -0 4 -0 2 E M -0 4 -0 3 E M -0 4 -0 4 E M -0 4 -1 1

a ab

ba

b

ba

a

a

a

a

a

a

a

aa

a

a

a

a

aa

ab

a

ab

ab

b

a

a

a

a

a

a

a

a

a

aa

aa

aa

A p o p ka

O ct N o v D e c J a n F e b M a r A p r M a y

FT4

(pg/

mL)

0

1

2

3

4

5

6

7

8

9AP -04 -10 AP -04 -W 2 AP -04 -W 10

a

bb

aa

a a

a

aaa

a

a

aa

a

aa

aaa

b

b

a

80

Testosterone

Estradiol

GHRH

Estradiol

Neuropeptide Y

GLP-I

Estradiol

Norepinephrine

Galanin

Somatotroph cellsof anterior pituitary

Growth Hormone Release

Ration SizeProtein Intake

StarvationAcute Stress

Chronic Stress

TSH

T3T4

DOPA

Dopamine

Norepinephrine

5-HydroxytryptamineSomatostatin-25

Somatostatin-28

SRIFIGF-I

NPY

NMA

GHTRH

SRIFGnRH

Estradiol

SRIF

Bombesin

ExerciseOvulation

TemperatureDaylength

Seawater Adaptation CCK

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

hormone. Adapted from Mommsen, 1998.

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

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

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

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

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

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 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

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 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

88

Table 2-8. Multiple linear regression analysis of hatchling alligator growth rates, thyroid hormone secretory rates and organochlorine 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

89

LIST OF REFERENCES

Bjöornsson, B.T., Johansson, V., Benedet, S., Einarsdottir, I.E., Hildahl, H., Agustsson, T., Jönsson, 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., Lofts, B. (1980) Annual cycles in levels of pituitary and plasma gonadotropin, gonadal steroids, and thyroid activity in the Chinese cobra (Naja naja). General and Comparative Endocrinology 42: 477-493.

Brouwer, A., Morse, D.C., Lans, M.C., Schuur, G., Murk, A.J., Klasson-Wehler, E., Bergman, A., Visser, T.J. (1998) Interactions 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., Eales, 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 juvenile rainbow trout (Oncorhynchus mykiss) to 3,3’4,4’,5- pentachlorobipheyl (PCB 126). Aquatic Toxicology 59(3-4): 139-152.

Brucker-Davis, F. (1998) Effects of environmental synthetic chemicals on thyroid function. Thyroid 8(9): 827-856.

Bustnes, J.O., Miland, O., Fjeld, M., Erikstad, K.E., Skaare, J.U. (2005) Relationships between ecological variables and four organochlorine pollutants in an artic glacous gull (Larus hyperboreous) population. Environmental 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..

90

Chang, L., Munro, L.A., Richardson, S.J., Schreiber, G. (1999) Evolution of thyroid hormone binding by transthyretins in birds 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: interaction of organochlorine compounds with the thyroid receptor, transthyretin, and thyroid-binding globulin. Environmental Health Perspectives 107(4): 273-278.

Conley, A.J., Elf, P., Corbin, C.J., Dubowsky, S., Fivizzani, A., Lang, J.W. (1997) Yolk steroids decline during sexual differentiation in the alligator. General and Comparative Endocrinology 107: 191-200.

Cooley, H.M., Fisk, A.T., Wiens, S.C., Tomy, G.T., Evans, R.E., Muir, D.C.G. (2001) Examination of behavior and thyroid histology of juvenile 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 hormone 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) Dependence of body growth on thyroid activity in turtles. The Journal of Experimental 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 poikilotherms. Can. J. Zool. 63: 1217-1231.

Elsey, R.M., Joanen, T., McNease, L., Kinler, N. (1992) Growth rates and body condition factors of Alligator mississippiensis in coastal Louisiana wetlands: a comparison of wild and farm-raised juveniles. Comparative Biochemistry and Physiology 103A(4): 667-672.

Elsey, R.M., Joanen, T., McNease, L., Lance, V. (1990) Growth rate and plasma corticosterone levels in juvenile alligators maintained at different stocking densities. The Journal of Experimental Zoology 255: 30-36.

Ferguson, M.W.J. (1985) The reproductive biology 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.

91

Fowles, J.R., Fairbrother, A., Trust, K.A., Kerkvliet, N.I. (1997) Effects of Arochlor 1254 on the thyroid gland, immune function, and 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 Journal of Physiology and Pharmacology 46:59-66.

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 disrupting 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 concentrations and male phallus size in juvenile alligators from seven Florida lakes. General and Comparative 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 disrupting contaminants (EDCs) during embryonic development in wildlife. Environmental Health Perspectives 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 Everglades, 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., Guillette 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.

92

Hulbert, A.J. (2000) Thyroid hormones and their effects: a new perspective. Biol. Rev. 75:519-631.

Hulbert, A.J. and Williams, C.A. (1988) Thyroid 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) Further evidence for reciprocal interactions between the annual sexual and thyroid cycles in male Peking ducks. General and Comparative Endocrinology 37: 44-51.

Jallegeas, M., Tamisier, A., Assenmacher, I. (1978) A comparative study of the annual cycles of sexual and thyroid function in male Peking ducks (Anas platyrhynchos) and teal (Anas crecca). General and Comparative Endocrinology 36: 201-210.

Jefferies, D.J. and French, M.C. (1972) Changes induced in the pigeon thyroid by p,p, DDE and dieldrin. Journal of Wildlife Management 36(1): 24-30.

Jefferies, D.J. and Parslow, J.L.F. (1972) Effect 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 alligator. International Zoo Yearbook 19: 61-66.

Kar, A. and Chandola-Saklani, A. (1984) Circulating thyroid hormone concentrations in relation to seasonal events in the male Indian Garden lizard, Calotes versicolor. General and Comparative Endocrinology 60: 14-19.

Kato, Y., Ikushiro, S., Haraguchi, K., Yamazaki, 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 polychlorinated 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) Thyroxine 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 thyroxine 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.

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) Temperature dependence of in vitro pituitary, testis, and thyroid secretion 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 Comparative Endocrinology 59: 130-139.

Licht, P., Wood, J.F., Wood, F.E. (1985) Annual and diurnal cycles in plasma testosterone and thyroxine 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 perchlorate exposure?. Toxicological Science 82: 106-113.

McNabb, M.A. and Fox, G.A. (2003) Avian thyroid development in chemically contaminated environments: is there evidence 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, their 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 contaminants. Journal of Wildlife Diseases 22(1): 60-70.

Moccia, R.D., Leatherland, J.F., Sonstegard, 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) Thyroxine concentrations in blood serum and pericardial fluid of metamorphosing tadpoles and of adult frogs. General and Comparative Endocrinology 39: 343-349.

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 Comparative Endocrinology 120: 198-211.

Nishimura, N., Miyabara, Y., Sato, M., Yonemoto, J., Tohyama, C. (2002) Immunohistochemical localization of thyroid stimulating hormone induced by a low oral dose of 2,3,7,8-tetrachlorobenzo-p-dioxin 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 endothermy in the red Winged Blackbird (Agelaius phoeniceus). General and Comparative Endocrinology 116: 204-212.

O’Steen, S. and Janzen, F.J. (1999) Embryonic temperature affects metabolic compensation and thyroid hormones in hatchling snapping turtles. 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 the developing brain to thyroid abnormalities: environmental insults to the thyroid system. Environmental Health Perspectives 102 (Suppl. 2): 125-130.

Prapunpoj, P., Richardson, S.J., Schreiber, G. (2002) Crocodile transthyretin: structure, function, and evolution. Am. J. Physiol. Inter. Comp. Physiol. 283: R885-R896.

Prati, M., Calvo, R., Morreale de Escobar, G. (1992) L-thyroxine and 3,5,3’- triiodothyronine concentrations in the chicken egg and in the embryo before and after the onset of thyroid function. Endocrinology 130(5): 2651-2659.

Rauschenberger, R.H. (2004) Developmental mortality in American alligators (Alligator mississippiensis) exposed to organochlorine pesticides. PhD dissertation. University of Florida.

Rauschenberger, H.R., Sepúlveda, M.S., Wiebe, J.J., Szabo, N.J., Gross, T.S. (2004) Predicting maternal body burdens of organochlorine pesticides from eggs and evidence of maternal transfer in Alligator mississippiensis. Environmental Toxicology and Chemistry 23(12): 2906-2915.

95

Rauschenberger, R.H., Wiebe, J.J., Buckland, J.E., Smith, J.T., Sepúlveda, 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, Switzerland and Cambridge UK.

Richardson, S.J., Monk, J.A., Shepherdley, C.A., Ebbesson, L.O.E., Sin, F., Power, D.M., Frappell, P.B., Köhrle, J., Renfree, M.B. (2005) Developmentally regulated thyroid hormone distributor in marsupials, a reptile 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-induced alterations in thyroid and vitamin a status from field studies of wildlife and fish. Journal of Wildlife Diseases 36(4): 615-635.

Rotstein, D.S., Schoeb, T.R., Davis, L.M., Glenn, 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 Thyroid and Its Diseases. Chapter 2: Thyroid Hormone Synthesis and Secretion from www.thyroidmanager.org/Chapter2/2-text.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, and insulin-like growth factor-I in ad libitum and food-restricted (altricial) European starlings and (precocial) Japanese quail. General and Comparative Endocrinology 101: 304-316.

Scollon, E.J., Carr, J.A., Cobb, G.P. (2004) The 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.

96

Sechman, A. and Bobeck, S. (1988) Presence of iodothyronines in the yolk of a hen’s egg. General and Comparative Endocrinology 69: 99-105.

Sellers, J.C., Wit, L.C., Ganjam, V.K., Ethridge, 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., Evans, B.K., Kuhn, E.R., Darras, V.M. (2002) Characterization of outer ring iodothyronine deiodinases in tissues of saltwater crocodiles. General and Comparative 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., Bragt, P.C., Notten, W.R.F. (1991) Interactions of halogenated industrial chemicals with transthyretin and effects on thyroid hormone levels in vivo. Archives of Toxicology 65: 15-19.

Verreault, J., Skaare, J.U., Jenssen, B.M., Gabrielsen, G.W. (2004) Effects of organochlorine contaminants on thyroid hormone levels in artic breeding Glaucous gulls, Larus hyperboreus. Environmental Health Perspectives 112(5): 532-537.

Very, N.M. and Sheridan, M.A. (2002) The role of somatostatins in the regulation of growth in fish. Fish Physiology and Biochemistry 27: 217-226.

Wade, M.G., Parent, S., Finnson, K.W., Foster, W., Younglai, E., McMahon, A., Cyr, D.G., Hughes, C. (2002) Thyroid toxicity due to subchronic mixture of 16 organochlorines, lead, and cadmium. Toxicological Sciences 67: 207-218.

Waritz, R.S., Steinberg, M., Kinoshita, F.K., Kelly, C.M., Richter, W.R. (1996) Thyroid function and thyroid tumors in toxaphene-treated 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.

97

Wiebe, J.J., Sepúlveda, 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 low-dose 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 days and thyroxine program American tree sparrows for seasonality: evidence for temporal flexibility for the breeding of euthyroid females. General and Comparative 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 status of mummichogs (Fundulus heteroclitus) from polluted versus a reference habitat. Environmental Toxicology and Chemistry 18(12): 2817-2823.

98

BIOGRAPHICAL SKETCH

Jonathan James Wiebe was born on December 15, 1969, in Pensacola, Florida, and

is the son of Ralph and Linda Wiebe. Jon graduated from Gainesville 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 professional career in the 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 variety of different species. Jon is particularly

proud of the collaborative work that he has achieved with Dr. Tim Gross, Dr. Heath

Rauschenberger and Janet Scarborough in the area of alligator ecotoxicology.