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ECOTOXICOLOGY OF MORELET'S CROCODILE IN BELIZE
by
THOMAS ROBERT RAINWATER, B.S., M.S.
A DISSERTATION
IN
ENVIRONMENTAL TOXICOLOGY
Submitted to the Graduate Faculty of Texas Tech University in
Partial Fulfillment of the Requirements for
the Degree of
DOCTOR OF PHILOSOPHY
Approved
August, 2003
ACKNOWLEDGMENTS
Numerous people were instrumental in the success of this project, and to thank
each of them here would likely double the length of this document. Thus, although I will
only mention a fraction of those who have helped me reach this point, I acknowledge the
contributions of the many others and thank them for their help over the last nine years.
First, I wish to thank Drs. Todd Anderson, George Cobb, Lou Densmore, Scott
McMurry, and Ernest Smith for serving as my doctoral committee and for their flexibility
and guidance throughout the duration of this project. 1 particularly thank my academic
advisor. Dr. McMurry, for a 1994 conversation in a Clemson University hallway where
he encouraged me to pursue a dissertation project on a topic in which I was truly
interested instead of just taking what was available at the time. That advice was well-
founded, and after two years of writing grant proposals, we finally secured funding for
what would become a dream project for me. The ensuing four years in Belize were some
of the most rewarding of my life, both personally and professionally. Dr. McMurry was a
superb mentor, and I sincerely appreciate his friendship and guidance over the years.
This research was funded by the U.S. Environmental Protection Agency (grant #
R826310 to Dr. Scott McMurry), the Royal Geographical Society (UK), Lamanai Field
Research Center (Belize), and Texas Tech University, Lubbock, Texas, USA. I am
particularly grateful to Mrs. Fran Carter and the ARCS Foundation, Inc. of Lubbock for
the C.B. Carter Memorial Scholarship which supported me from 2001-2003. I also wish
to thank Dr. Ron Kendall and The Institute of Environmental and Human Health
(TIEHH)/Department of Environmental Toxicology for constant support throughout my
tenure as a doctoral student.
In Belize, research permits were granted by the Ministry of Natural Resources,
Forestry Department. Emil Cano, Earl Codd, Raphael Manzanero, Natalie Rosado, and
Marcelo Windsor assisted with the permitting process and gladly provided us with export
permits necessary to transport samples to the U.S., often with short notice. Robert
Noonan generously allowed access to Gold Button Ranch, and Victor Cmz and his family
often provided meals and accommodations there.
u
The success of this project is largely due to Mark and Monique Howells,
proprietors of the Lamanai Outpost Lodge. In 1995, Mark, Monique, and the late Colin
Howells invited Dr. McMurry and me to come to Belize and develop a Morelet's
crocodile research project at Lamanai. 1 showed up on their doorstep in 1997, and since
that time they have generously provided accommodations, meals, and logistical support
to me and the rest of our research team. My four years living and working with Mark and
Monique at Lamanai were highly enjoyable, and I am most grateful for their continued
friendship.
While at Lamanai, I was fortunate to meet and work with numerous people from a
variety of backgrounds, and I learned much from them. 1 thank Dr. Brock Fenton for his
professional advice and for allowing me to assist him in netting bats in the Lamanai
Reserve. Dr. Elizabeth Graham's enthusiasm for archaeology was contagious and her
energy inspiring. In addition, many times she allowed us to use her truck when the
Clemson Belle II was not operational, and 1 thank her for her generosity. I am grateful to
Dr. Hal Markowitz for his constant advice and support over the years and for his
hospitality when I visited San Francisco. Dr. Steve Reichling is thanked for sharing his
knowledge about tarantulas, allowing me to join him on his evening snake walks, and for
inviting me to speak on the crocodile project at the Memphis Zoo in 2000. Denver Holt
is thanked for allowing me to assist him in netting birds near Irish Creek and particularly
for inviting my father and me to join him on a trip to Tanzania in 1998. I am especially
grateful to my fellow graduate students and field technicians at Lamanai for four years of
countless memories: Tanny Brown, Amanda Colombo, Travis Crabtree, Leslie Cornick,
Jen Dever, Marcus England, Debbie Green, Laura Howard, Steve Lawson, Blanca
Manzanilla, Araba Oglesby, Terry Powis, Brenda Salgado, Patti Schick, David Ray,
Thomas Rhott, Norbert Stanchly, Hilary Swarts, and Amy Webbeking. John Ratcliffe is
thanked for inviting me to help collect vampire bats in the caves near San Ignacio. I also
thank Mrs. Betty Rowe for her friendship, special interest in the crocodile project, and
zest for life. I am especially grateful to Katie Eckert for help in the field, hospitality in
California, and most of all, four years of patience, friendship and companionship.
i l l
Numerous others in Belize also contributed to the success of this project.
Richard and Carol Foster generously provided accommodations, meals, freezer space,
and good conversation when work took me to central Belize. They also shared
equipment and field rations with Dr. McMurry and me during an expedition on the upper
Macal River after ours were lost in a storm. Mike, Anita, and Christine Tupper,
proprietors of Cheers on the Western Highway, provided an additional base camp in
central Belize, complete with great food, jumbo lime juice, freezer space, and a fax
machine. John and Carolyn Carr allowed access to Banana Bank Lagoon. Bmce
Cullerton, Rainforest Mechanic, is thanked for assistance in catching crocodiles at Cox
Lagoon and for constructing a canoe rack for the Belle. Matt and Marga Miller of
Monkey Bay Wildlife Sanctuary provided a place to pitch a tent and unlimited access to
the biogas latrine. Mick Mulligan and his family provided meals and accommodations in
Belize City and took care of my dog when 1 had to leave the country unexpectedly. Dr.
Sheila Schmeling of Corozal Town provided veterinary advice and a place to store gear
before entering Mexico. And, she is sincerely thanked for helping CD before she made
the joumey. Sharon Matola is thanked for allowing me to sample crocodiles at the Belize
Zoo. Jan Meerman and Martin Meadows assisted in the field and generously shared their
observations on crocodiles in other regions of Belize. Alan Kutcher provided constant
motivation in 1997, despite his concern over the Rambo aspect of the project. Special
thanks go to Ruben Arevalo, Benjamin Cruz, Luis Gonzales, and Jose Torres for
assistance in the field under what were at times adverse conditions.
I am also grateful for the friends who came to Belize to visit and assist in
fieldwork. It was always nice to see a face from home, and visitors usually came bearing
a fresh supply of mail, books, and papers. Tony and Julie Hawkes, Lynn Sholtis, and Ted
Wu are thanked for making the trip and helping with crocodile captures and egg
collections. Jim Stoker shook off being stung by a Portuguese Man-of-War and helped
catch crocodiles at Gold Button Lagoon. Sean Richards helped with nest surveys and
hatchling measurements while camped at Barter Town and is especially thanked for
photographing the large blacktail indigo captured at Gold Button Lagoon.
IV
I am grateful to Dr. Lou Guillette, Dr. Drew Grain, Matthew Milnes, Thea
Edwards, and Gerry Binczik for inviting me to the University of Florida in 2001 and
helping me troubleshoot the sex-steroid hormone radioimminoassays. Likewise, I thank
Dr. Kyle Selcer at Duquesne University for providing the vitellogenin antibody and
helping me with the vitellogenin assays.
Dr. Phil Smith and Kevin Reynolds are thanked for nine years of camaraderie in
the McMurry lab at both Clemson and Texas Tech. Our numerous field excursions in
South Carolina, Alabama, and the fertile cottonmouth-hunting grounds of east Texas
were most enjoyable, and 1 look forward to our future collaborations.
Dr. Steve Piatt is the godfather of the Belize crocodile research project, and
without his pioneering efforts in the country during the early 1990s, the three masters
degrees (with a fourth soon to be completed), three doctoral degrees (not including his
own), and numerous scientific publications that have resulted from this project would not
have come to fruition. Steve and Trouble, his feisty canine companion, first arrived in
Belize in 1992 and subsequently laid the groundwork for what would eventually blossom
into a multi-year, multi-disciplinary research project on Morelet's crocodiles. As a Ph.D.
student at Clemson University, Steve examined the status and ecology of Morelet's
crocodile in Belize. In the spring of 1995, he invited me to come to Belize to collect non
viable crocodile eggs for environmental contaminant analysis. Dr. McMurry and I took
Steve up on his offer and met him in Belize in July of that year. During that trip, two
things occurred that set the stage for the expansion of the crocodile project: (1) we
collected eggs that were found to contain multiple environmental contaminants,
providing a basis for an ecotoxicological investigation, and (2) Steve introduced us to
Colin, Mark, and Monique Howells who invited us to come to their lodge and develop a
crocodile project at Lamanai. I retumed to Belize in February 1997 to begin the
crocodile ecotoxicology project and spent the first month with Steve surveying Morelet's
and American crocodiles in the coastal zone. During that time, Steve imparted to me the
many skills necessary for working in Belize, the least of which was catching crocodiles.
In addition, he introduced me to numerous people that would eventually become good
friends and vital to the success of the project. Steve's reputation as a researcher
throughout the country and his respectful relationship with Belize Forestry Department
gave me instant credibility (founded or not), and 1 was able to start my work in Belize
with a fraction of the obstacles I would have otherwise faced without his influence.
Steve's involvement in the crocodile project continues today, and it has been a privilege
to call him a colleague and friend.
Lastly, I thank my family for 36 years of love and encouragement. My
grandparents, Lonnell Hiers and Madison and Mattie Virginia Rainwater have provided
constant support, despite the concems they must have had (and perhaps still do!) about a
36-year-old grandson who is still in school. My brother John has always been highly
enthusiastic and supportive of my graduate school career, and has constantly helped me
in many ways over the years. And, he and his wife Kelly have provided me with an
incredible nephew, Hiers, and two incredible nieces, Tumer and Price. My uncle. Dr.
Thorn Hiers, instilled in me a propensity to travel and constantly provided me with
humorous and encouraging anecdotes about his graduate school and travel experiences,
which always helped me keep things in perspective. And, his home on Sullivan's Island,
South Carolina has always been a sanctuary for me when I have wanted to get back to the
low country. Finally, I thank my parents, James and Anna Rose for their love, support,
patience, and friendship. In addition to teaching me the values of honesty, respect, and
hard work, they have also been incredibly patient and understanding during what at times
may have seemed like "The Degree That Would Never Be Attained". Most special to me
has been that we have been able to enjoy the ride together. Over the years, they have
visited me at numerous field sites or residences in South Carolina, Colorado, Iowa, and
California. In 2000, they traveled to Belize and helped catch crocodiles, and just last
week they came to Lubbock to attend my dissertation defense. I greatly look forward to
what lies ahead and thank them for teaching me to enjoy the joumey.
VI
TABLE OF CONTENTS
ACKNOWLEDGMENTS ii
ABSTRACT x
LIST OF TABLES xiii
LIST OF FIGURES xv
CHAPTER
1. INTRODUCTION 1
References 11
II. PLASMA VITELLOGENIN EXPRESSION IN MORELET'S C R 0 C 0 D D : . E S F R O M C O N T A M I N A T E D HABITATS IN NORTHERN BELIZE 27
Abstract 27 Introduction 28 Materials and Methods 31
Study Sites 31 Animal Collections and Sampling 32 Gel Electrophoresis 33 Immunoblotting 33 Enzyme-Linked Immunosorbent Assay (ELIS A) 34 Statistical Analyses 35
Results 35 Discussion 36 Conclusions 41 References 43
Vll
III. SEX-STEROID HORMONE CONCENTRATIONS IN MORELET'S CROCODILES FROM CONTAMINATED HABITATS IN NORTHERN BELIZE 56
Abstract 56 Introduction 57 Materials and Methods 61
Study Sites and Sample Collection 61 Steroid Hormone Radioimmunoassays 63 Statistical Analyses 64
Results 65 Mean Body Size 65 Mean Hormone Concentrations 65 Body Size and Hormone Concentrations 66
Discussion 67 Conclusions 74 References 76
IV. PHALLUS SIZE AND PLASMA TESTOSTERONE CONCENTRATIONS IN MALE MORELET'S CROCODILES FROM CONTAMINATED HABITATS IN NORTHERN BELIZE 99
Abstract 99 Introduction 100 Materials and Methods 104
Study Sites 104 Animals, Blood Sampling, and Morphometries 104 Steroid Hormone Radioimmunoassay 105 Statistical Analyses 107
Results 107 Animals Captured and Sampled 107 Phallus Morphometries 108 Plasma Testosterone Concentrations 109
Discussion 110 Conclusions 114 References 116
Vlll
CONCLUSIONS 136
Study Summary 136 Comparison of this Study with Studies on Florida Alligators 139 Uncertainties 141 Future Research Directions 143 Conclusions 144 References 146
IX
ABSTRACT
Over the last two decades, population declines and reproductive impairment have
been observed in American alligators (Alligator mississppiensis) inhabiting Lake
Apopka, a highly contaminated lake in Florida, USA. Juvenile alligators from the lake
have exhibited altered sex-steroid hormone concentrations, abnormal gonadal
morphology, and reduced phallus size compared to alligators from a reference lake. No
direct cause-effect relationship has been established between these reproductive and
endocrine anomalies and environmental contaminants, but results of laboratory and field
investigations suggest the potential for contaminant-induced endocrine disruption at
various levels of organization in these animals. Although various environmental
contaminants considered to be endocrine disrupters have been found in eggs and tissues
of crocodilians worldwide, no studies have yet investigated endpoints of endocrine
disruption in wild crocodilians outside of Florida. The primary objective of this study
was to address this data gap by examining ecotoxicological endpoints in another
crocodilian species living in habitats contaminated with endocrine-disrupting chemicals
(EDCs), and where appropriate, compare results from this study with those observed for
alligators in Florida.
During a pilot study in 1995, multiple organochlorine (OC) pesticides considered
to be EDCs were found in eggs of Morelet's crocodiles (Crocodylus moreletii) from three
localities in northem Belize. Based on these findings and previous data from Florida
showing egg contamination, population declines, and reproductive abnormalities in
alligators exposed to many of the same chemicals, a multi-year study was initiated to
examine various endpoints of contaminant exposure and response in Morelet's crocodiles
living on contaminated and reference sites in northem Belize. Gold Button Lagoon, a
man-made lagoon from which contaminated crocodile eggs were collected in 1995, was
selected as the contaminated site for this study, while New River Watershed, a more
remote site with fewer anthropogenic inputs than Gold Button Lagoon, was selected as
the reference site.
Three primary endpoints of endocrine disruption were evaluated in this study.
First, vitellogenin induction was examined as an endpoint of exposure to exogenous
estrogens or estrogen-mimicking contaminants. Vitellogenin is an egg-yolk precursor
protein expressed in all oviparous and ovoviviparous vertebrates. Male and juvenile
females normally have no detectable concentration of vitellogenin in their blood but can
produce it following stimulation by an exogenous estrogen, such as an EDC. Thus, the
presence of vitellogenin in the blood of male or juvenile female crocodiles can serve as
an indicator of exposure to an estrogen-mimicking EDC. Of 358 males and juvenile
females sampled in this study, no vitellogenin induction was observed, suggesting these
animals were likely not exposed to estrogenic xenobiotics. However, many of the
animals sampled were later found to contain OC pesticides in their caudal scutes,
confirming they had in fact been exposed to OCs (and EDCs). These data suggest the
lack of a vitellogenic response should not necessarily be interpreted as an indication that
no exposure or other contaminant-induced biological response has occurred.
Second, plasma sex-steroid hormone concentrations were examined as an
endpoint of response to EDC exposure in crocodiles from the two study sites. The
selection of this endpoint was based on numerous studies reporting altered concentrations
of estradiol-17P (E2) and testosterone (T) in alligators from Lake Apopka and other
contaminated lakes in Florida. In the present study, few inter-site differences in plasma
hormone concentrations were noted. No significant differences in plasma E2
concentrations were detected between sites. However, juvenile males and females from
the contaminated site exhibited significantiy reduced plasma T concentrations compared
to juvenile males and females from the reference site, respectively. This finding was
consistent with results from previous studies on alligators in Florida. No other inter-site
differences in hormone concentrations were observed. Relationships between body size
and hormone concentrations were variable and showed no clear pattem.
Third, male phallus size was examined as a second endpoint of response to EDC
exposure in crocodiles from the two study sites. Concurrent with reductions in plasma T
concentrations, male alligators from Lake Apopka and other contaminated lakes in
Florida have exhibited smaller phallus size compared to animals from a reference lake.
XI
Researchers speculate that abnormal hormone concentrations during early life stages may
affect anatomical stmctures dependent on these hormones for proper growth and
development (i.e., genitalia). p,p '-DDE, a known anti-androgen, has been detected in
alligator eggs and serum from Lake Apopka and was also detected in eggs and scutes of
Morelet's crocodiles from the two Belize study sites. Thus, in the present study, male
crocodile phallus size and plasma T concentrations were examined as endpoints of
response to p,p '-DDE exposure as well as exposure to other contaminants. No
differences in mean phallus size were observed between sites, whereas mean plasma T
concentrations in juveniles from Gold Button Lagoon were significantly reduced
compared with those from New River Watershed.
It was discovered late in the study that New River Watershed exhibited a
contaminant profile similar to that observed at Gold Button Lagoon, with multiple OCs
detected at similar concentrations in sediments, crocodile eggs, and crocodile caudal
scutes at both sites. With the lack of a suitable reference site, it is thus unclear if steroid
hormone concentrations and male phallus size observed in this study are within the
normal range exhibited by Morelet's crocodiles living in non-contaminated habitats or if
they are altered in some way (e.g., increased, reduced). In addition, it is also unclear if
inter-site differences in plasma T are the result of exposure to EDCs, natural variation,
one or more undetermined factors (e.g., stress), or a combination of these factors.
In general, the results of this study indicate few or no effects of EDC exposure on
Morelet's crocodiles inhabiting contaminated wetlands in northem Belize. However,
multiple uncertainties encountered in this study make inter-site and inter-study (crocodile
to alligator) comparisons difficult and some results equivocal. Thus, the potential effects
of EDCs and other contaminants on crocodiles inhabiting these sites should not be
assumed to be negligible. Long-term studies are essential to adequately assess the effects
of EDCs on crocodilian populations, as these animals are long-lived and many
contaminant-induced effects are organizational in nature, occurring during embryonic
development but not appearing until later in life.
Xll
LIST OF TABLES
2.1 Sex, number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for vitellogenin induction during this study 51
3.1 Sex, number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for plasma estradiol- 17p concentrations in this study 84
3.2 Sex. number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for plasma testosterone concentrations in this study 84
3.3 Mean (±SE) plasma concentrations of estradiol-17P (E2) (pg/ml) and testosterone (T) (ng/ml) in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize study 85
3.4 Results of linear regression analysis of hormone concentrations as a function of body size (TL) in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize 86
4.1 Sex, number, and size range (cm total length [TL]) of male crocodiles from New River Watershed and Gold Button Lagoon for which phallus size was measured in this study 122
4.2 Sex, number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for plasma testosterone concentrations in this study 122
4.3 Results of linear regression analysis comparing snout-vent length (SVL) and phallus size (tip length and cuff diameter) in male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize 123
4.4 Mean (±SE) of phallus tip length (mm) and cuff diameter (mm) of juvenile and adult male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize 123
4.5 Results of linear regression analysis comparing body size (SVL) and plasma testosterone (T) concentrations of male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize 124
Xlll
4.6 Mean (±SE) plasma testosterone (T) concentrations (ng/ml) in juvenile and adult male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize 124
4.7 Results of linear regression analysis comparing phallus size (tip length and cuff diameter) and plasma testosterone concentrations of male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize 125
5.1 Measured endpoints of response to endocrine-dismpting chemicals in juvenile male American alligators and Morelet's crocodiles living in contaminated habitats 151
5.2 Measured endpoints of response to endocrine-dismpting chemicals in juvenile female American alligators and Morelet's crocodiles living in contaminated habitats 153
5.3 Comparison of human-health protective concentration levels (PCLs) for various organochlorine (OC) contaminants in sediments in Texas, USA, ecological benchmarks for OCs in sediments, ecological screening values for sediments, and maximum concentrations of OCs in sediments collected from Gold Button Lagoon and New River Watershed, Belize. All contaminant concentrations are presented in mg/kg 154
5.4 Mean p,p'-DDE concentrations detected in eggs of various crocodilian species 156
XIV
LIST OF FIGURES
2.1 Map of Belize showing locations of the two study sites. GBL = Gold Button Lagoon, NRW = New River Watershed, NR = New River, NRL = New River Lagoon 52
2.2 SDS-PAGE gel (A) and Western blot (B) of plasma samples from vitellogenic (females) and non-vitellogenic (males) Morelet's crocodiles from northem The letter "S" indicates the lane containing pre-stained molecular weight standards. In the gel, two large molecular weight proteins were present in all six females (adults; samples collected during the breeding season) and none of the six males (4 adults, 2 juveniles) examined (A). In the Western blot, both proteins cross-reacted with vitellogenin antiserum (B), confirming both proteins to be vitellogenin. The two proteins may represent different vitellogenin forms, or the lower molecular weight protein may be a breakdown product of the larger protein 53
2.3 Vitellogenin induction (as a function of optical density at 450 nm) in plasma of Morelet's crocodiles from northem Belize. Numbers inside bars indicate the number of animals sampled within that group. Bars with different superscripts are significantly different. Only plasma from adult females contained vitellogenin (also see results of a gel electrophoresis and immunoblotting in Figure 2.2). AF = adult females; JF = juvenile females; AM = adult males; JM = juvenile males 54
2.4 Vitellogenin induction in the plasma of Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. Numbers above bars indicate the number of animals sampled per site within a group. Vitellogenin was only detected in the plasma of adult females. No significant difference in vitellogenin induction (adult females) and background absorbance (remaining groups) was observed between sites. AF = adult females; JF = juvenile females; AM = adult males; JM = juvenile males 55
3.1 Map of Belize showing locations of the two study sites. GBL = Gold Button Lagoon, NRW = New River Watershed, NR = New River, NRL = New River Lagoon 87
xv
3.2 Representative standard curves for hormone RlAs. A = estradiol-np (E2), using Endocrine Sciences antibody; B = E2, using ICN antibody; C = testosterone, using Endocrine Sciences antibody 88
3.3 Inter-site comparison of mean (±SE) plasma estradiol-17p (E2) concentrations in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. Numbers above bars indicate the number of animals sampled per site within a group. No significant difference in E2 concentrations within a group was observed. SJM = small juvenile males; SJF = small juvenile females; LJM = large juvenile males; LJF = large juvenile females; AM = aduh males; AF = adult females 89
3.4 Intra-site comparison of mean (±SE) plasma estradiol-17p (E2) concentrations in Morelet's crocodiles from New River Watershed (top) and Gold Button Lagoon (bottom), northern Belize. Numbers above bars indicate the number of animals sampled per site within a group. Bars with different superscripts are significantly different. LJM = large juvenile males; LJF = large juvenile females; AM = adult males; AF = adult females 90
3.5 Inter-site comparison of mean (±SE) plasma testosterone (T) concentrations in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. Numbers above bars indicate the number of animals sampled per site within a group. Asterisks indicate a significant (p < 0.05) difference in T concentrations within a group. SJM = small juvenile males; SJF = small juvenile females; LJM = large juvenile males; LJF = large juvenile females; AM = adult males; AF = adult females 91
3.6 Intra-site comparison of mean (±SE) plasma testosterone (T) concentrations in Morelet's crocodiles from New River Watershed (top) and Gold Button Lagoon (bottom), northem Belize. Numbers above bars indicate the number of animals sampled per site within a group. Bars with different superscripts are significantly different. SJM = small juvenile males; SJF = small juvenile females; LJM = large juvenile males; LJF = large juvenile females; AM = adult males; AF = adult females 92
3.7 Relationship between estradiol-17P (E2) concentration and body size in small juvenile (TL < 80 cm) crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. No E2-body size relationship was detected in females or males from either site 93
XVI
3.8 Relationship between testosterone (T) concentration and body size in small juvenile (TL < 80 cm) crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. No T-body size relationship was detected in females or males from either site : 94
3.9 Relationship between estradiol-17p (E2) concentration and body size in large juvenile (females, TL = 80-149.9 cm; males, TL = 80-179.9 cm) crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. A positive relationship between body size and E: was detected in females from New River Watershed but not in Gold Button Lagoon females or males from either site 95
3.10 Relationship between testosterone (T) concentration and body size in large juvenile (females, TL = 80-149.9 cm; males, TL = 80-179.9 cm) crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. No T-body size relationship was detected in females or males from either site 96
3.11 Relationship between estradiol-17P (E2) concentration and body size in adult (females, TL > 150 cm; males, TL > 180 cm) crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. A positive relationship between body size and E2 was detected in females and males from Gold Button Lagoon but not from New River Watershed site 97
3.12 Relationship between testosterone (T) concentration and body size in adult (females, TL > 150 cm; males, TL > 180 cm) crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. A positive relationship between body size and T was detected in females from Gold Button Lagoon but not in females from New River Watershed or males from either site 98
4.1 Map of Belize showing locations of the two study sites. GBL = Gold Button Lagoon, NRW = New River Watershed, NR = New River, NRL = New River Lagoon 126
4.2 Diagram of the crocodilian (alligator) phallus, showing its primary components (top) and points from which measurements were taken (bottom) from Morelet's crocodiles in this study 127
4.3 Relationship between total length (TL) and snout-vent length (SVL) in Morelet's crocodiles sampled in this study 128
XVll
4.4 Relationship between penis tip length (top) or penis cuff diameter (bottom) and snout-vent length (SVL) of juvenile male Morelet's crocodiles from two habitats in northern Belize. A significant relationship existed between SVL and both measures of phallus size at both sites 129
4.5 Relationship between penis tip length (top) or penis cuff diameter (bottom) and snout-vent length (SVL) of adult male Morelet's crocodiles from two habitats in northem Belize. A significant relationship existed between SVL and tip length at both sites and between SVL and cuff diameter at New River Watershed but not Gold Button Lagoon 130
4.6 Mean (±SE) phallus size of male Morelet's crocodiles sampled from New River Watershed and Gold Button Lagoon in northem Belize. No significant (p < 0.05) difference in either morphometric was observed between sites 131
4.7 Relationship between snout-vent length (SVL) and plasma testosterone (T) concentration in juvenile (top) and adult (bottom) male Morelet's crocodiles from two habitats in northem Belize. A positive SVL-T relationship was observed only in juveniles from New River Watershed 132
4.8 Mean (±SE) plasma testosterone (T) concentrations in male Morelet's crocodiles sampled from New River Watershed and Gold Button Lagoon in northem Belize. Numbers above bars indicate the number of animals sampled per site within a group. Bars with different superscripts are significantly different. An asterisk above a bar indicates a significant difference within that pair 133
4.9 Relationship between plasma testosterone (T) concentration and penis tip length (top) and cuff diameter (bottom) in juvenile Morelet's crocodiles from two habitats in northem Belize. Significant relationships were observed between T and both measures of phallus size at New River Watershed but not Gold Button Lagoon. However, these significant relationships disappear if the one individual with the exceptionally high T concentration (27.25 ng/ml) is removed from the analysis 134
XVlll
4.10 Relationship between plasma testosterone (T) concentration and penis tip length (top) and cuff diameter (bottom) in adult male Morelet's crocodiles from two habitats in northem Belize. No significant relationship was observed between T and either measure of phallus size 135
XIX
CHAPTER 1
INTRODUCTION
The term "ecotoxicology" was first introduced by Tmhaut (1977) as the "branch
of toxicology concerned with the study of toxic effects, caused by natural or synthetic
pollutants, to the constituents of ecosystems, animal (including human), vegetable and
microbial, in an integral context" (p. 152). Since that time, numerous variations of this
definition have been provided. Jorgensen (1990) maintained that ecotoxicology is the
study of toxic substances in the environment and their impact on living organisms, while
Cairns and Mount (1990) proposed it to be the study of the fate and effects of toxic
substances in ecosystems. Newman (1995) submitted that ecotoxicology is "the
organization of knowledge about the fate and effects of toxic agents in ecosystems on the
basis of explanatory principles" (p. 2). Hodgson et al. (1998) defined ecotoxicology as
"the study of environmental toxicants on populations and communities of living
organisms" (p. 176). Moriarty (1999) and Kendall et al. (2001) contended that
ecotoxicology is concerned with the effects of toxic substances on ecosystems. More
recently, Hoffman et al. (2003) suggested that in a broad sense, ecotoxicology can be
defined as "the science of predicting effects of potentially toxic agents on natural
ecosystems and on non-target species" (p. 1), while in a more restrictive sense it is "the
science of assessing the effects of toxic substances on ecosystems with the goal of
protecting entire ecosystems, and not merely isolated components" (p. 1). While most of
these definitions primarily emphasize contaminant effects at the ecosystem level, others
(Jorgensen, 1990; Hodgson et al., 1998; Hoffman et al., 2003) also stress effects on living
organisms at lower levels of organization (e.g., community, population, species). To
date, most studies on the ecotoxicology of wildlife have subscribed to the general
definition of Truhaut (1977), focusing broadly on the effects of environmental
contaminants on wildlife ecology; that is, the effect of contaminants on the interaction
between wildlife species and their environment.
Historically, the field of ecotoxicology primarily examined exposure and response
of fish, birds, and mammals to environmental contaminants; little was known concerning
the effects of contaminants on reptiles and amphibians (Hall, 1980; Martin, 1983; Hall
and Henry, 1992; Guillette et al , 1995b; Brisbin et al., 1998; Spariing et al., 2000). In
the few cases where contaminant effects on reptiles and amphibians were considered, it
was presumed that tests conducted on fish, birds and mammals would yield a range of
toxicity data that, when evaluated and implemented with appropriate safety factors,
would also protect reptiles and amphibians (Hall and Henry, 1992). Consequentiy, few
studies examining exposure and response of reptiles and amphibians to environmental
contaminants were conducted. Over the last 10 years, however, a substantial increase in
reptile and amphibian ecotoxicological research has occurred. This is largely the result of
heightened concerns over reports of deformities (Ouellet et al., 1997; Rainwater et al.,
1999), mortalities (Schoeb et al., 2002), reproductive abnormalities (Jennings et al., 1988;
Woodward et al., 1993; Guillette et al., 1994, 1996a; Grain et al., 1998) and declining
populations in various species (Gibbons et al., 2000). Subsequently, this surge in reptile
and amphibian ecotoxicology research has demonstrated the sensitivity of these animals
to multiple pollutants (Sparling et al., 2000) and illustrated their utility as biomonitors of
environmental contamination (Lambert, 1997a,b; Grain and Guillette, 1998; Sparling et
al., 2000).
Reptiles are particularly useful species in ecotoxicological research due to various
aspects of their life history and biology. First, reptiles exhibit a wide geographic
distribution and persist in a diversity of aquatic and terrestrial habitats (Halliday and
Adler, 1988; Grain and Guillette, 1998). Second, reptiles are often top camivores in their
respective communities, making them susceptible to bioaccumulation and
biomagnification of chemicals through trophic transfer (Olafsson et al., 1983; Bryan et
al., 1987; Burger, 1992; Hall and Henry, 1992). Third, reptiles are long-lived (Neil,
1971; Halliday and Adler, 1988; Zug et al., 2001), thereby increasing their likelihood of
exposure and accumulation of contaminants. Fourth, reptiles generally have smaller
home ranges (stronger site fidelity) compared to predatory birds or mammals, making
them more precise indicators of contaminant sources (Bauerle et al., 1975; Heinz et al.,
1980; Rootes and Chabreck, 1993; Brisbin et al., 1998). Fifth, some reptiles exhibit
sensitivity to contaminants similar to birds and mammals (Hall and Clark, Jr., 1982) and
exhibit a higher incidence of embryonic mortality and deformity with elevated
contaminant concentrations (Bishop et al., 1991). Lastiy, different reptile species exhibit
variations in certain aspects of their reproduction (e.g., temperature-dependent sex
determination) that make them excellent models for elucidating mechanisms by which
certain contaminants affect the structure and function of the reproductive system (Grain et
al., 1997; Bull et al., 1988; Deeming and Ferguson, 1988, 1989; Janzen and Paukstis,
1991; Lance and Bogart, 1994; Lang and Andrews, 1994; Grain and Guillette, 1998;
Milnes et al., 2002a).
In recent years, one particular group of reptiles, crocodilians (crocodiles,
alligators, caimans, and gharials), has been pushed to the forefront of ecotoxicological
research. Crocodilians function as keystone species in their environments by selectively
preying on fish species, culling diseased or weak animals from their respective
populations, recycling nutrients, and maintaining wet refugia during periods of drought,
thereby shaping the stmcture of associated animal and plant communities (Craighead,
1968; Fittkau, 1970, 1973; Robinson and Bolen, 1984; Thorbjarnarson, 1992; Mazzotti
and Brandt, 1994). As such, adverse effects on crocodilians may have dramatic effects
on the overall systems they support. Although habit loss is currently the most noticeable
threat to crocodilian conservation, exposure to environmental contaminants may present a
subtle yet significant long-term risk to populations in contaminated areas
(Thorbjarnarson, 1992; Gibbons et al., 2000).
The recent emphasis on crocodilians in ecotoxicology is largely the result of
numerous studies showing population declines and reproductive impairment in American
alligators (Alligator mississippiensis) inhabiting contaminated lakes in Florida, USA
(Jennings et al., 1988; Woodward et al., 1993; Guillette et al., 1994, 1996a, 1999a; Grain
et al., 1998a). Over the last 30 years, several reports have documented exposure of wild
crocodilians to various environmental contaminants including organochlorine (OC)
pesticides (BiUings and Phelps, 1972; Best 1973; Ogden et al., 1974; Vermeer et al.,
1974; Wheeler et al., 1977; Hall et al., 1979; Wessels at al., 1980; Matthiessen et al..
1982; Phelps et al., 1986, 1989; Delany et al., 1988; Heinz et al., 1991; Skaare d al.,
1991; Guillette et al., 1999b; Wu d al., 2000a,b; Pepper et al., 2003), polychlorinated
biphenyls (PCBs) (Hall d al., 1979; Phelps d al., 1986; Delany d al., 1988; Heinz et al.,
1991; Cobb d al., 1997, 2002; Bargar d al., 1999; Guillette d al., 1999b), and metals
(Ogden et al., 1974; Vermeer et al., 1974; Stonebumer and Kushlan 1984; Phelps et al.,
1986; Delany et al., 1988; Hord d al., 1990; Ware d al., 1990; Heinz d al., 1991;
Yoshinaga et al., 1992; Ruckel 1993; Heaton-Jones et al., 1994, 1997; Facemire d al.,
1995a; Odierna, 1995; Bowles 1996; Yanochko et al., 1997; Jagoe et al., 1998; Brisbin et
al., 1998; Rhodes 1998; Elsey d al., 1999; Burger et al., 2000; Ding d al., 2001;
Rainwater et al., 2002). The recent studies in Florida, however, are the first to report
reproductive abnormalities and potential population level effects in crocodilians
inhabiting contaminated habitats.
Associated laboratory studies have demonstrated that many of the chemicals to
which wild Florida alligators have been exposed exhibit the ability to dismpt the normal
function of the endocrine system (Vonier et al., 1996; Amold et al., 1997; Grain et al.,
1998b), potentially leading to alterations in reproduction, growth, and survival (Grain et
al., 2000). These endocrine-dismpting chemicals (EDCs) are thought to be at least partly
responsible for reproductive impairment observed in Florida alligators (Woodward et al.,
1993; Grain et al., 1997, 1998a, 2000; Guillette et al., 1994, 1995b, 1996a, 1997, 1999a,
2000; Grain and Guillette, 1997, 1998) and other wildlife, including fish (Johnson et al.,
1988, 1993; Spies and Rice, 1988; Munkittrick et al., 1992, 1994; Hontela et al., 1995;
Lye et al., 1997), birds (Frye and Toone, 1981; Burger et al., 1995), and mammals
(Delong et al., 1973; Gilmartin et al., 1976; Subramanian et al., 1987; Beland et al., 1993;
Facemire et al., 1995b; McCoy et al., 1995; Henny et al., 1996).
Research on EDCs has been conducted for decades but increased dramatically
during the 1990s (Colborn and Clements, 1992; Grain and Guillette, 1997; Kendall et al.,
1998; Guillette and Grain, 2000). Currentiy, EDCs are a major focus of toxicology,
endocrinology, and reproductive physiology research (Grain and Guillette, 1997; Grain et
al., 2000; Guillette, 2000) and the subject of various policy and legislative debates
(Ankley et al., 1998). Numerous chemicals, mostiy anthropogenic but including some
naturally-occurring compounds, have been shown to have endocrine-disrupting properties
(Colborn et al., 1993; Guillette et al., 1996b). These chemicals include a variety of
herbicides, fungicides, insecticides, nematocides, and industrial chemicals (Colbom et al.,
1993; Guillette et al., 1996b) that have the capacity to dismpt normal endocrine function
by altering (1) the hypothalamic-pituitary axis of endocrine control, (2) the activity of
steroidogenic enzymes, (3) the function of steroid binding molecules (plasma proteins),
(4) the activity of steroid hormone receptors by acting as hormone agonists (mimics) or
antagonists (anti-hormones), and (5) the hepatic clearance rate of steroids (Grain and
Guillette, 1997). Such alterations can dismpt the normal production, availability, action,
biotransformation, and excretion of endogenous hormones (see reviews in Grain and
Guillette, 1997; Grain et al., 2000), which in tum can cause a variety of organizational
and activational effects in exposed organisms.
Organizational effects are permanent modifications in the morphology or function
of tissue (e.g., gonads, reproductive ducts, liver) occurring during the period from gamete
production to juvenile development (Guillette et al., 1995a; Grain and Guillette, 1997),
whereas activational effects involve the temporary alteration in the function of normally
organized tissue (temporary insults during mature life stages) (Guillette et al., 1995a;
Grain and Guillette, 1997). That is, organizational effects occur early in an organism's
lifetime and are permanent, while activational effects occur during adulthood and are
usually temporary (Guillette et al., 1995a). Numerous organizational effects including
increased embryonic deformities and mortality, gonadal abnormalities, altered steroid
hormone concentrations, and sex reversal have been observed in various wildlife species
exposed to environmental contaminants (Frye and Toone, 1981; Hose et al., 1989; Bishop
et al., 1991; Bergeron et al., 1994; Guillette et al., 1994; Grain et al., 1997). Likewise,
numerous activational effects such as decreased fertility and fecundity, low clutch
viability, and abnormal mating behavior have also been observed (Frye and Toone, 1981;
Subramanian et al , 1987; Hose et al., 1989; Woodward et al., 1993).
Specific cause-effect relationships between EDCs and reproductive abnormalities
in wildlife remain difficult to discern. Laboratory studies have demonstrated the
sensitivity of developing embryos to chemical signals (Bem, 1992) and confirmed that in
ovo and in utero exposure to EDCs can cause irreversible alterations of the reproductive
systems in multiple wildlife species (Guillette et al., 1995a). Because EDC-induced
organizational effects are initiated during eariy life stages and do not become apparent
until later in life, most field studies have examined activational effects (Guillette et al.,
1995a). However, few large-scale research projects with both laboratory and field
components have been undertaken. As a result, the effects of EDC-induced embryonic
modifications on the health and reproductive potential of adults is still unclear and
difficult to predict (Guillette et al., 2000). Further, population-level effects of EDC
exposure in wildlife have been sparsely studied and are generally unknown (Brisbin et
al., 1998; Guillette et al., 2000). Grain and Guillette (1997, 1998) stressed that studies
with a multi-scale approach (e.g., gene to ecosystem) are needed to adequately examine
and understand the mechanisms and effects of EDCs on wildlife at different levels of
organization.
Studies examining exposure and response of American alligators to
environmental contaminants have provided perhaps the most comprehensive assessment
of EDC effects on a wildlife species to date. A combination of laboratory and field
research over the last two decades has demonstrated the sensitivity of alligators to EDCs
under controlled and field conditions and revealed endocrine dismption and reproductive
abnormalities in these animals at the molecular, cellular, tissue, organism, and population
levels (Grain and Guillette, 1998; Guillette et al., 2000). The primary study site for this
research has been Lake Apopka, a 12,960 ha freshwater lake approximately 20 km west-
northwest of Oriando, Florida, USA (Matter et al., 1998; Guillette et al., 2000). Lake
Apopka is one of the most polluted lakes in Florida as the result of extensive agricultural
pesticide and nutrient mnoff, municipal wastewater discharge, and a major OC pesticide
spill in 1980 (Matter et al., 1998; Guillette et al., 2000). In the five years following the
pesticide spill, significant declines in egg (clutch) viability and juvenile alligator density
were observed on Lake Apopka when compared to other lakes (Woodward et al., 1993).
Egg viability and juvenile recmitment remained depressed until the 1990s, and although
both have since increased, pre-1980 levels have not been observed (Woodward et al.,
1991; Rice et al., 1996). In addition, alligator eggs from Lake Apopka were found to
contain numerous OC pesticides, many of which have been identified as EDCs, at higher
concentrations than eggs from other lakes (Heinz et al., 1991). Laboratory studies later
demonstrated that many of the same contaminants found in Apopka alligator eggs exhibit
an affinity for the alligator estrogen and progesterone receptors (Vonier et al., 1996;
Arnold et al., 1997; Guillette et al., 2002). Additional studies revealed that many EDCs
do not bind to alligator cytosolic binding proteins (vom Saal et al., 1995; Amold et al.,
1996; Grain et al., 1998b), suggesting that these EDCs may go unregulated in the plasma
or cytoplasm, thereby increasing their availability to target cells (Grain and Guillette,
1997; Guillette et al., 2000). Further, Matter et al. (1998) found that some OCs (e.g.,
p,p'-DDE) which have been found in alligator eggs and semm from Lake Apopka (Heinz
et al., 1991; Guillette et al., 1999b) can override the temperature-dependent sex
determination mechanism in crocodilians (Lance and Bogart, 1994; Lang and Andrews,
1994; Lance, 1997) and induce sex reversal (male to female).
During the 1990s and early 2000s, examination of hatchlings and juveniles from
Lake Apopka revealed numerous abnormalities in their reproductive and endocrine
systems when compared to alligators from a reference population. Hatchling and
juvenile males from Lake Apopka exhibited depressed circulating concentrations of
testosterone (T) (Guillette et al., 1994, 1996a, 1997, 1999a; Grain et al., 1998a) and
elevated concentrations of estradiol-17p (E2) (Milnes et al., 2002b), while hatchling
females exhibited elevated circulating concentrations of E2 (Guillette et al., 1994) and
juvenile females exhibited reduced E2 concentrations (Guillette et al., 1999a). In
addition, testes from juvenile Apopka males and ovaries from juvenile Apopka females
exhibited elevated and depressed E2 production, respectively (Guillette et al. 1995b).
Moreover, juvenile Apopka females exhibited abnormal ovarian morphology with
numerous polyovular follicles and polynuclear oocytes, while Apopka males exhibited
poorly organized seminiferous tubules (Guillette et al., 1994). Grain et al. (1997) found
that juvenile Apopka females also had depressed gonadal aromatase (enzyme responsible
for estrogen production; Simpson et al., 1994; Norris, 1996) activity. Abnormal hormone
concentrations during critical early life stages suggest that anatomical structures
dependent on these hormones for proper growth and development may also be altered
(Guillette et al., 2000). Indeed, multiple studies have also shown reduced phallus (penis)
size in juvenile male alligators from Lake Apopka (Guillette et al., 1994, 1996a, 1999a;
Pickford et al., 2000).
American alligators are one of the only wildlife species for which a multi-scale
approach for examining endocrine disruption has been implemented. The vast amount of
data compiled on exposure and response of alligators to EDCs in both the laboratory and
at Lake Apopka illustrates the sensitivity of these reptiles to EDCs and strongly suggests
the potential for contaminant-induced endocrine disruption at various levels of
organization in wild crocodilians inhabiting polluted systems. It is important to note that
in the past few years, many of the reproductive alterations observed in alligators from
Lake Apopka have also been observed in other, lesser contaminated lakes in Florida
(Grain et al., 1998; Guillette et al., 1996a, 2000; Hewitt et al., 2002). Hence,
reproductive abnormalities are not confined to Lake Apopka only, and studies are
currently in progress to examine endpoints of endocrine dismption at other localities in
Florida. But what about other crocodilian species living in contaminated habitats? The
wide distribution of crocodilians in developing, tropical countries where chemical use is
often poorly regulated (Murray, 1994; Thorbjamarson, 1992) suggests a similar EDC
exposure scenario to that observed in Lake Apopka and other Florida wetlands. Are
similar reproductive anomalies manifested in crocodilians inhabiting these areas as well?
Regulations goveming the production, distribution, and use of chemicals in
developing countries are scant or inadequately enforced (Murray, 1994). In Central
America, no training or certification is required for a person to buy or apply pesticides
(Castillo et al., 1997). As a result, large quantities of chemicals are routinely used in the
tropics for agriculture, mining, crop storage, and vector control (Cawich and Roches,
1981; Lacher and Goldstein, 1997; Alegria et al., 2000) at rates often comparable to or
higher than those in developed countries (Castillo et al., 1997). In addition, many
compounds banned in most industriaHzed countries are still commonly used in tropical
areas. For example, the persistent OC (and EDC) DDT is still readily available in many
South Asian countries (Mengech et al., 1995) and is still used for vector control in
Central America (Grieco et al., 2000; Roberts et al., 2002; Alegria et al., 2000).
Chemical storage conditions in many developing countries are also inadequate, further
increasing the potential for environmental contamination (Alegria, 1998). Numerous
environmental contaminants including heavy metals, polycyclicaromatic hydrocarbons
(PAHs), and OCs have been found in sediments in several tropical countries (Hall and
Chang-Yen, 1986; Phuong d al., 1989; Gonzalez, 1991; Bernard, 1995; Gutierrez-
Galindo et al., 1996; Gibbs and Guerra, 1997; Marins et al., 1998; Michel and Zengel,
1998; Norena-Barroso et al, 1998; Carvalho et al., 1999). In addition, elevated
concentrations of multiple OCs have recently been detected in ambient air in Central
America (Alegria et al., 2000). However, despite the wide use and occurrence of these
chemicals in developing countries and the high biodiversity of the tropics (Wilson, 1992),
few studies have examined the exposure and response of tropical wildlife to
environmental contaminants (Goldstein et al., 1996; 1999a,b; Castillo et al., 1997).
In 1994, during a study of the ecology and status of Morelet's crocodile
(Crocodylus moreletii) in Belize (Piatt, 1996), Steven Piatt observed low crocodile egg
viability at a lagoon surrounded by actively-farmed sugarcane fields (Steven Piatt,
personal communication). Morelet's crocodile is a medium-sized freshwater crocodile
found in the Atiantic and Caribbean lowlands of Mexico, Guatemala, and Belize and is
currentiy recognized as an endangered species (Groombridge, 1987; Lee, 1996; Ross,
1998). Piatt pondered the possible influence of agricultural chemicals on crocodile
reproduction in the lagoon and brought this observation to my attention. Shortly
thereafter, Piatt invited Scott McMurry and me to come to Belize and collect non-viable
crocodile eggs for environmental contaminant analysis. In July 1995, McMurry and I
traveled to Belize and with Piatt collected 31 non-viable eggs from crocodile nests at
three localities in the northem portion of the country. Subsequent residue analyses
revealed detectable concentrations of multiple environmental contaminants, including
EDCs, in the eggs. Based on the egg contamination, population declines, and
reproductive abnormalities reported in alligators exposed to many of the same chemicals
in Lake Apopka, the impetus was provided to develop a multi-year project to examine the
ecotoxicology of Morelet's crocodile in Belize. In 1997, with funding from the U.S.
Environmental Protection Agency, the study was initiated. Since that time, our research
team has documented numerous EDCs, particulariy OC pesticides, in a variety of
matrices, including sediment and nest material as well as crocodile eggs, chorioallantoic
membranes, and caudal scutes (Wu, 2000; Wu et al., 2000a,b; DeBusk, 2001; Pepper,
2001; Rainwater et al., 2002; Pepper et al., 2003). The focus of this dissertation was to
further examine the ecotoxicology of Morelet's crocodile in Belize by examining one
additional endpoint of EDC exposure and two additional endpoints of crocodile response
to this exposure in populations residing in contaminated habitats.
The following chapters will describe in detail my approach and assessment of
endpoints of endocrine dismption in Morelet's crocodile. Chapter II describes results of
a study examining vitellogenin induction as an indicator of EDC exposure in male
crocodiles inhabiting contaminated sites in Belize. Vitellogenin is an egg-yolk precursor
protein expressed in all oviparous and ovoviviparous vertebrates (Palmer and Palmer,
1995). Males normally have no detectable level of vitellogenin in their blood, but can
produce it following stimulation by an exogenous estrogen (Palmer and Palmer, 1995),
such as an EDC. Thus, the presence of vitellogenin in the blood of male crocodiles can
serve as an indicator of exposure to an estrogen-mimicking EDC. Chapters III and IV
describe the results of two studies examining crocodile response to EDC exposure.
Chapter III focuses on circulating concentrations of the plasma steroid hormones T and
E2 in crocodiles inhabiting contaminated habitats, while Chapter IV examines phallus size
in male crocodiles from these same sites. Chapter V provides a summary of the
dissertation. Significant findings from the three previous chapters are presented, and data
from this study are compared to those from ecotoxicological studies on alligators in
Florida. Uncertainties associated with this study are then discussed and future research
directions proposed. Lastly, final conclusions are presented.
10
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25
Wu, T.H., T.R. Rainwater, S.G. Piatt, S.T. McMurry and T.A. Anderson. 2000a. Organochlorine contaminants in Morelet's crocodile (Crocodylus moreletii) eggs from Belize. Chemosphere. 40:671-678.
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26
CHAPTER II
PLASMA VITELLOGENIN INDUCTION IN
MORELET'S CROCODILES FROM
CONTAMINATED HABITATS
IN NORTHERN BELIZE
Abstract
Many environmental contaminants exhibit estrogenic activity in wildlife. Such
pollutants may pose a risk or risks to animal populations by disrupting normal
reproductive and developmental processes in exposed individuals. Over the last decade,
population declines and various reproductive and endocrine abnormalities have been
observed in American alligators (Alligator mississippiensis) inhabiting Lake Apopka and
other contaminated lakes in Florida, USA. Alligator eggs and semm from Lake Apopka
contain multiple organochlorine (OC) pesticides known to have an affinity for the
alligator estrogen receptor, suggesting a possible role of these compounds in the observed
reproductive abnormalities. Many of these same contaminants have been detected in
eggs and tissues of other crocodilian species worldwide; however, no studies have yet
examined estrogenic responses in wild crocodilians outside of Florida. Induction of the
egg-yolk precursor protein vitellogenin has been successfully used as a biomarker of
estrogenicity in wildlife, primarily fish. Males normally have no detectable
concentrations of vitellogenin in their blood, but can produce it following stimulation by
exogenous estrogens, such as certain OC pesticides. The presence of vitellogenin in the
blood of males can thus serve as an indicator of exposure to xenobiotic estrogens. In
1995, multiple OC pesticides were found in eggs of the endangered Morelet's crocodile
(Crocodylus moreletii) in Belize. Shortly after, a multi-year study was initiated to
examine exposure and response of these crocodiles to environmental contaminants. The
primary objective of the present study was to examine plasma vitellogenin induction in
crocodiles from contaminated and reference habitats in northem Belize. Of 381
crocodiles examined, 8 animals contained vitellogenin in their plasma. These were adult
females sampled during the breeding season. No males or juvenile females exhibited
27
vitellogenin induction. However, many of the animals sampled were later found to
contain OC pesticides in their caudal scutes, confirming contaminant exposure. The lack
of a vitellogenic response in these animals may be due to several factors including
insufficient contaminant concentrations to induce vitellogenesis or no affinity of these
particular compounds for the Morelet's crocodile estrogen receptor. Plasma vitellogenin
induction may still serve as a reliable biomarker of estrogen exposure in these and other
crocodilians, but the lack of a vitellogenic response should not be interpreted as an
indication of no exposure or a lack of another contaminant-induced biological response.
Introduction
Numerous environmental pollutants are believed to exhibit endocrine disrupting
properties in wildlife (Colbom et al., 1993). Many of these compounds are suspected to
elicit their effects by enhancing or impairing the actions of the natural hormone estrogen,
thereby dismpting normal reproductive and developmental processes (Palmer and
Palmer, 1995; Palmer and Selcer, 1996; Vonier et al., 1996; Grain and Guillette, 1997;
Palmer et al., 1998; Guillette et al., 2002). Over the last decade, laboratory and field
studies examining the effects of xenobiotic estrogens on wildlife have increased
substantially (Colbom and Clements, 1992; Grain and Guillette, 1997; Kendall et al.,
1998; Grain et al., 2000; Guillette and Grain, 2000). AUhough reptiles have received
little attention in ecotoxicological research compared to other vertebrate classes (Hall,
1980; Martin, 1983; Hall and Henry, 1992; Guillette et al., 1995; Brisbin et al., 1998;
Sparling et al., 2000), recent investigations examining exposure and response of
American alligators (Alligator mississippiensis) to environmental contaminants in
Florida, USA have provided one of the most comprehensive assessments of endocrine
dismption in a wildlife species to date (for a review, see Grain and Guillette, 1998;
Guillette et al., 2000). Numerous organochlorine (OC) pesticides considered to be
xenobiotic estrogens have been detected in alligator eggs and semm from Lake Apopka
and other Florida lakes (Heinz et al., 1991; Vonier et al., 1996; Guillette et al., 1999b;
2002). Juvenile alligators from these lakes, particularly lake Apopka, have exhibited
altered steroid hormone concentrations, reduced phallus size, depressed gonadal enzyme
28
expression, and multiple abnormal gonadal morphologies compared to animals from a
reference population (Guillette d al., 1994, 1995, 1996, 1997a, 1999a,b; 2000; Grain d
al., 1997,1998a; Pickford d al., 2000; Hewitt d al., 2002; Milnes d al., 2001; 2002a,b).
Although no direct cause-effect relationship has been established between these
reproductive and endocrine anomalies and environmental contaminants, results of
laboratory and field observations over the last 20 years strongly suggest the potential for
contaminant-induced endocrine disruption at various levels of organization in these
animals (Grain and Guillette, 1998).
Increasing concern over contaminant-induced endocrine dismption in wildlife has
elevated the need for a rapid, sensitive, and inexpensive biomarker of exposure to
environmental estrogens (Palmer and Palmer, 1995). In recent years, induction of the
semm protein vitellogenin has shown promise as such a biomarker, and numerous studies
have investigated its efficacy in both the laboratory and field (Purdom et al., 1994;
Heppell et al., 1995; Palmer and Palmer, 1995; Sumpter and Jobling, 1995; Folmar et al.,
1996; Palmer and Selcer, 1996; Palmer et al., 1998; Allen et al., 1999; Orlando et al.,
1999; Irwin et al., 2001; Selcer at al., 2001; Shelby and Mendonca, 2001; Brasfield et al.,
2002; Hecker et al., 2002; Okoumassoun et al., 2002a,b; Vethaak et al., 2002).
Vitellogenin is the precursor molecule for egg-yolk, expressed in all oviparous
and ovoviviparous vertebrates and essential as the source of metabolic energy for the
developing embryo (Selcer et al., 2001). Although under multi-hormonal control (Ho,
1987; Ho et al., 1982, 1985), vitellogenin production is primarily regulated by estrogen
(Ho, 1987; Palmer and Palmer, 1995; Selcer et al., 2001). Following stimulation of the
hypothalamo-pituitary axis, gonadotropins released by the pituitary stimulate the
production of estrogen in the ovaries (Ho, 1987). Increasing concentrations of estrogen
in tum stimulate the liver to produce vitellogenin, which is released into the bloodstream,
taken up by developing oocytes, and cleaved into egg-yolk proteins (Ho, 1987, Selcer et
al., 2001). Under normal conditions, vitellogenin is present only in mature females at
times corresponding to elevated estrogen concentrations (e.g., reproductive periods)
(Palmer and Selcer, 1996). Conversely, vitellogenin in males and immature animals is
normally non-detectable, due to their normally low endogenous estrogen concentrations
29
(Palmer and Palmer, 1995). However, the liver of males and immature females can
produce vitellogenin in response to exogenous estrogen stimulation (Ho, 1987; Palmer
and Palmer, 1995). Indeed, numerous studies have demonstrated vitellogenin induction
in males exposed to natural, synthetic, and xenobiotic estrogens (for example. Palmer and
Palmer, 1995; Sumpter and Jobling, 1995; Palmer et al., 1998). Thus, the presence of
vitellogenin in males can serve as evidence of exposure to endogenous estrogens or
exogenous estrogens, including OC pesticides and other environmental contaminants
with an affinity for the estrogen receptor (Vonier et al., 1996; Guillette et al., 2002).
Palmer and Selcer (1996) reported that the utility of vitellogenin as a biomarker of
estrogenicity is based on several factors: (1) vitellogenin induction is a physiological
response to estrogen or estrogen-mimicking compounds, (2) the mechanism of
vitellogenin production has been extensively studied and is well-understood, (3)
vitellogenin is readily quantifiable and produced by all non-mammalian vertebrates in a
dose-dependant manner, (4) male oviparous vertebrates normally will not have
vitellogenin in their blood unless they have been exposed to estrogen or chemicals with
estrogenic properties; thus the presence of vitellogenin in the blood of male oviparous
vertebrates can serve as an indicator of exposure to estrogenic contaminants. These
factors meet the necessary criteria for a functional biomarker of xenobiotic estrogen
exposure set forth by Palmer and Palmer (1995). Vitellogenin induction has been shown
to be a particularly reliable biomarker of environmental estrogen exposure in fish
(Purdom et al , 1994; Sumpter and Jobling, 1995; Folmar et al, 1996; Allen et al., 1999;
Oriando et al., 1999; Vethaak et al., 2002; Okoumassoun et al., 2002a,b; Hecker et al.,
2002), while fewer studies have applied this biomarker to other vertebrates (Palmer and
Palmer, 1995; Irwin et al., 2001; Shelby and Mendonca, 2001; Brasfield et al., 2002).
Environmental contaminants, many considered to exhibit estrogenic properties,
have been found in crocodilian eggs and bodily tissues throughout tropical and
subtropical areas worldwide (see Rainwater et al. 2002; Chapter I). However, despite the
apparent sensitivity of alligators to OC pesticides and other pollutants (Guillette et al.,
2000), no study has yet examined endpoints of endocrine dismption in crocodilians
outside of Florida.
30
In 1995, we conducted a pilot study to examine exposure of Morelet's crocodile
(Crocodylus moreletii) to environmental contaminants in Belize. Morelet's crocodile is a
medium-sized, freshwater crocodilian found in the Atlantic and Caribbean lowlands of
Mexico, Guatemala, and Belize (Groombridge, 1987; Lee, 1996; Ross, 1998) and is
currently recognized as endangered under the United States Endangered Species Act
(Endangered and Threatened Wildlife and Plants, 1991). Detectable concentrations of
p,p '-DDE, p,p '-DDT, p,p '-DDD, and heptachlor epoxide were found in crocodile eggs
from three localities in the northern portion of Belize (Rainwater et al., 2002; Rainwater
et al., unpublished data). Based on these findings, a multi-year study was initiated to
examine various endpoints of contaminant exposure and response in Morelet's crocodiles
living in polluted habitats in Belize. This paper describes one component of that study in
which plasma vitelloginin induction was examined in crocodiles from contaminated and
reference habitats to determine exposure to xenobiotic estrogens. We hypothesized that
male crocodiles living in habitats contaminated with estrogenic pollutants would exhibit
plasma vitellogenin induction, while males from reference areas would not. Few studies
have examined contaminant-induced vitellogenesis in wild reptiles (Irwin et al., 2001;
Shelby and Mendonca, 2001), and to our knowledge this is the first study to examine this
endpoint in wild crocodilians.
Materials and Methods
Study Sites
Crocodiles were captured and sampled from two sites in northem Belize, Gold
Button Lagoon and the New River Watershed. Gold Button Lagoon (17°55'N, 88°45'W)
is a large man-made impoundment located on Gold Button Ranch, a 10,526 ha private
cattle ranch approximately 25 km southwest of Orange Walk Town, Orange Walk
District (Figure 2.1). Gold Button Ranch is situated adjacent to an intensively farmed
settiement, and past use of OC pesticides in this area (on crops) as well as on the ranch
itself (in cattie feed and dip) is believed to have occurred (Martin Meadows, pers.
comm.). New River Watershed is comprised of the New River, New River Lagoon, and
associated tributaries in the Orange Walk and Corozal Districts (Figure 2.1). New River
31
Lagoon (17°42'N, 88°38'W; ca. 23 km long) and the southern-most 18 km of New River
constituted the New River Watershed study site for this project. This section of the New
River Watershed is relatively remote, bordered by semi-evergreen seasonal forest
(Stafford, 2000) to the west and seasonally flooded savanna to the east. Both Gold
Button Lagoon and New River Watershed contain two of largest Morelet's crocodile
populations in Belize (Piatt, 1996; Rainwater et al., 1998; Piatt and Thorbjarnarson,
2000). During the 1995 pilot study, multiple OCs and other contaminants were found in
crocodile eggs from Gold Button Lagoon (Rainwater et al., 2002; Rainwater et al.,
unpublished data). Thus, when designing the present study. Gold Button Lagoon was
selected as the contaminated site. Although no samples had been collected from New
River Watershed, based on its remote location, surrounding topography limiting large-
scale agriculture, and logistical advantages, this area was selected as the reference site.
Animal Collections and Sampling
Crocodiles from both sites were hand- or noose-captured at night from a boat
during March-October, 1998-2001 under permit from the Behze Ministry of Natural
Resources. For each animal, sex was determined by cloacal examination of the genitalia
(Allsteadt and Lang, 1995; see Chapter IV) and measures of total length (TL; measured
ventrally), snout-vent length (SVL; measured ventrally from the tip of the snout to the
anterior margin of the cloaca), and mass were obtained. Animals were categorized into
one of the following groups based on size: (1) juvenile males (TL < 179.9 cm), (2)
juvenile females (TL < 159.9 cm), (3) adult females (TL > 150 cm), and (4) adult males
(TL > 180 cm) (Table 2.1). The sizes at which male and female Morelet's crocodiles
become reproductively active (adults) are unknown. Thus, for this study the adult size
class for males was based on that reported for alligators (> 1.8 m; Ferguson, 1985), while
the adult size class for females was based on the smallest known nesting female
Morelet's crocodile on either of our study sites (150 cm TL; Piatt, 1996).
Blood (ca. 1.0 to 10.0 ml, depending on animal size) was collected from the post-
cranial sinus, transferred to an dhylenediaminetetraacetic acid (EDTA)-treated
Vacutainer®, and centrifuged at 2000 rpm for 10 minutes. The plasma supernatant was
32
then transferred to a collection tube and frozen at -25°C until shipment to Texas Tech
University. Samples were then stored at -80°C until assayed for the presence of
vitellogenin using the methods of Selcer et al. (2001). Following sample collection, each
crocodile was marked and released at its site of capture.
Gel Electrophoresis
Plasma samples were electrophoresed under denaturing conditions in
polyacrylamide gels using sodium dodecylsulfate-polyacrylamide gel electrophoresis
(SDS-PAGE) with BioRad 4-15% gradient Tris-HCl gels. Plasma samples were diluted
1:25 with milli-Q water and mixed 1:1 v/v with Laemmli sample buffer (BioRad, 2%
SDS, 62.5 mM Tris-HCl, pH 6.8, 0.01% w/v bromophenol blue, 25% w/v glycerol,
combined with 10% v/v 2-mercaptoethanol) and heated for 4 min in a boiling water bath.
Samples were run at constant current (100 V) in an electrophoresis buffer (BioRad, 25
mM Tris, 192 mM glycine, 0.1% SDS, pH 8.3). Gels were either Coomassie stained
(BioRad, Coomassie Brilliant Blue R-250) and rinsed for 30 min in milli-Q water or used
directly for immunoblotting.
Immunoblotting
Proteins separated by SDS-PAGE were transferred to polyvinylidene fluoride
(PVDF) membranes (BioRad) for immunoblotting. Transfers were performed in a
BioRad Trans-Blot apparatus packed in ice, at 100 V for 2 hr. Transfer buffer was 25
mM Tris, 192 mM glycine, 20% methanol, pH 8.3. Following transfer, the PVDF
membranes were blocked using 5% nonfat dry milk ovemight at 4°C, with shaking.
Membranes were then exposed to antiserum (# 498, rabbit anti-frog [Xenopus laevis]
vitellogenin; generously provided by Dr. K.W. Selcer, Duquesne University) diluted in
5% nonfat dry milk for 2 hr at 37°C, with shaking (Selcer et al., 2001). PVDF
membranes were then washed (10 min, with shaking) once with Tris-tween (50 mM Tris-
HCl (pH 7.5), 0.9% NaCl, 0.05% Tween-20 (BioRad)) and twice with Tris-saline (50
mM Tris-HCl (pH 7.5), 0.9% NaCl). PVDF membranes were then incubated in 5%
nonfat dry milk containing peroxidase-coupled goat anti-rabbit semm (BioRad), diluted
33
1:1000 for 2 hr at 37°C, with shaking. The PVDF membranes were then washed again as
above. Finally, each PVDF membrane was developed with peroxidase-substrate
(diaminobenzidine and urea-hydrogen peroxide tablets, Sigma) until a color change
occurred.
Enzyme-Linked Immunosorbent Assay (ELISA)
Plasma samples were diluted 1:1000 with phosphate buffered saline (PBS;
Dulbecco's, Sigma; pH 7.2), and 100 [i\ were added to individual wells (in triplicate) of
an Immunosorb microliter plate (Nunc-Immuno™, Fisher). PBS alone was added in
triplicate to wells designated as blanks. Antigen was allowed to bind to the plate
ovemight at 4°C in a container filled with water and covered to create a humid chamber.
The plate was then washed five times with PBS. Next, 200 l of PBS-blotto (5 g nonfat
milk in 100 |xl PBS) was used to block for 1 hr at room temperature, with shaking. The
PBS-blotto was replaced with 100 \i\ of polyclonal antisemm (#498, anti-vitellogenin;
Selcer et al., 2001) diluted 1:1000 in PBS-blotto, and the plate was incubated for 2 hr at
room temperature, with shaking. The plate was then washed five times with PBS and
incubated for 2 hr on the shaker at room temperature with goat anti-rabbit
immunoglobulin conjugated to horseradish peroxidase (BioRad) diluted 1:1000 in PBS-
blotto. The plate was again washed five times with PBS, then developed using a
tetramethylbenzidine (TMB) Peroxidase EIA Substrate Kit (BioRad), and incubated for
10 min at room temperature. Optical density was then recorded using a SpectraMax®
microplate reader (Molecular Devices) at 655 nm at 10 min. The reaction was then
stopped with 1 N H2SO4, and the optical density was determined at 450 nm. Results of
the ELISA were first examined qualitatively to determine the presence or absence of
vitellogenin in each of the crocodile groups. If vitellogenin was found in any animals
other than adult females sampled during the breeding season, ELISA resuhs were then
examined quantitatively to determine actual plasma concentrations of the protein(s).
34
Statistical Analyses
All analyses were performed using program JMPin statistical software (Version
3.2, SAS Institute, Gary, NC, USA). Both non-transformed and transformed data did not
meet the assumptions of an ANOVA. Thus, possible significance in plasma vitellogenin
variation (as a function of optical density) in crocodile groups was tested using the non-
parametric Wilcoxon rank sums test (two groups) or the Kruskal-Wallis test (more than
two groups). All statistical tests were considered significant when p < 0.05.
Results
No antiserum specific for vitellogenin in crocodilians is currentiy available, but
antiserum # 498 has exhibited cross-reactivity with vitellogenin in various fishes,
amphibians, and reptiles, including alligators (Selcer et al., 2001). To validate that
antibody # 498 recognized Morelet's crocodile vitellogenin, plasma from 6 adult female
crocodiles sampled during the breeding season (presumptive vitellogenic) and 6 male
crocodiles (4 adults, 2 juveniles; presumptive non-vitellogenic) was electrophoresed
using SDS-PAGE. Two high molecular weight proteins (approximately 204 and 168
kDa) were present in the plasma of all 6 females and absent in all male plasma samples
(Figure 2.2). The Westem blot with the anti-vitellogenin antibody detected these
proteins, confirming them to be vitellogenin (Figure 2.2). These proteins and others
visible in female crocodile plasma on the Western blot may represent two separate forms
of vitellogenin, or the smaller protein may be a breakdown product of the larger, primary
protein (Kyle Selcer, pers. comm.). The visible bands in male crocodile plasma samples
are attributed to non-specific binding.
A total of 381 crocodiles, 294 from New River Watershed and 87 from Gold
Button Lagoon, were examined for plasma vitellogenin induction using ELISA. The
vitellogenin antibody showed high reactivity with plasma samples from 8 adult females,
confirming the presence of vitellogenin in these animals. No vitellogenin was detected in
any of the remaining 373 samples analyzed by the ELISA. Because no vitellogenin was
detected in the 6 male samples using SDS-PAGE and Westem blot analyses, the slight
absorbance observed in these and the remaining non-vitellogenic samples is attributed to
35
background absorbance and minor non-specific binding of the polyclonal antibody (see
Figure 2.2) (Palmer and Palmer, 1995).
As a group, aduh female crocodiles exhibited significant (X^ = 50.42, d.f. = 3,
381, p < 0.0001) vitellogenin induction compared to juvenile females, adult males, and
juvenile males (Figure 2.3). No significant difference in vitellogenin induction was
observed between adult females from New River Watershed and Gold Button Lagoon (X^
= 1.83, d.f. = 1, 23, p = 0.1756) (Figure 2.4).
Discussion
Over the last decade, increasing evidence of contaminant-induced endocrine
disruption in wildlife has highlighted the need for sensitive and reliable assays to screen
populations for exposure to hormone-altering compounds (Colborn and Clements, 1992;
Palmer and Selcer, 1996; Grain and Guillette, 1997; Kendall et al., 1998; Guillette and
Grain, 2000). Vitellogenin induction has shown promise as a sensitive and non-
destmctive biomarker of wildlife exposure to xenobiotic estrogens, particularly in aquatic
systems (Palmer and Palmer, 1995; Sumpter and Jobling, 1995; Folmar et al., 1996;
Purdom et al., 1994). The majority of research examining contaminant-induced
vitellogenesis in wildlife has involved laboratory and in situ studies on fish (for example,
Sumpter and Jobling, 1995; Purdom et al., 1994), and a considerable number of studies
on wild fish have validated the use of this biomarker in the field (Folmar et al., 1996;
Allen et al., 1999; Oriando et al., 1999; Vethaak et al., 2002; Okoumassoun et al ,
2002a,b; Hecker et al., 2002). Comparatively few studies have examined the efficacy of
vitellogenin induction as a biomarker of estrogen exposure in other animals. Vitellogenin
induction has been observed in frogs, turtles, and lizards exposed to estrogenic
compounds in the laboratory (Palmer and Palmer, 1995; Brasfield et al., 2002),
suggesting the utility of this endpoint as a biomarker of environmental estrogen exposure
in wild amphibians and reptiles. However, despite evidence of population declines and
widespread exposure to xenobiotic estrogens in these animals (Gibbons et al., 2000;
Sparling et al., 2000), few studies have examined vitellogenin in wild amphibians and
reptiles living in contaminated habitats (Irwin et al., 2001; Shelby and Mendonca, 2001).
36
This is largely due to a lack of available antibodies specific for vitellogenin in these
animals and the little inter-species cross-reactivity of the antibodies that are available
(Selcer et al., 2001). By using a recentiy-developed polyclonal antibody which,
recognizes vitellogenins in multiple species of fish, reptiles, and amphibians (Selcer et
al., 2001), the present study examined vitellogenin induction in Morelet's crocodiles
inhabiting contaminated and reference habitats in northem Belize.
Of the 381 crocodiles sampled in this study, eight (2%) exhibited vitellogenin
induction. These were all adult females sampled during the peak of the breeding season
(Piatt, 1996; Perez-Higareda et al., 1989; see Chapter 111). Based on previous studies on
crocodilian reproduction, these findings are consistent with those observed for
populations assumed to be exhibiting normal endocrine function (Lance, 1987, 1989;
Kofron, 1990; Guillette et al., 1997b). During this period, ovarian follicles in breeding
females increase in size and secrete estradiol-17p, which in tum stimulates the liver to
produce large amounts of vitellogenin (Lance, 1987). Vitellogenin is then released into
the blood stream and transported to the ovaries where it is absorbed by developing ova
and transformed into yolk (Lance, 1987; 1989). Because vitellogenesis is induced by
estrogen, the presence of vitellogenin in the blood is concomitant with elevated
concentrations of estrogen (Lance, 1987, 1989; Guillette et al., 1997b). This has been
routinely reported in studies of crocodilian reproduction and was also observed in this
study (see Chapter III). That only 35% (8 of 23) of the females sampled in this study
exhibited vitellogenin induction is also consistent with previous observations of
reproductive pattems in other crocodilians. Numerous studies have reported that
significant numbers of females in a given population fail to breed each year (Cott, 1961;
Joanen and McNease, 1980; Wilkinson, 1984; Jacobsen and Kushlan, 1986). The
estimated percentage of non-breeding adult female alligators in various populations in the
southeastem United States has ranged from > 90% to 37% (Joanen and McNease, 1980;
Wilkinson, 1984; Jacobsen and Kushlan, 1986; Lance, 1989; Guillette et al., 1997b). In
addition, Cott (1961) reported that approximately 20% of large (TL > 3 m) female Nile
crocodiles (C. niloticus) in Uganda and northem Zimbabwe (then Rhodesia) fail to breed
each year. Multiple factors including population density, female size and health,
37
available mating and nesting habitat, etc., likely influence the number of breeding
females in a given population. The proportion of vitellogenic females and timing of
vitellogenesis observed in this study are in agreement with reports on other crocodilians
and appear normal.
The fact that vitellogenin induction was not observed in any males (n = 264)
sampled during this study suggests that these and other crocodiles at the two study sites
have likely not been exposed to estrogenic contaminants. However, this is not the case.
We found comparable concentrations of multiple OCs including p,p '-DDE, p,p '-DDT,
and mdhoxychlor in crocodile eggs from both New River Watershed and Gold Button
Lagoon (Wu et al., 2000a), suggesting contaminant exposure in neonates and maternal
females. More definitively, caudal (tail) scutes of crocodiles from both sites were found
to contain p,p '-DDE, p,p '-DDT, p,p '-DDD, and mdhoxychlor (DeBusk, 2001). Of the
animals in this study for which data on scute contamination and vitellogenin induction
are available (n = 83), 73% were exposed to mdhoxychlor, 59% to DDE, 46% to DDT,
and 23% to DDD (DeBusk, 2001). In addition, numerous other OCs including aldrin,
endosulfan I, endrin, heptachlor epoxide, and lindane were found in comparable
concentrations (< 300 ppb) in sediments and crocodile nest material at both sites (Wu et
al., 2000a). Each of these chemicals is believed to have endocrine-dismpting properties
(Colbom et al., 1993), and most have been shown to interact with the alligator estrogen
receptor (Vonier et al., 1996; Guillette et al., 2002).
Following the discovery that New River Watershed and Gold Button Lagoon
exhibited comparable contaminant profiles, considerable effort was made to locate non-
contaminated crocodile habitat to use as a reference site for comparisons of vitellogenin
induction and other ecotoxicological endpoints. Two additional sites in northem Belize
and four additional sites in southem Belize were examined, but all crocodile eggs from
each locality were shown to contain environmental contaminants (Wu et al., 2000b;
Rainwater et al., 2002), suggesting contamination of each site and the crocodiles
inhabiting them. Similar contaminant concentrations were also found in American
crocodile (C. acutus) eggs from four sites in the coastal zone of Belize (Wu et al., 2000b),
further illustrating the ubiquitous nature of environmental contamination in the country.
38
Numerous OC pesticides were at one time commonly sold in Belize for agricultural
purposes (Cawich and Rhodes, 1981), but the amounts and locations in which these
compounds were applied are unknown. DDT was used agriculturally in Belize until
1988, and its use in vector control continues today (Roberts et al., 2002). Our
conversations with locals indicate agricultural use of DDT and other OCs also persists.
Multiple OCs including DDTs, PAHs, and PCBs have been found in sediments in
Chdumal Bay, Mexico (Norena-Barroso et al., 1998), approximately 100 km from New
River Watershed and Gold Button Lagoon, and several toxic metals have been detected in
sediments in Belize City Harbor (Gibbs and Guerra, 1997). In addition, Alegria et al.
(2000) recentiy found elevated concentrations of multiple OC pesticides in air samples
from Belize, suggesting the potential for contamination of this region through
atmospheric deposition. Due to the remote location of New River Watershed and relative
absence of nearby agriculture, atmospheric deposition may be the most significant and
continual source of OC contamination at this site (Eisenreich et al., 1981; Rapaportet al.,
1985; Alegria d al., 2000).
Other studies have also reported a lack of vitellogenin induction in reptiles
exposed to xenobiotic estrogens. Matter et al. (1998) observed no plasma vitellogenin in
juvenile alligators exposed in ovo to various OCs, and concluded that vitellogenin
induction may not be a viable biomarker for alligators of this age, as the biochemical
pathways responsible for vitellogenin synthesis may not be functional in immature
animals. In addition, the contaminant doses administered in this study and subsequent
levels of exposure in neonates may have been insufficient to induce vitellogenesis
(Matter et al., 1998). Amold et al. (2002) reported that topical treatment of alligator eggs
with OC contaminants results in poor contaminant absorption into the yolk. Indeed,
Matter et al. (1998) stressed that the actual chemical dose received by embryos in their
study must be considered lower than the amount applied to the shell, as chemical
transport across the eggshell was incomplete. Thus, it is uncertain if the lack of
vitellogenin induction in neonatal alligators was due to insufficientiy low dose
concentrations or failure of a sufficient dose to fully reach target tissues.
39
In a recent field study, no vitellogenin induction was observed in male painted
turtles (Chrysemvs picta) living in cattle farm ponds containing natural estrogens at
concentrations similar to those observed in streams receiving sewage treatment plant
effluent (Irwin et al., 2001). Based on previous studies demonstrating vitellogenin
induction in male painted turtles injected with high concentrations of estradiol-17p (Ho et
al., 1981; Palmer and Palmer, 1995), Irwin et al. (2001) speculated that adult male turtles
may require a significant previous exposure to estrogen to "prime" their livers to
estrogenic signals. Following this initial exposure, turtles would then respond to
subsequent lower estrogen exposures with greater sensitivity (Irwin et al, 2001; Ho et al,
1985). Thus, the authors surmised that environmentally relevant estrogen concentrations
in the farm ponds were likely not sufficient to induce a vitellogenic response in
unsensitized male turtles (Irwin et al., 2001). This may be the case in the present study as
well. Multiple OCs have been detected in sediments and crocodile tissues from both
New River Watershed and Gold Button Lagoon (Wu et al., 2000a,b; DeBusk, 2001), and
many of these contaminants exhibit an affinity for the alligator estrogen receptor in vitro
(Vonier et al., 1996; Guillette et al., 2002). However, despite exposure to many of these
chemicals, the male crocodiles examined in this study exhibited no vitellogenin
induction, suggesting the concentrations to which these animals are exposed may not be
sufficient to induce a vitellogenic response. It should be noted, however, that OC binding
to the alligator estrogen receptor in vitro does not provide evidence that these compounds
are estrogenic in vivo, nor does it provide evidence that similar OC binding occurs with
the Morelet's crocodile estrogen receptor. Controlled laboratory studies in which
Morelet's crocodiles are dosed with increasing concentrations (singly and in
combination) of contaminants commonly detected in their habitats are needed to
adequately examine the ability of these chemicals to induce vitellogenin production in
this species. However, due to the endangered status of Morelet's crocodile (Ross, 1998),
dosing studies may not be feasible. Thus, inter-species extrapolations based on data from
more commonly studied crocodihans (e.g., American alligators) will continue to provide
the most relevant data on dose-response relationships in these reptiles.
40
Conclusions
This study indicates that Morelet's crocodiles living in contaminated habitats in
northern Belize do not exhibit contaminant-induced vitellogenin induction. Of 381
animals sampled, vitellogenin induction was observed in breeding females only, a pattern
considered indicative of normal crocodilian reproduction. However, caudal scutes from
crocodiles sampled at both sites contained detectable concentrations of OCs known to
have estrogenic properties. The fact that many of these animals have been exposed to
xenobiotic estrogens but do not exhibit vitellogenin induction suggests multiple
possibihties: (1) each of the OCs found in crocodile scutes does not have an affinity for
the Morelet's crocodile estrogen receptor, (2) each OC found in crocodile scutes has the
ability to interact with the Morelet's crocodile estrogen receptor and thus the capacity to
induce vitellogenin, but only at higher concentrations, (3) the specific OCs found in
crocodile scutes have the ability to interact with the Morelet's crocodile estrogen receptor
but act antagonistically on each other, precluding a quantifiable estrogenic effect, (4) the
specific OCs found in the scutes of a given crocodile are indicative of all the OCs to
which that animal is exposed, but exposure has been gradual and at concentrations too
low to induce vitellogenin production; however, mobilization of accumulated OCs
sequestered in other tissues (e.g., fat) over time may introduce into circulation a dose
capable of inducing vitellogenesis, or (5) contaminant profiles in scutes are not indicative
of all the OCs to which a crocodile has been exposed, and other OCs present in other
tissues have an influence (e.g., block the estrogen receptor) on the animal's overall
response to contaminant exposure.
The results of this study agree with recent studies in which male reptiles living in
environments contaminated with natural and xenobiotic estrogens did not exhibit
vitellogenin induction (Irwin et al., 2001; Shelby and Mendonca, 2001). Although not
specifically determined, turtles in these studies were likely exposed to estrogenic
chemicals present in their respective aquatic habitats. Palmer and Palmer (1995) stressed
that as a biomarker, vitellogenin induction demonstrates a biological effect, not simply
the presence of a contaminant in bodily tissues or fluids. The present study supports this
notion, as vitellogenin induction was not observed in male crocodiles, although the
41
presence of estrogenic contaminants in tissues was confirmed analytically for many
animals. However, a lack of vitellogenin induction does not eliminate the possibility of
other biological responses to xenobiotic estrogen exposure. For example, laboratory
studies (Matter et al., 1998a,b; Milnes et al., 2002b) have observed alterations in alligator
sex ratios at p,p '-DDE concentrations insufficient to induce vitellogenin production
(Matter et al., 1998a). Thus, vitellogenin induction in reptiles may serve as a useful
biomarker of exposure to environmental estrogens, but the lack of a vitellogenic response
should not be interpreted as an indication that no exposure or other biological response
has occurred.
42
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50
Table 2.1. Sex, number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for vitellogenin induction during this study.
site
New River Watershed Gold Button Lagoon Group Sex Number Size range (cm TL) Number Si?e range (cm TL)
Adults F 12 156.0-233.0 11 152.0-197.0
M 43 180.1-290.0 8 183.2-298.7
Juveniles F 60 36,0-145.5 34 35.0-148.6
M 179 35.9-176.8 34 35.3-161.0
51
MX? .. /Y Caribbean
BZ Sea
\
17°N-
16»N
Figure 2.1. Map of Behze showing locations of the two study sites. GBL = Gold Button Lagoon, NRW = New River Watershed, NR = New River, NRL = New River Lagoon.
52
Males Females
A. W M M M kv" fM feS ^- * •^ 204 kDa protein
168 kDa protein
SDS-PAGE Gel
B. Protein - antiserum complex
Western Blot
Figure 2.2. SDS-PAGE gel (A) and Western blot (B) of plasma samples from vitellogenic (females) and non-vitellogenic (males) Morelet's crocodiles from northem Belize. The letter "S" indicates the lane containing pre-stained molecular weight standards. In the gel, two large molecular weight proteins were present in all six females (adults; samples collected during the breeding season) and none of the six males (4 adults, 2 juveniles) examined (A). In the Westem blot, both proteins cross-reacted with vitellogenin antiserum (B), confirming both proteins to be vitellogenin. The two proteins may represent different vitellogenin forms, or the lower molecular weight protein may be a breakdown product of the larger protein.
53
0.20
I 0.15 O
(0 0.10 c 0) •o
75 _o Q. 0.05
O
0.00
p < 0.0001
a
-
-
-
•
2 3
-
b h _ b
94 51 213*-
-
•
AF JF AM
Group
JM
Figure 2.3. Vitellogenin induction (as a function of optical density at 450 nm) in plasma of Morelet's crocodiles from northem Behze. Numbers inside bars indicate the number of animals sampled within that group. Bars with different superscripts are significantly different. Only plasma from adult females contained vitellogenin (also see results of a gel electrophoresis and immunoblotting in Figure 2.2). AF = aduh females; JF = juvenile females; AM = adult males; JM = juvenile males.
54
0.25
Figure 2.4. Vitellogenin induction in the plasma of Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. Numbers above bars indicate the number of animals sampled per site within a group. Vitellogenin was only detected in the plasma of adult females. No significant difference in vitellogenin induction (adult females) and background absorbance (remaining groups) was observed between sites. AF = adult females; JF = juvenile females; AM = aduh males; JM = juvenile males.
55
CHAPTER 111
SEX-STEROID HORMONE CONCENTRATIONS IN
MORELET'S CROCODILES FROM
CONTAMINATED HABITATS
IN NORTHERN BELIZE
Abstract
Numerous studies have reported altered concentrations of the sex-steroid
hormones estradiol-17p (E2) and testosterone (T) in American alligators (Alligator
mississippiensis) inhabiting contaminated lakes in Florida, USA. However, despite the
apparent sensitivity of alligators to endocrine-dismpting contaminants (EDCs), no studies
have examined these endpoints in other crocodilians living in contaminated habitats. The
primary objective of this study was to examine plasma E2 and T concentrations in
Morelet's crocodiles (Crocodylus moreletii) from contaminated and reference sites in
northem Belize. Data were first examined by comparing hormone concentrations among
males and females within different size groups (small juveniles, large juveniles, adults)
from the contaminated site. Gold Button Lagoon, and the reference site. New River
Watershed. No significant (p < 0.05) differences in plasma E2 concentrations were
detected between sites. Large juvenile males and females from Gold Button Lagoon
exhibited significantiy (p < 0.05) reduced plasma T concentrations compared to large
juveniles males and females from the New River Watershed, respectively. No other
inter-site differences in hormone concentrations were observed. Data were then
examined for relationships between body size and hormone concentrations within each
size group. Significant body size-E2 relationships were observed in large juvenile
females from New River Watershed (r = 0.26 , p = 0.03) and aduh males (r = 0.93, p =
0.04) and females (r = 0.50, p = 0.05) from Gold Button Lagoon. Body size was
positively related to T in Gold Button Lagoon females only (r = 0.77, p = 0.003). The
overall similarity in hormone concentrations and body size-hormone relationships
between sites might be explained by the discovery midway through the study that New
River Watershed exhibits a contamination profile closely resembling that of Gold Button
56
Lagoon. Due to the lack of a legitimate reference site, it is unclear whether steroid
hormone concentrations observed at these two sites are normal or altered by some
stressor (e.g., EDCs). Thus, the biological significance of the few site differences in
hormone concentrations observed in this study is difficult to interpret. An inability to
locate a non-contaminated reference site after sampling multiple localities throughout the
country underscores the ubiquitous nature of environmental contamination in Belize and
demonstrates the inherent difficulty of acquiring suitable reference sites for
ecotoxicological field studies. These difficulties may be more pronounced in developing
countries where chemical use is often unregulated and information on environmental
contamination non-existent.
Introduction
Over the past two decades, studies examining exposure and response of
American alligators (Alligator mississippiensis) to endocrine-dismpting contaminants
(EDCs) have provided one of the most comprehensive ecotoxicological assessments on a
wildhfe species to date (for a review, see Grain and Guillette, 1998; Guillette et al.,
2000). A combination of laboratory and field research has demonstrated exposure and
sensitivity of alligators to EDCs and revealed endocrine dismption and reproductive
abnormalities in these animals at multiple levels of organization (Grain and Guillette,
1998; Guillette et al., 2000). The primary study site for this research has been Lake
Apopka, a large freshwater lake in central Florida, USA (Matter et al., 1998a; Guillette et
al., 2000). Lake Apopka is one of the most polluted lakes in Florida as the result of
extensive agricultural pesticide and nutrient runoff, municipal wastewater discharge, and
a major organochlorine (OC) pesticide spill in 1980 (Matter et al., 1998a; Guillette et al.,
2000). In the five years following the pesticide spill, significant declines in egg (clutch)
viability and juvenile alligator density were observed on Lake Apopka when compared to
other lakes (Jennings et al., 1988; Woodward et al , 1993). Egg viabihty and juvenile
recmitment remained depressed until the 1990s, and although both have since increased,
pre-1980 levels have not been observed (Woodward et al., 1991; Rice et al , 1996). In
addition, alligator eggs from Lake Apopka were found to contain numerous OC
57
pesticides, many of which have been identified as EDCs, at higher concentrations than
eggs from other lakes (Heinz et al., 1991). Laboratory studies later demonstrated that
many of the same contaminants found in Apopka alligator eggs and serum exhibit an
affinity for the alligator estrogen and progesterone receptors (Vonier et al., 1996; Amold
et al., 1997; Guillette et al., 2002). Additional studies revealed that many EDCs do not
bind to alligator cytosolic binding proteins (vom Saal et al., 1995; Amold et al., 1996;
Grain et al., 1998b), suggesting that these EDCs may go unregulated in the plasma or
cytoplasm, thereby increasing their availability to target cells (Grain and Guillette, 1997;
Guillette et al., 2000). Further, Matter et al. (1998a) found that some OCs (e.g., p,p'-
DDE) which have been found in alligator eggs and serum from Lake Apopka (Heinz et
al., 1991; Guillette et al., 1999b) can override the temperature-dependent sex
determination mechanism in crocodilians (Lance and Bogart, 1994; Lang and Andrews,
1994; Lance, 1997) and induce sex reversal (male to female).
During the 1990s and early 2000s, examination of hatchlings and juvenile
alligators from Lake Apopka revealed numerous abnormalities in their reproductive and
endocrine systems when compared to alligators from a reference population. Hatchling
and juvenile males from Lake Apopka exhibited depressed circulating concentrations of
testosterone (T) (Guillette et al., 1994, 1996, 1997a, 1999a; Grain et al., 1998a) and
elevated concentrations of estradiol-17p (E2) (Milnes et al., 2002), while hatchling
females exhibited elevated circulating concentrations of E2 (Guillette et al., 1994) and
juvenile females exhibited reduced E2 concentrations (Guillette et al., 1999a). In
addition, testes from juvenile Apopka males and ovaries from juvenile Apopka females
exhibited elevated and depressed E2 production, respectively (Guillette et al., 1995).
Moreover, juvenile Apopka females exhibited abnormal ovarian morphology with
numerous polyovular follicles and polynuclear oocytes, while Apopka males exhibited
poorly organized seminiferous tubules (Guillette et al., 1994). Grain et al. (1997) found
that juvenile Apopka females also exhibited depressed activity of gonadal aromatase, the
enzyme responsible for estrogen production. Abnormal hormone concentrations during
critical early life stages suggest that anatomical structures dependent on these hormones
for proper growth and development may also be altered (Guillette et al., 2000). Indeed,
58
multiple studies have also shown reduced phallus (penis) size in juvenile male alligators
from Lake Apopka (Guillette et al., 1994, 1996, 1999a,b; Pickford d al., 2000).
Although no direct cause-effect relationship has been established between the
reproductive abnormalities observed in Apopka alligators and environmental
contaminants in the lake, results of laboratory and field observations over the last 20
years strongly suggest the potential for contaminant-induced endocrine dismption at
various levels of organization in these animals (Grain and Guillette, 1998).
Recentiy, many of the reproductive alterations observed in alligators from Lake
Apopka have also been observed in other, lesser contaminated lakes in Florida (Grain et
al., 1998; Guillette et al., 1996, 2000; Hewitt et al., 2002), illustrating that reproductive
abnormalities are not confined to Lake Apopka only, and contaminant-induced endocrine
disruption in other wild crocodilians inhabiting polluted systems may occur. However,
despite many reports of environmental contaminant exposure in other crocodihan species
worldwide (see Rainwater et al., 2002; Chapter I), no other studies have examined
endpoints of endocrine dismption in crocodilians outside of Florida.
Regulations goveming the production, distribution, and use of chemicals in
developing countries are scant or inadequately enforced (Murray, 1994), increasing the
potential for environmental contamination and subsequent contaminant exposure in
wildlife. In much of Central America, no training or certification is required for a person
to buy or apply pesticides (Castillo et al., 1997). As a result, large quantities of chemicals
are routinely used in the tropics for agriculture, mining, crop storage, and vector control
(Cawich and Roches, 1981; Lacher and Goldstein, 1997) at rates often comparable to or
higher than those in developed countries (Castillo et al., 1997). In addition, many
compounds banned in most industrialized countries are still commonly used in tropical
areas. For example, the persistent OC (and EDC) DDT is still easily available in many
South Asian countries (Mengech et al., 1995) and is still used for vector control in
Central America (Grieco et al., 2000; Roberts et al., 2002). In addition, chemical storage
conditions in many developing countries are often inadequate, further increasing the
potential for environmental contamination (Alegria, 1998). Numerous environmental
contaminants including heavy metals, polycyclicaromatic hydrocarbons (PAHs), and OCs
59
have been found in sediments in several tropical countries (Hall and Chang-Yen, 1986;
Phuong d al., 1989; Gonzalez, 1991; Bemard, 1995; Gutierrez-Galindo d al., 1996;
Gibbs and Guerra, 1997; Marins et al., 1998; Michel and Zengel, 1998; Norena-Barroso
et al., 1998; Carvalho et al., 1999). However, despite the wide use and occurrence of
these chemicals in developing countries and the high biodiversity of the tropics (Wilson,
1992), few studies have examined the exposure and response of tropical wildlife to
environmental contaminants (Goldstein et al., 1996, 1999a,b; Castillo et al., 1997).
The majority of the 23 recognized extant species of crocodilians occur in tropical
regions within the boundaries of developing countries (Ross, 1998), suggesting the
potential for contaminant exposure in these animals. In 1995, we found detectable
concentrations of numerous contaminants, including multiple chemicals considered to be
EDCs. in non-viable Morelet's crocodile (Crocodylus moreletii) eggs from three
localities in northem Belize (Rainwater et al., 2002; Rainwater et al., unpublished data).
Morelet's crocodile is a freshwater crocodilian found in the Atlantic and Caribbean
lowlands of Mexico, Guatemala, and Belize (Groombridge, 1987; Lee, 1996; Ross, 1998)
and is currently recognized as endangered under the United States Endangered Species
Act (Endangered and Threatened Wildlife and Plants, 1991). Based on these findings
and previous data from Lake Apopka showing egg contamination, population declines,
and reproductive abnormalities in alligators exposed to many of the same chemicals, a
multi-year study was initiated to examine various endpoints of contaminant exposure and
response in Morelet's crocodiles living on contaminated and reference sites in northem
Belize. This paper describes one component of that study in which plasma
concentrations of ET and T were examined in juvenile and adult crocodiles from
contaminated and reference habitats. Circulating concentrations of E2 and T have been
found to be consistentiy different in similar-sized alligators from contaminated and
reference lakes in Florida (Guillette et al., 1994, 1996, 1997a, 1999a,b; Grain et al.,
1998a; Milnes et al, 2002), suggesting disruption of normal endocrine function in
animals exposed to environmental contaminants. Based on these observations, we
hypothesized in this study that crocodiles living in contaminated habitats in Belize would
60
also exhibit altered hormone concentrations compared to crocodiles from non- or less-
contaminated lagoons.
Materials and Methods
Study Sites and Sample Collection
Crocodiles were captured and samples collected from two sites in northern Belize,
Gold Button Lagoon and the New River Watershed. Gold Button Lagoon (17°55'N,
88°45'W) is a large man-made lagoon located on Gold Button Ranch, a 10,526 ha private
cattie ranch approximately 25 km southwest of Orange Walk Town, Orange Walk
District (Figure 3.1). Gold Button Ranch is situated adjacent to an intensively farmed
settlement, and past use of OC pesticides in this area (on crops) as well as on the ranch
itself (in cattle feed and dip) is believed to have occurred (Martin Meadows, pers.
corrmi.). New River Watershed is comprised of the New River, New River Lagoon, and
associated tributaries in the Orange Walk and Corozal Districts (Figure 3.1). New River
Lagoon (17°42'N, 88°38'W; ca. 23 km long) and the southem-most 18 km of New River
constituted the New River Watershed study site for this project. This section of New
River Watershed is relatively remote, bordered by semi-evergreen seasonal forest
(Stafford, 2000) to the west and seasonally flooded savanna to the east. Both Gold
Button Lagoon and New River Watershed contain two of the largest Morelet's crocodile
populations in Belize (Piatt, 1996; Rainwater et al., 1998; Piatt and Thorbjarnarson,
2000). During a pilot study to examine exposure of Morelet's crocodiles to
environmental contaminants in Belize, we found p,p '-DDE, p,p '-DDT, p,p '-DDD,
heptachlor epoxide, and mercury in crocodile eggs from Gold Button Lagoon (Rainwater
et al., 2002; Rainwater et al., unpublished data). Thus, when designing the present study.
Gold Button Lagoon was selected as the contaminated site. Although no samples had
been collected from New River Watershed, based on its remote location, surrounding
topography limiting large-scale agriculture, and logistical advantages, this area was
selected as the reference site.
Crocodiles from both sites were hand- or noose-captured at night from a boat
under permit from the Belize Ministry of Natural Resources. To minimize temporal
61
effects of sampling, animals were collected during the same two-month period (1 April to
29 May) for four consecutive years (1998-2001). This period corresponds to the peak of
the breeding season for Morelet's crocodile in northem Belize (Piatt, 1996; see Chapter
II). For each animal, sex was determined by cloacal examination of the genitaha
(Allsteadt and Lang, 1995; see Chapter IV) and measurements (total length [TL;
measured ventrally], snout-vent length [SVL; measured ventrally from the tip of the
snout to the anterior margin of the cloaca], mass) obtained. Animals were categorized
into one of the following groups based on size: (1) small juvenile males (TL < 80 cm),
(2) small juvenile females (TL < 80 cm), (3) large juvenile females (TL = 80-149.9 cm),
(4) large juvenile males (TL = 80-179.9 cm), (5) aduh females (TL > 150 cm), and (6)
adult males (TL > 180 cm). Separation of juveniles into small and large groups was
necessary because hormone concentrations in smaller juveniles are likely to exhibit a
higher degree of variation compared to those in larger juveniles. Prior to this study, no
data were available concerning hormone concentrations in Morelet's crocodiles.
However, studies on juvenile alligators have demonstrated that relationships between
body size and plasma hormone concentrations are not observed until animals reach
approximately 80 cm TL (Guillette et al., 1996, 1999a). In addition, the sizes at which
male and female Morelet's crocodiles become reproductively active (adults) are
unknown. Thus, for this study, the aduh size class for males was based on that reported
for alligators (> 1.8 m; Ferguson, 1985), while the aduh size class for females was based
on the smallest known nesting female Morelet's crocodile on either of our study sites
(150 cm TL; Piatt, 1996). Blood (volume relative to animal size but not exceeding 1.6%
body mass) was collected from the post-cranial sinus, transferred to an EDTA-treated
Vacutainer®, placed on ice in the field, and later centrifuged at 2000 rpm for 10 minutes.
The plasma supematant was then transferred to a collection tube and frozen at -25°C until
to shipment to Texas Tech University. Samples were then stored at -80°C until assayed
for E2 and T concentrations. Following sample collection, each crocodile was marked
and released at its site of capture.
62
Steroid Hormone Radioimmunoassays
E2 and T were analyzed using radioimmunoassays previously validated for
alligator plasma (Guillette et al., 1997a) and modified by Grain et al. (1997). Each
plasma sample (250 \i\ for E2, 40 ^1 for T) was extracted 2X with ethyl ether. Briefly,
plasma was mixed with 4 ml ether for 1 min. The aqueous layer was frozen in a dry ice-
methanol bath (-30°C) and the ether supematant decanted into a borosilicate glass assay
tube. The ether extract (supernatant) was then dried using a vortex evaporator
(Labconco, Lenexa, KS) for 15 min. The aqueous pellet was re-extracted with ether and
this second ether extract added to the assay tube. The ether extract was then dried by
vortex evaporation. Extraction efficiency averaged 63% for E2 and 71% for T, and the
assay had a linear range of at least two orders of magnitude (1.56 - 800 pg hormone/100
\i\ borate buffer) (Figure 3.2). Dried samples were resuspended with borate buffer (100
\i\; 0.5 M; pH 8.0). To reduce nonspecific binding, 100 ^1 of borate buffer with bovine
semm albumin (BSA, fraction V; Fisher Scientific) at a final assay concentration of
0.15% for T and 0.19% for E2 was added to each tube. Antibody was then added to each
tube (200 ^1; final concentration of 1:25,000 for T, 1:55,000 for E2). Antibody for all T
samples and for large juvenile and adult E2 samples was obtained from Endocrine
Sciences, Calabasas Hills, CA, USA. Endocrine Sciences stopped producing its E2
antibody before our small juvenile E2 samples were analyzed; thus these samples were
analyzed using antibody from ICN Biomedicals, Costa Mesa, CA, USA. Cross
reactivities of the T antibody to other ligands are as follows: dihydrotestosterone, 44%;
A-1-testosterone, 41%; A-1-dihydrotestosterone, 18%; 5 a-androstan-3p, 17p-diol, 3%; 4-
androsten-3p, 17p-diol, 2.5%; A-4-androstenedione, 2%; 5p-androstan-3p, 17p-diol,
1.5%; estradiol, 0.5%; all other ligands <0.2%. Cross reactivities of the E2 antibody used
for large juveniles and adults to other ligands are as follows: estrone, 1.3%; estriol, 0.6%;
16-keto-estriol, 0.2%; all other ligands <0.2%. Cross reactivities of the E2 antibody used
for small juveniles to other ligands are as follows: estrone, 1.3%; estriol, 2.5%; estradiol
17a, 1.3%; estrone-sulfate, 0.2%; ethinyl estradiol, 0.1%; ah other hgands <0.01%.
Finally, 100 \il of radiolabeled steroid was added to each tube (12,000 cpm per 100 \x\;
[2,4,6,7,16,17-^H]estradiol @ 1 mCi/ml; [l,2,6,7-^H]testosterone @ 1 mCi/ml; both from
63
Amersham International, Piscataway, NJ, USA). For standard tubes, either T or E2 was
added at 0, 1.56. 3.13, 6.25, 12.5, 25, 50, 100, 200, 400, and 800 pg/tube. All tubes were
vortexed for 1 min and incubated overnight at 4°C. All standards and samples were
prepared in duplicate. Separation of bound and free hormone was accomplished by
adding 500 \i\ 5% charcoal-0.5% dextran and immediately centrifuging at 2500 rpm at
4°C for 30 min. Lastiy, 500 \i\ of the supernatant was added to 5 ml scintillation cocktail,
and the tubes were counted on a Beckman LS 6500 scintillation counter. Interassay
variance was 22% for T, 25% for large juvenile and aduU E2, and 28% for small juvenile
E2. Intraassay variance was 4.70% for T, 4.69% for large juvenile and adult E2, and
3.79% for small juvenile E2. Samples were analyzed by size group and were randomly
selected (both sexes) for each assay.
Statistical Analyses
All statistical tests were performed using program JMPin statistical software
(Version 3.2, SAS Institute, Gary, NC, USA). Data were log- or square root-transformed
to obtain normality and homogeneity of variance (the latter checked using Levene's test).
Non-parametric Wilcoxon rank sums test (two groups) or the Kmskal-Wallis test (more
than two groups) was used if an assumption of the ANOVA was not met following
transformation. The majority of the data were not normally distributed; so to maintain
consistency, all data were analyzed using non-parametric statistics. For each crocodile
size group, samples falling below detection limits were assigned concentrations of 0.5X
the minimum detected value for that group (Milnes et al., 2002). To examine the
relationship of body size and hormone concentration, linear regression analysis was
conducted for each group at each site. All statistical tests were considered significant
when p < 0.05.
64
Results
Mean Body Size
For each group of animals, no significant differences in mean hormone
concentrations were observed by month (April, May) or among years (1998-2001). Thus,
data within each group were pooled for further analysis. Because steroid hormone
concentrations are influenced by body size in crocodilians (Lance, 1989; Guillette et al.,
1996, 1999a; Grain et al., 1998a), mean crocodile body size was then examined for
differences between sites within each size group. For animals sampled for plasma E2
analyses (n = 180; Table 3.1), mean body size (TL) was not significantiy different
between sites for large juvenile males (X^ = 0.03; d.f. = 1, 54; p = 0.87), large juvenile
females (X^= 1.48; d.f. = 1, 31; p = 0.22), aduh males (X^ = 0.72; d.f. = 1, 12; p = 0.40),
and adult females (X" = 0.02; d.f. = 1, 13; p = 0.88). However, differences in mean body
size were observed in small juveniles, with males (X^ = 14.32; d.f. = 1, 48; p = 0.0002)
and females (X^ = 11.76; d.f. = 1, 22; p = 0.0006) from New River Watershed being
larger than those from Gold Button Lagoon. Similarly, for animals sampled for plasma T
analyses (n = 188; Table 3.2), mean body size was not significantly different for large
juvenile males (X^ = 0.61; d.f. = 1, 56; p = 0.44), large juvenile females (X^ = 1.17; d.f. =
1, 31; p = 0.28), aduh males (X^ = 0.72; d.f. = 1, 12; p = 0.40), and aduh females (X^ =
0.02; d.f. = 1, 13; p = 0.88). But again, small juveniles from New River Watershed were
th
females: X^ = 9.26; d.f. = 1, 24; p = < 0.002).
larger than those from Gold Button Lagoon (males: X^ = 11.08; d.f. = 1, 52; p < 0.0009;
Mean Hormone Concentrations
Following body size comparisons, data were then examined for site differences in
mean hormone concentrations within each group and for differences among groups
within the same site (Table 3.3). Because a different antibody was used to analyze E2 in
small juveniles, data from these animals were not compared to those collected for large
juveniles and adults. Different antibodies have different cross-reactivities with other
steroids, and using multiple antibodies could resuh in differences in the magnitude of the
steroid measured (McMaster et al., 2001). Between sites, no differences in mean E2
65
concentrations were observed in small juvenile males (X^ = 0.004; d.f. = 1, 48; p = 0.95),
small juvenile females (X' = 2.30; d.f. = 1, 22; p = 0.13), large juvenile males (X^ = 3.59;
d.f. = 1, 54; p = 0.06), large juvenile females (X^ = 0.95; d.f. = 1, 31; p = 0.33), aduh
males (X' = 3.54; d.f. = 1, 12; p = 0.06), and adult females (X^ = 0.77; d.f. = 1, 13; p =
0.38) (Figure 3.3). Within sites, mean E2 concentrations were significantly higher in
adult females than adult males and large juveniles at both New River Watershed (X^ =
35.23; d.f. = 3, 79; p < 0.0001) and Gold Button Lagoon (X^ = 19.19; d.f. = 3, 31; p =
0.0002) (Figure 3.4). No difference was observed in mean E2 concentrations between
small juveniles at either site (New River Watershed: X^ = 0.02; df = 1, 42; p = 0.88;
Gold Button Lagoon: X^ = 2.14; d.f. = 1, 28; p = 0.14).
Mean T concentrations were not different between small juvenile males (X^ =
0.68; d.f. = 1, 52; p = 0.41), small juvenile females (X^ = 0.34; d.f. = 1, 24; p = 0.56),
aduh males (X^ = 2.34; d.f. = 1, 12; p = 0.13), and aduh females (X^ = 0.02; d.f. = 1, 13;
p = 0.88) from the two sites (Figure 3.5). However, T concentrations were significantly
higher in large juveniles from New River Watershed compared to large juveniles from
Gold Button Lagoon (males: X^ = 10.61; d.f. = 1, 56; p = 0.001; females: X^ = 5.77; d.f.
= 1, 31; p = 0.02) (Figure 3.5). Within sites, T concentrations at New River Watershed
were significantly higher in adult males than in all other groups (X^ = 38.15; d.f. = 5,
125; p < 0.0001) (Figure 3.6). However, at Gold Button Lagoon, T concentrations in
adult males were significantly higher than large juvenile females only (X = IQ.ll; d.f. =
5, 63; p = 0.0009) (Figure 3.6).
Body Size and Hormone Concentrations
The relationship between body size and hormone concentrations in crocodiles
from New River Watershed and Gold Button Lagoon was examined using regression
analysis (Table 3.4). For small juveniles, no relationship between body size and E2
(Figure 3.7) or T (Figure 3.8) was observed at either site. For large juveniles, a positive
relationship between body size and E2 was apparent in females from New River
Watershed (r = 0.26, p = 0.03) but not in females from Gold Button Lagoon (r = 0.20, p
= 0.15) (Figure 3.9). No body size- E2 relationship was observed in males from either
66
site. In addition, T was not positively related to body size in large juveniles of either sex
at either site (Figure 3.10). For adults, a positive relationship between body size and E2
was detected in males (r = 0.93, p = 0.04) and females (r = 0.50, p = 0.05) from Gold
Button Lagoon but not from New River Watershed (Figure 3.11). Body size was
positively related to T in Gold Button Lagoon females (r = 0.77, p = 0.003), but no body
size-T relationship was detected in Gold Button Lagoon males or in either sex from New
River Watershed (Figure 3.12).
Discussion
Few inter-site differences in steroid hormone concentrations were observed in
crocodiles from New River Watershed and Gold Button Lagoon. No differences in E2
concentrations were observed between sites for any of the crocodile groups examined,
and no differences in T concentrations were observed in small juveniles and adults
between the two sites. The only inter-site differences in steroid hormone concentrations
observed were T concentrations in large juveniles, which were higher in animals from
New River Watershed than those from Gold Button Lagoon. Grain et al. (1998a) also
found no differences in plasma E2 concentrations, but reported significant differences in
T concentrations in juvenile alligators from contaminated and reference lakes in Florida.
Numerous studies have documented reduced plasma T concentrations in juvenile
alligators from contaminated lakes (Guillette et al., 1994, 1996, 1997a, 1999a,b; Grain et
al., 1998a), suggesting a possible influence of contaminants on circulating T
concentrations. No correlation was observed between serum hormone concentrations and
semm contaminant concentrations in alligators from these localities, leading researchers
to speculate that reproductive abnormalities (e.g., alterations in hormone concentrations,
gonadal morphology, and enzyme expression) previously observed on the more
contaminated lakes (particularly Lake Apopka) could be due to chemical exposure during
embryonic development rather than more recent exposures (Guillette et al., 1999b).
It is important to note that in addition to the potential influence of environmental
contaminants, significant variation in T concentrations may exist naturally among
populations of similar-sized crocodihans (Lance and Elsey, 1986; Guillette et al., 1999a).
67
Lance and Elsey (1986) reported inter-site variation in plasma T in male alligators (TL =
140-380 cm) sampled from North Carolina, South Carolina, Florida, and Louisiana,
USA. In addition, Guillette et al. (1999a) found T concentrations in juvenile (TL = 80-
130 cm) male alligators from one Florida lake to be almost two orders of magnitude
higher than concentrations in males from six other lakes and speculated that the variation
in hormone concentrations among sites could be, in part, related to differences in body
size, seasonal variation in plasma T in individuals, and differing hormonal cycles by site.
In the present study, differences in crocodile T concentrations between large juvenile
crocodiles from New River Watershed and Gold Button Lagoon are likely not related to
body size, as no inter-site differences in body size in this group was detected. Only small
juveniles exhibited significant differences in body size between sites in this study, and T
concentrations in these groups were not significantiy different. The source of inter-site
variation in T concentrations observed in this study are unknown, but may be a function
of natural variation in T concentrations between sites, exposure to one or more
environmental contaminants, one or more undetermined factors (e.g., stress), or a
combination of these.
Intra-site differences in E2 and T concentrations among crocodiles from New River
Watershed and Gold Button Lagoon were also observed. Adult females exhibited
significantly higher E2 concentrations than all other groups at both sites, and adult males
at New River Watershed exhibited significantly higher T concentrations than all other
groups from that site. These results are consistent with other studies which have
demonstrated that concentrations of E2 and T peak during the breeding season in adult
female and male crocodilians, respectively, while hormone concentrations in immature or
non-breeding animals remain comparatively low (Lance, 1987, 1989; Kofron, 1990;
Guillette et al., 1997b). Concentrations of T in adult males at Gold Button Lagoon were
significantiy higher than those of large juvenile females but not any other group within
the site. This is likely explained by the small number of adult males (n = 4) sampled at
Gold Button Lagoon. In addition, four small juvenile males exhibited T values
approximately 2.5- to 4.5-fold higher than the rest of the individuals in that group (n =
18). When those four animals are removed from the analysis, T concentrations in adult
68
males are significantly higher than all other groups except adult females (X^ ^ 16.41; d.f.
= 5, 59; p = 0.0058), and concentrations in large juvenile females are no longer
significantly different from other juvenile groups. Overall, relative concentrations of E2
and T (e.g., higher concentrations in adults compared to juveniles) in crocodiles within
each site resemble those observed in other crocodilian species sampled during the
breeding season (Lance, 1987, 1989; Kofron, 1990; Guillette et al., 1997b).
Relationships between body size and hormone concentrations varied by sex, site,
and the steroid hormone examined. The lack of a relationship between body size and T
in small juvenile crocodiles (TL < 80 cm) supports previous reports in which a body size-
T relationship was not found in alligators less than 75-80 cm TL (Grain et al., 1998a;
Guillette et al., 1996). This finding further supports our reasoning for dividing juvenile
crocodiles into two size groups. The variabihty in body size-hormone relationships
observed in adults may be partly due to the small number of males and females sampled
at both sites. In addition, not all adult female crocodilians breed every year, and non-
breeding animals exhibit markedly lower steroid hormone concentrations than breeding
individuals (Lance, 1987, 1989; Kofron, 1990; Guillette et al., 1997b). Based on analysis
of plasma vitellogenin, not all adult female crocodiles sampled in this study were in
breeding condition (see Chapter II). Thus, the variability in hormone concentrations
between breeding and non-breeding adult females likely influenced the body size-
hormone relationship. For large juveniles, a positive relationship between body size and
E2 was observed in females from New River Watershed but not Gold Button Lagoon.
Similarly, Grain et al. (1998a) found a positive relationship between body size and E2 in
juvenile female alligators from a reference lake, while no clear relationship was detected
for juveniles from more contaminated lakes. However, while the same inter-site
difference in body size-hormone relationships was also observed for T in juvenile male
aUigators (Grain et al., 1998a), in the present study no differences in the relationship
between body size and T concentrations in juvenile males were detected between sites,
despite the fact that T concentrations at Gold Button Lagoon were reduced compared to
those at New River Watershed. Because contaminant concentrations are similar at both
69
study sites, these data suggest the possible influence of one or more sources of variation
in hormone concentrations (e.g., sex, size, habitat) not related to contaminant exposure.
A major finding midway through this study may explain much of the similarity in
hormone concentrations and body size-hormone relationships observed between New
River Watershed and Gold Button Lagoon. During the planning stages of this study,
sample collection sites were selected based on multiple factors including abundance of
crocodiles, likelihood of contamination, and logistical considerations. The discovery of
multiple OCs and mercury in crocodile eggs from Gold Button Lagoon during the pilot
study (Rainwater et al., 2002; Rainwater et al., unpublished data) formed the basis for this
lagoon being selected as the contaminated site. Conversely, New River Watershed had
not been sampled during the pilot study, but based on its remote location, distance from
large-scale agriculture, and various logistical advantages, this area was designated as the
reference site. However, approximately two years into the study, we found comparable
concentrations (< 300 ppb) of aldrin, heptachlor epoxide, lindane, and other OCs in
sediments from both sites (Wu et al., 2000a). In addition, crocodile eggs, and crocodile
scutes from both Gold Button Lagoon and New River Watershed were found to contain
p,p'-DDE, p,p'-DDT, and methoxychlor. Concentrations of/?,/?'-DDE were three-fold
higher in eggs from Gold Button Lagoon (mean [±SE] = 127.9 ± 8.8 ng/g) compared to
New River Watershed (41.2 ± 3.5 ng/g) (X^ = 45.84; d.f. = 1, 170; p < 0.0001) but not
different in scutes from both sites (Gold Button Lagoon: 81.5 ± 33.4 ng/g; New River
Watershed: 70.0 ± 16.0 ng/g; X^ = 0.15; d.f. = 1, 84; p = 0.70) (Wu, 2000; DeBusk,
2001). Following this discovery, considerable effort was made to locate non-
contaminated crocodile habitat to use as a reference site for comparisons of hormone
concentrations and other ecotoxicological endpoints. Two additional sites in northem
Belize and four additional sites in southern Belize were examined, but all crocodile eggs
from each locality were shown to contain environmental contaminants (Wu et al., 2000b;
Rainwater et al., 2002), suggesting contamination of each site and the crocodiles
inhabiting them. Similar contaminant concentrations were also found in American
crocodile (C. acutus) eggs from four sites in the coastal zone of Behze (Wu et al., 2000b),
further illustrating the ubiquitous nature of environmental contamination in the country.
70
In ensuing years, additional discussions with locals indicated that contamination with OC
pesticides (primarily DDT) associated with treated cattle feed had likely occurred on
Gold Button Ranch, and contamination related to past logging operations (i.e., treatment
of recentiy cut timber with aldrin for aphid control before floating logs down-river) in
New River Lagoon may have also resulted in site contamination (Martin Meadows,
personal communication).
Researchers examining E2 and T concentrations in alligators from contaminated
lakes in Florida have also been unable to find non-contaminated reference sites for
comparative purposes (Guillette et al., 1999a,b). However, based on well-known
contaminant inputs (e.g., documented chemical spills, agricultural and municipal mnoff
events) and land use practices associated with different lakes, researchers have been able
to identify highly contaminated sites (i.e.. Lake Apopka) and lesser contaminated sites
(i.e.. Lake Woodmff, Florida, USA) from which to make comparisons of hormone
concentrations and other endpoints (Guillette et al., 1994, 1995, 1996, 1997a, 1999a,b;
Grain et al., 1997, 1998a; Milnes et al., 2001, 2002; Pickford et al., 2000). Conversely,
records pertaining to past and present pesticide use in Belize, particularly in remote areas
associated with crocodile habitat, are non-existent. Large amounts (e.g., 8.8 metric tons
in 1979; Cawich and Rhodes, 1981) of OC pesticides including aldrin, chlordane,
dieldrin, endrin, lindane, and toxaphene were at one time commonly sold in Belize for
agricultural purposes, but no records exist on the locations and rates at which these
compounds have been used. Largely due to its relationship to human health, the use of
DDT for eradication of disease vectors in Belize has been comparatively well-
documented. The use of DDT for malaria control in Belize (then British Honduras) was
initiated in 1950, but its use in agriculture had begun earlier (Roberts et al., 2002).
Agricultural use of DDT was banned in 1988, and spraying for malaria control was
temporarily discontinued in 1993 (Roberts et al., 2002). However, recent research
(Alegria et al., 2000) and our conversations with locals suggest DDT is still used in both
agriculture and vector control. In addition, poor storage conditions of large amounts of
unused DDT may further contribute to environmental contamination in Belize (Alegria,
1998).
71
Contamination of both New River Watershed and Gold Button Lagoon with
multiple OCs, many of which are associated with endocrine dismption in alligators and
other wildlife (Vonier et al., 1996; Grain and Guillette, 1997; Grain et al., 1998b, 2000;
Guillette et al., 2002), precludes the comparison of crocodile hormone concentrations
measured in this study to legitimate reference values. This, in turn, makes it difficult to
interpret the biological significance of absolute hormone concentrations and associated
inter-site differences observed in this study. To our knowledge no other data on steroid
hormone concentrations in Morelet's crocodiles exist to which we can compare our
values. Thus, we are unsure if the hormone concentrations we have observed are within
the normal range exhibited by Morelet's crocodiles living in non-contaminated
environments or if they are abnormal (e.g., elevated, depressed). It is possible that the
concentrations of E2 and T observed in this study, although comparable between sites,
may be altered, with this condition going undetected due to the lack of suitable reference
concentrations. Conversely, it is also possible that the hormone concentrations at both
sites are unaltered and fall within the normal range for crocodiles residing in non-
contaminated habitats. Ideally, values for E2 and T concentrations in wild Morelet's
crocodiles from non-contaminated habitats in close proximity to our current study sites
would be used for comparative purposes, but based on the apparent ubiquitous nature of
environmental contamination in Belize such values may be impossible to obtain.
Although multiple studies have demonstrated inter-site variability in plasma hormone
concentrations in alhgators of the same size (Lance and Elsey, 1986; Guillette et al.,
1999a), E2 and T values for wild Morelet's crocodiles inhabiting non-contaminated
environments, even if collected outside of northern Belize, would provide at least some
basis for comparison. Considerable research on the biology of Morelet's crocodile has
been conducted in Belize and Mexico, but no other study has yet examined hormone
concentrations in this species.
Another factor that may have influenced our ability to detect differences in
hormone concentrations between sites is the potential effect of capture-induced stress on
crocodile E2 and T concentrations. GuiUette et al. (1997a) found that capture-induced
stress did not affect plasma E2 or T in juvenile alligators following 2 hr of captivity.
72
However, others have reported that capture stress induces a decline in E2 and T in adult
alligators by 4 hr post-capture (Lance and Elsey, 1986; Elsey at al., 1991). During the
present study, the timing of blood collection differed between sites. At Gold Button
Lagoon, all blood samples were collected within 5 min of capture, while at New River
Watershed blood was collected from 5 min to several hours after capture. Variability in
the timing of blood collection at New River Watershed precludes statistical analysis to
examine the potential influence of capture stress on E2 and T concentrations observed in
this study. However, although samples sizes are small, mean hormone concentrations in
blood collected from juvenile males approximately 12 hr after capture appear to be
generally higher than those from blood collected at 5 min. This is the opposite pattem
previously observed in alligators (Lance and Elsey, 1986); however. Lance and Elsey
(1986) cautioned that the effects of capture-induced stress on alligators may not
necessarily apply to other reptilian species. Indeed, while plasma T in male painted
turtles (Chrysemys picta) declined sharply within 24 hr of capture (Licht et al., 1985), T
concentrations in male snapping turtles (Chelydra serpentina) rose during the first 24 hr
post-capture (Mahmoud et al., 1989), and concentrations of T were apparentiy unaffected
by captivity in male sleepy lizards (Tiliqua rugosa) (Bourne et al., 1986). Based on this
variation in reptilian response, it is unclear if capture stress influenced plasma hormone
concentrations in Morelet's crocodiles in this study.
All crocodiles at Gold Button Lagoon were sampled within 5 min of capture,
presumably negating any effect of capture stress on hormone concentrations (Guillette et
al., 1997a), while most animals at New River Watershed were sampled several hours
after capture, increasing the likelihood of stress-induced alteration of hormone
concentrations (Lance and Elsey, 1986; Elsey at al., 1991). Despite these differences in
the time of blood collection, only two significant differences in 12 inter-site comparisons
(Figure 3.3 and Figure 3.5) between hormone concentrations at the two sites were
detected. This suggests three primary possibihties: (1) crocodile hormone concentrations
at the two sites were similar (naturally or due to some alteration), and although samples
from New River Watershed were collected at a post-capture time at which stress-induced
hormone alteration has been observed in some reptiles, the fact that few significant
73
differences in crocodile hormone concentrations were detected between sites suggests
that stress did not significantly affect hormone concentrations measured in this study, (2)
crocodile hormone concentrations at the two sites were dissimilar (naturally or due to
some alteration), and capture-induced stress at New River Watershed altered (e.g.,
reduced, increased) hormone concentrations to concentrations similar to those at Gold
Button Lagoon, and (3) capture-induced stress was size- and hormone-specific, altering T
concentrations in large juveniles only.
The fact that relative concentrations of E2 and T in crocodiles within New River
Watershed and Gold Button Lagoon closely resemble those observed in other crocodilian
species sampled during the breeding season (Lance, 1987, 1989; Kofron, 1990; Guillette
et al., 1997b) may indicate that both exposure to EDCs and capture-induced stress had
little or no effect on hormone concentrations measured in this study, and that the few
inter-site differences in hormone concentrations observed are most likely the result of
natural variation between sites. However, it is possible that hormone concentrations in
each crocodile group are in fact altered, but altered in the same manner (e.g., all
depressed or all elevated) so that these alterations go undetected without suitable
reference values for comparative purposes. Direct comparisons of hormone
concentrations from New River Watershed and Gold Button Lagoon with concentrations
from a non-contaminated site in northem Belize are required to adequately examine the
influence of contaminated habitats on steroid hormone concentrations in resident
crocodiles. Moreover, samples from all sites should be collected within a uniform time
frame minutes after capture to minimize the potential influence of stress on hormone
concentrations.
Conclusions
This study was conducted to examine plasma E2 and T concentrations in
Morelet's crocodiles living in contaminated and non-contaminated habitats in Belize.
Large juvenile crocodiles from the designated contaminated site. Gold Button Lagoon,
exhibited reduced plasma T concentrations compared to large juveniles at the reference
site. New River Watershed. No other inter-site differences in hormone concentrations
74
were observed. Discovery midway through the study that New River Watershed is
contaminated with multiple OCs at similar concentrations to those at Gold Button Lagoon
may explain the similarity in crocodile hormone concentrations between these two sites.
However, due to the lack of a non-contaminated reference site, it is unclear whether the
steroid hormone concentrations observed in this study are normal or altered by some
stressor (e.g., contaminants). Thus, the biological significance of these resuhs is difficuh
to interpret. Our inability to locate a non-contaminated reference site after sampling six
additional localities throughout the country underscores the ubiquitous nature of
environmental contamination in Belize and demonstrates the inherent difficulty of
acquiring suitable reference sites for ecotoxicological field studies (Matter et al., 1998b).
These difficulties may be exacerbated in developing countries where past or present use
of OCs and other chemicals is often widespread and unregulated, and records indicating
chemical use pattems and identifying contaminated areas do not exist. The high potential
for contamination in these areas combined with the high biodiversity characteristic of the
tropics emphasizes the need for more studies examining exposure and response of
tropical wildlife to environmental contaminants. Continuing efforts are underway to
locate non-contaminated Morelet's crocodile habitat in Belize to obtain reference steroid
hormone values for this species and further examine hormone concentrations in animals
living in contaminated and non-contaminated wetlands.
75
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Milnes, M.R., A.R. Woodward, A.A. Rooney and Guillette, Jr. 2002. Plasma steroid concentrations in relation to size and age in juvenile alligators from two Florida lakes. Comparative Biochemistry and Physiology A. 131:923-930.
Murray, D.L. 1994. Cultivating Crisis: The Human Cost of Pesticides in Latin America. University of Texas Press, Austin. 177 pp.
Norena-Barroso, E., O. Zapata-Perez, V. Ceja-Moreno and G. Gold-Bouchot. 1988. Hydrocarbon and organochlorine residue concentrations in sediments from Bay of Chetumal, Mexico. Archives of Environmental Contamination and Toxicology. 61:80-87.
Norris, D.O. 1996. Vertebrate endocrinology. 3^"^ ed. Academic Press, San Diego, CA. 634 pp.
81
Phuong, P.K., C.P.N. Son, J-J. Sauvain and J. Tarradellas. 1989. Contamination by PCBs, DDTs, and heavy metals in sediments of Ho Chi Minh City's canals, Vietnam. Bulletin of Environmental Contamination and Toxicology. 60:347-354.
Pickford, D.B., L.J.Guillette, Jr., D.A. Grain, A.A. Rooney and A.R. Woodward. 2000. Plasma dihydrotestosterone concentrations and phallus size in juvenile American alligators (A. mississippiensis) from contaminated and reference populations. Journal of Herpetology. 34:233-239.
Piatt, S.G. 1996. The ecology and status of Morelet=s crocodile in Belize. Ph.D. Dissertation. Clemson University, Clemson, SC. 187 pp.
Piatt, S.G. and J.B. Thorbjamarson. 2000. Population status and conservation of Morelet's crocodile, Crocodylus moreletii, in northern Belize. Biological Conservation. 96:21-29.
Rainwater, T.R., S.G. Piatt and S.T. McMurry. 1998. A population study of Morelet's crocodile (Crocodylus moreletii) in the New River watershed of northem Belize, pp. 206-220. In: Crocodiles. Proceedings of the 14th Working Meeting of the Crocodile Specialist Group, lUCN - The World Conservation Union, Gland, Switzerland and Cambridge UK.
Rainwater, T.R., B.M. Adair, S.G. Piatt, T.A. Anderson, G.P. Cobb and S.T. McMurry. 2002. Mercury in Morelet's crocodile eggs from northern Behze. Archives of Environmental Contaminants Toxicology. 42:319-324.
Rice, K.G., H.F. Percival, A.R. Wooward, CL. Abercrombie and P.M. Wilkinson. 1996. Clutch viability, population trends and nesting female demographics. In: Effects of environmental contaminants on the demographics and reproduction of Lake Apopka's alligators and other taxa. Florida Cooperative Fish and Wildlife Research Unit, University of Florida, GainesviUe, FL. Pp. 7-35.
Roberts, D.R., E. Vanzie, M.J. Bangs, J.P. Grieco, H. Lenares, P. Hshieh, E. Rejmankova, S. Manguin, R.G. Andre and J. Polanco. 2002. Role of residual spraying for malaria control in Belize. Joumal of Vector Ecology. 27:63-69.
Ross, J.P. (ed.). 1998. Crocodiles: status survey and conservation action plan. SSC-IUCN Crocodile Specialist Group, 2"'' ed. Gland, Switzeriand. 136 pp.
Stafford, P.J. and J.R. Meyer. 2000. A Guide to the Reptiles of Belize. Academic Press, San Diego, CA. 356 pp.
82
vom Saal, F.S., S.C. Nagel, P. Palanza, M. Boechler, S. Parmigiani and W.V. Welshons. 1995. Estrogenic pesticides: binding relative to estradiol in MCF-7 cells and effects of exposure during fetal life on subsequent territorial behavior in male mice. Toxicology Letters. 77:343-350.
Vonier, P.M., D.A. Grain, J.A. McLachlan, L.J. Guillette, Jr. and S.F. Amold. 1996. Interaction of environmental contaminants with estrogen and progesterone receptors from the oviduct of the American alligator. Environmental Health Perspectives. 104:1318-1322.
Wilson, E.O. 1992. The Diversity of Life. W.W. Norton & Company, New York, NY. 424 pp.
Woodward, A.R., H.F. Percival, M.L. Jennings and C.T. Moore. 1993. Low clutch viability of American alligators on Lake Apopka. Florida Scientist. 56:52-63.
Wu, T.H. 2000. Evaluation of organochlorine residues in Morelet's and American crocodile eggs from Belize. M.S. Thesis, Texas Tech University, Lubbock, TX. 120 pp.
Wu, T.H., T.R. Rainwater, S.G. Piatt, S.T. McMurry and T.A. Anderson. 2000a. Organochlorine contaminants in Morelet's crocodile (Crocodylus moreletii) eggs from Belize. Chemosphere. 40:671-678.
Wu, T.H., T.R. Rainwater, S.G. Piatt, S.T. McMurry and T.A. Anderson. 2000b. DDE in eggs of two crocodile species from Belize. Joumal of Agricultural and Food Chemistry. 48:6416-6420.
83
Table 3.1. Sex, number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for plasma estradiol-17p concentrations in this study.
site
Group Sex
New River Watershed
Number Size range (cmTL)
Gold Button Lagoon
Number Size range (cm TL)
Adults 158.6- 184.0 153.0-197.0
M 186.0-267,7 183.2-262,0
Large juveniles 19 86,5-145.5 12 87.2-148.6
M 47 81.4-176.8 103.1 -161.0
Small juveniles F 10 43.5 - 74.8 12 35,0-52.1
M 32 37.0-73.7 16 34,7-49.5
Table 3.2. Sex, number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for plasma testosterone concentrations in this study.
site
Group Sex
New River Watershed
Number Size range (cm TL)
Gold Button Lagoon
Number Size range (cm TL)
Adults 158.6-184,0 153.0-197,0
M 186,0-267.7 183.2-262.0
Large juveniles
Small juveniles
M
18
49
M 34
86.5 - 145.5
81.4-175.5
36,0 - 74.8
35.9-73.7
13
13
87,2-148,6
99.5-158.3
35,0-52,1
34.7 - 59,0
84
Table 3.3. Mean (±SE) plasma concentrations of estradiol-17p (E2) (pg/ml) and testosterone (T) (ng/ml) in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize.
Mean hormone concentration
Group Hormone
New River Watershed
Males Females
Gold Button Lagoon
Males Females
Adults 57,0 ±12.1 533,1 ±206.1 16.5 ±16,0 1147,8±411.3
11,0±5.5 1.7 ±0,7 2.9 ±1,3 1.5 + 0,4
Large juveniles E: 83.1+23.3 185,8 ±18.4 40.5 ± 26,3 149,9 + 26,4
3.4 ±0.6 0.9 ±0.1 0.7 ±0.2 0.5 ±0.1
Small juveniles 317.3 + 49,1 310,7 ±29.6 293.9 ± 35.8 236,9 ±19.5
1,6 ±0.2 1.0 ±0,1 2.2 ±0,5 1.0±0,1
85
Table 3.4. Results of linear regression analysis of hormone concentrations as a function of body size (TL) in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize.
Group
Adults
Large juveniles
Small juveniles
Hormone
E;
T
E:
T
E:
T
r
r
P
r
x"
P
r
r"
r
r
?
P
r
?
P
r
?
P
NewRi
Males
0.0458
0.0021
0.9144
0,4642
0.2155
0.2465
0,0400
0,0016
0,7876
0.1442
0.0208
0.3227
0,1887
0.0356
0,2774
0,0510
0,0026
0,7645
iver Watershed
Females
0,1131
0.0128
0,8561
0,5810
0.3376
0.3042
0,5100
0,2601
0,0257
0.0825
0.0068
0.7447
0.2121
0.0450
0.5310
0,0781
0,0061
0.8098
Gold Button
Males
0,9621
0,9257
0,0379
0,7157
0.5122
0,2843
0.3288
0.1081
0,4714
0,6840
0.4679
0.0901
0,2369
0,0561
0.3770
0,4090
0,1673
0,0919
Lagoon
Females
0.7102
0,5044
0.0484
0,8803
0,7749
0.0039
0,4467
0.1995
0.1455
0,5032
0,2532
0,0797
0.0964
0.0093
0,7653
0,2784
0,0775
0.3572
86
Figure 3.1. Map of Belize showing locations of the two study sites. GBL = Gold Button Lagoon, NRW = New River Watershed, NR = New River, NRL = New River Lagoon.
87
•D C 3 O
.a
c
.a vP 2
T3 C 3 O
'
A
, • Observed
Expected
r = 0.99
•
:
:
•o
i 0 o L -2
C 3 O
C -1 3 O •O -2
B
r =0.94 ;
• ^ \ • r = 0.96
1.56 3.12 6.25 12.5 25 50 100 200 400 800
Hormone concentration (pg/100 MO
Figure 3.2. Representative standard curves for hormone RIAs. A = estradiol-17P (E2), using Endocrine Sciences antibody; B = E2, using ICN antibody; C = testosterone, using Endocrine Sciences antibody.
88
1800
1600
1400
"£ 1200
3 1000 "o ^ 800 (0
W 600 UJ
400
200
0
I New River Watershed H Gold Button Lagoon
Group
Figure 3.3. Inter-site comparison of mean (±SE) plasma estradiol-17p (E2) concentrations in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. Numbers above bars indicate the number of animals sampled per site within a group. No significant difference in E2 concentrations within a group was observed. SJM = small juvenile males; SJF = small juvenile females; LJM = large juvenile males; LJF = large juvenile females; AM = aduh males; AF = adult females.
89
Q.
•a (D •J (0 LU
800
600
400
200
-New River Watershed b
: 5
-
_
a 19
a
•
47 T
X CO
CO
•
•
a.
(0 <-• (A
2000
1500
1000
500
Gold Button Lagoon
a 7
a 12 a
4
b 8
r^^«fij}
LJM LJF AM
Group
AF
Figure 3.4. Intra-site comparison of mean (±SE) plasma estradiol-17P (E2) concentrations in Morelet's crocodiles from New River Watershed (top) and Gold Button Lagoon (bottom), northern Belize. Numbers above bars indicate the number of animals sampled per site within a group. Bars with different superscripts are significantly different. LJM = large juvenile males; LJF = large juvenile females; AM = adult males; AF = adult females.
90
20
15
o c o o *^ (0 o V) 0)
10
I New River Watershed D Gold Button Lagoon
11 13
SJM SJF AM AF
Figure 3.5. Inter-site comparison of mean (±SE) plasma testosterone (T) concentrations in Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. Numbers above bars indicate the number of animals sampled per site within a group. Asterisks indicate a significant (p < 0.05) difference in T concentrations within a group. SJM = small juvenile males; SJF = small juvenile females; LJM = large juvenile males; LJF = large juvenile females; AM = aduh males; AF = adult females.
91
20
1 ''' C
0> c o 0) (0
o 4 - i (0 0)
10
c o
w o (0
u
New River Watershed
a
a : •
34
• • • - : • .
a 11
49 T
ml-
b 8
a 18
1 • • 1
< 4
•
a
T : •
SJM SJF LJM LJF
Group
Figure 3.6. Intra-site comparison of mean (±SE) plasma testosterone (T) concentrations in Morelet's crocodiles from New River Watershed (top) and Gold Button Lagoon (bottom), northem Belize. Numbers above bars indicate the number of animals sampled per site within a group. Bars with different superscripts are significantiy different. SJM = small juvenile males; SJF = small juvenile females; LJM = large juvenile males; LJF = large juvenile females; AM = adult males; AF = adult females.
92
E a
•D (0 w (0 UJ
1000
800
600
400
200
• New River Watershed O Gold Button Lagoon
r2 = 0.05. p = 0.53
r2 = 0.01, p = 0.77
Females
- V - ^ ' sP— •
_j I , L_
30 40 50 60 70 80
1000
800 -
O) Q .
- O
• o (0
^ (0 UJ
600
400
200 -
30
r2 = 0.06, p = 0.38
• o , •
• • •o o
o . • • • o •
cf o ,• • • ,
•
• •
• • • •
J 1 1 L .
•
, , ,
Males
•
•
40 50 60
Total length (cm)
70 80
Figure 3.7. Relationship between estradiol-17P (E2) concentration and body size in small juvenile (TL < 80 cm) crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. No E2-body size relationship was detected in females or males from either site.
93
I ' 0)
0) 4->
o • 4 (A
0
12
10
E "O) 8
0)
i 6 0) 4^ (0 2 4 (A 0)
• New River Watershed O Gold Button Lagoon
r" = 0.01, p = 0.81
r' = 0.08, p = 0.36
Females
o
^ ^ = " = ^ -CL..
O
r' = 0.002, p = 0.76
1 = 0.17, p = 0.09
Males
O ^tr-o»
^ ^ . ^ • • . * " * . • •
30 40 50 60 70
Total length (cm)
80 90
Figure 3.8. Relationship between testosterone (T) concentration and body size in small juvenile (TL < 80 cm) crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. No T-body size relationship was detected in females or males from either site.
94
500
400 -
E "B> 300 Q.
"jo "•5 2 200 *^ (A UJ
100 h
80
• New River Watershed • O Gold Button Lagoon
r' = 0.26, p = 0.03
r'= 0.20.0 = 0.15
•
O O
• • °°_,--<
o • o o
•
Females ,
• •
-
•
o
90 100 110 120 130 140 150 160
,-^
(pg/
m
Est
rad
iol
IHUU
1200
1000
800
600
400
200
n
r = 0.002, p = 0.79
: r = 0.11, p = 0.47
•
•
^i t# •••v4w^r4fT
o • • • • ^
0 4 W M
- f^
Males :
•
•
•
-
•
•
60 80 100 120 140 160
Total length (cm)
180 200
Figure 3.9. Relationship between estradiol-17P (E2) concentration and body size in large juvenile (females, TL = 80-149.9 cm; males, TL = 80-179.9 cm) crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. A positive relationship between body size and E2 was detected in females from New River Watershed but not in Gold Button Lagoon females or males from either site.
95
E "& 2 C V c o
(A
2 1 (A 0)
• New River Watershed O Gold Button Lagoon
40
^^ 30
c 0) c O 20 0) (A
o (A V I - 10
r'= 0.01, p = 0.74
X' = 0.25, p = 0.08 Females
o • • » o ____
• • . n . • •
. — - O "
• o o •
80 90 100 110 120 130 140 150 160
f' = 0.02, p = 0.32
l' = 0.47, p = 0.09
Males
-^^^ 60 80 100 120 140
Total length (cm)
200
Figure 3.10. Relationship between testosterone (T) concentration and body size in large juvenile (females, TL = 80-149.9 cm; males, TL = 80-179.9 cm) crocodiles from New River Watershed and Gold Button Lagoon, northern Belize. No T-body size relationship was detected in females or males from either site.
96
•D (0
w UJ
5000
4000
O) 3000 Q.
2000
1000
• New River Lagoon O GoltJ Button Lagoon
• —
r = 0.01, p = 0.86 Females l' = 0.50, p = 0.05
0
^ - - ' 6 ' 0 — o
•
• . ° ^ " • , ° • . • 150 160 170 180 190 200
160
140 I-
^ 120
E 2 100
•D (0
(0 UJ
80
60
40
20
r = 0.002, p = 0.91
r' = 0.93, p = 0.04
160 180
Males
200 220 240 260
Total length (cm)
280
Figure 3.11. Relationship between estradiol- 17P (E2) concentration and body size in aduh (females, TL > 150 cm; males, TL > 180 cm) crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. A positive relationship between body size and E2 was detected in females and males from Gold Button Lagoon but not from New River Watershed.
97
•5* o>
0) c o V
4.^ (A O
4-» (0 0)
• New River Watershed O Gold Button Lagoon
r' = 0.34, p = 0.30
r' = 0.77, p = 0.003
Females
.-o
o
o
150 160 170 180 190 200
"&
0) c o *^ (0
o (A 0>
I -
60
40
20
•^ = 0.22, p = 0.25
r = 0 .51, p = 0.28
Males
o ] o I
160 180 200 220 240
Total length (cm)
260 280
Figure 3.12. Relationship between testosterone (T) concentration and body size in adult (females, TL > 150 cm; males, TL > 180 cm) crocodiles from New River Watershed and Gold Button Lagoon, northem Belize. A positive relationship between body size and T was detected in females from Gold Button Lagoon but not in females from New River Watershed or males from either site.
98
CHAPTER IV
PHALLUS SIZE AND PLASMA TESTOSTERONE
CONCENTRATIONS IN MALE MORELET'S
CROCODILES FROM CONTAMINATED
HABITATS IN NORTHERN BELIZE
Abstract
Over the last decade, multiple reproductive abnormalities have been observed in
juvenile American alligators (Alligator mississippiensis') inhabiting Lake Apopka,
Florida, USA. Lake Apopka is heavily polluted with organochlorine (OC) pesticides and
other contaminants, primarily as the result of extensive agricultural mnoff and a major
pesticide spill. Juvenile male alligators from this lake consistentiy exhibit reduced
phallus size concurrent with reduced plasma androgen concentrations. It has been
hypothesized that the demasculinization of these animals may result from exposure to
contaminants with antiandrogenic properties. p,p '-DDE, one of the primary
contaminants of concern at Lake Apopka, preferentially binds the mammalian androgen
receptor and inhibits normal androgen function in vivo. This persistent OC has been
detected in alligator eggs and serum from Lake Apopka, suggesting its potential role in
the reproductive anomalies observed in juvenile males. The primary objective of this
study was to determine if similar pattems of demasculinization are prevalent in other
crocodilian species living in habitats contaminated with p,p '-DDE and other pollutants.
Morelet's crocodiles (Crocodylus moreletii) were sampled from two habitats in northern
Belize, New River Watershed and Gold Button Lagoon. The southem portion of New
River Watershed is a remote system with relatively few modem anthropogenic impacts,
while Gold Button Lagoon is a man-made lagoon directly adjacent to areas of large-scale
agriculture. Concentrations ofp,p '-DDE and other contaminants have previously been
found in crocodile eggs from Gold Button Lagoon. Phallus size and plasma testosterone
(T) concentrations were measured in juvenile and aduh male crocodiles from these sites
from 1998-2000. Mean phallus size did not differ within crocodile size groups (adults,
juveniles) between sites. However, the mean plasma T concentration in juveniles at Gold
99
Button Lagoon was reduced compared to that at New River Watershed. In addition, no
relationships between plasma T and body size or phallus size was observed in juveniles
from Gold Button Lagoon, while at New River Watershed these relationships were
positively correlated. For adults, no significant inter-site differences were observed in
phallus size, plasma T concentrations, or relationships between plasma T and body size
or phallus size. The contradictory nature of these results might be explained by the
discovery late in the study that New River Watershed exhibits a contamination profile
closely resembling that of Gold Button Lagoon. Due to the lack of a less-contaminated
reference site, it is unclear whether phallus size and plasma T concentrations observed at
these two sites are normal or altered by some stressor (e.g., endocrine-disrupting
chemicals). Thus, the biological significance of the few site differences observed in this
study is difficult to interpret. An inability to locate a non-contaminated reference site
after sampling multiple localities throughout the country underscores the ubiquitous
nature of environmental contamination in Belize and demonstrates the inherent difficulty
of acquiring suitable reference sites for ecotoxicological field studies. Future wildhfe
studies in tropical, developing countries where chemical use is often unregulated should
address possible influences of contaminants on research endpoints even in seemingly
pristine localities.
Introduction
Little is known concerning the embryological development of crocodilian
genitaha (Allsteadt and Lang, 1995; Guillette et al, 1996). Sex determination and
gonadal differentiation occur prior to hatching in aU crocodilian species examined to
date, but differentiation of the genitals varies among species (Allsteadt and Lang, 1995).
However, at hatching, the copulatory organ, also called the clitero-penis, exhibits sexual
dimorphism in size and shape. Generally, in males the chtero-penis is larger, rounder,
and redder, while in females the organ is smaller and whiter (Allsteadt and Lang, 1995).
This sexual dimorphism is primarily the result of differential growth rates, with growth
being more rapid in males than females (Allsteadt and Lang, 1995). Examining
100
differences in clitero-penis size is currently the most reliable and least invasive method of
determining sex in crocodilians.
While it is known that sexual differentiation, development, and growth of the
genitalia in most reptiles (e.g., snakes, turtles, and lizards) are regulated by sex steroid
hormones (Raynaud and Pieau, 1985), little is currently known concerning the specific
influences of steroids on the genitalia during embryonic development in crocodilians
(Raynaud and Pieau, 1985). However, the data available suggest crocodilian genitalia are
also responsive to steroid hormones. Lang and Andrews (1994) observed reductions in
clitero-penis size in hatchling alligators (Alligator mississippiensis) and muggers
(Crocodylus palustris) following topical treatment of eggs with estradiol-17P (E2) during
incubation. Conversely, juvenile alligators and muggers dosed with testosterone (Forbes,
1938a) and testosterone proprionate (Forbes, 1939; Ramaswami and Jacob, 1965)
exhibited marked clitero-penis growth, while administration of estrone had no effect
(Forbes, 1938b). Recent studies on male genital size and plasma steroid concentrations
in wild alligators have also suggested that clitero-penis development and growth are
androgen dependent (Guillette et al., 1996; Pickford et al., 2000). Based on this
relationship, Guillette et al. (1996) proposed that phallus size could serve as an indicator
of abnormalities in androgen concentrations or function in reptiles.
In recent years, multiple reproductive abnormalities suggestive of disruption of
normal endocrine function have been reported in alligators inhabiting Lake Apopka,
Florida, USA (Grain et al., 1997, 1998a; Guillette et al., 1994, 1995a, 1996, 1997a,
1999a,b; Milnes et al., 2001, 2002; Pickford et al., 2000). Lake Apopka is a highly
contaminated lake in central Florida as the result of extensive agricultural pesticide and
nutrient mnoff, municipal wastewater discharge, and a major organochlorine (OC)
pesticide spill in 1980 (Matter et al., 1998a; Guillette et al., 2000). Over the past two
decades, laboratory and field studies examining exposure and response of alligators to
endocrine-dismpting contaminants in Lake Apopka have provided one of the most
comprehensive ecotoxicological assessments on a wildlife species to date (for a review,
see Grain and Guillette, 1998; Guillette et al., 2000).
101
Among the many endpoints examined, juvenile male alligators from Lake Apopka
have consistentiy exhibited reduced plasma T concentrations when compared to juveniles
from a relatively non-contaminated lake, Lake Woodruff (Guillette et al., 1994, 1996,
1997a, 1999a,b; Grain et al., 1998). Abnormal hormone concentrations during critical
eariy life stages suggest that anatomical structures dependent on these hormones for
proper growth and development (e,g., genitalia) may also be altered (Guillette et al.,
2000). Indeed, juvenile males from Lake Apopka have also consistentiy exhibited
smaller clitero-penis size than juveniles from Lake Woodmff (Guillette et al., 1994, 1996,
1999a,b; Pickford et al., 2000). The primary OC pesticide detected in alligator eggs and
semm from Lake Apopka is p,p '-DDE, a persistent metabolite of DDT (Heinz et al.,
1991; Guillette et al., 1999b). Although untested in crocodilians, this compound has been
found to preferentially bind the androgen receptor in rats, inducing multiple
antiandrogenic effects in both pubertal and adult males (Kelce et al., 1995). These data
suggest that alligators and other wildlife exposed to p,p '-DDE and other xenobiotic anti-
androgens during development may experience disruption of normal androgen activity,
leading to reproductive abnormalities later in life. This, in tum, suggests the reductions
in the size of the clitero-penis, hereafter referred to as phallus or penis, observed in
juvenile male alligators from Lake Apopka may be related to contaminant-induced
endocrine dismption.
Although no direct cause-effect relationship has been established between the
reproductive abnormalities observed in Apopka alligators and environmental
contaminants in the lake, results of laboratory and field observations over the last 20
years strongly suggest the potential for contaminant-induced endocrine dismption at
various levels of organization in these animals (Grain and Guillette, 1998). Further,
many of the reproductive alterations observed in Apopka alligators have also been
observed in alhgators from other, lesser contaminated lakes in Florida (Grain et al.,
1998a; Guillette et al., 1996, 2000; Hewitt et al., 2002), demonstrating that these
anomalies are not limited to Lake Apopka and may occur in alligators and other wild
crocodilians inhabiting polluted systems. However, despite numerous reports of
environmental contaminant exposure in crocodilian species woridwide (see Rainwater et
102
al , 2002) and a paucity of data concerning endocrine responses of crocodilians to these
contaminants (Guillette and Milnes, 2000), no studies have examined endpoints of
endocrine disruption in crocodilians outside of Florida.
To address this data gap, we conducted a pilot study in 1995 to examine exposure
of Morelet's crocodile (C. moreletii) to environmental contaminants in Belize. Morelet's
crocodile is a moderate-sized, freshwater crocodilian found in the Atlantic and Caribbean
lowlands of Mexico, Guatemala, and Belize (Groombridge, 1987; Lee, 1996; Ross, 1998)
and is currentiy recognized as endangered under the United States Endangered Species
Act (Endangered and Threatened Wildlife and Plants, 1991). Detectable concentrations
of multiple contaminants, including p,p '-DDE and mercury, were found in eggs from
three localities in the northem portion of the country (Rainwater et al., 2002; Rainwater et
al., unpublished data). Based on these findings and data from Lake Apopka showing egg
contamination, population declines, and reproductive abnormalities in alligators exposed
to many of the same chemicals (Jennings et al., 1988; Heinz et al., 1991; Woodward et
al., 1993; Grain et al., 1997, 1998a; Guillette et al., 1994, 1995a, 1996, 1997a, 1999a,b;
Milnes et al., 2001, 2002; Pickford et al., 2000), a multi-year study was initiated to
examine various endpoints of contaminant exposure and response in Morelet's crocodiles
living on contaminated and reference sites in Belize. This paper describes one
component of that study in which male phallus size and plasma T concentrations were
examined in juvenile and adult crocodiles from contaminated and reference habitats. In
light of the reductions in phallus size and circulating T concentrations observed in
alligators from Lake Apopka (Guillette et al., 1994, 1996, 1997a, 1999a,b; Grain et al.,
1998a), we hypothesized in this study that crocodiles living in contaminated habitats in
Behze would also exhibit smaller phalli and depressed T concentrations compared to
crocodiles from non- or less-contaminated wetlands.
103
Materials and Methods
Study Sites
Crocodiles were captured and samples collected from two sites in northern Belize,
Gold Button Lagoon and the New River Watershed. Gold Button Lagoon (17°55'N,
88°45'W) is a large man-made lagoon located on Gold Button Ranch, a 10,526 ha private
cattle ranch approximately 25 km southwest of Orange Walk Town, Orange WaUc
District (Figure 4.1). Gold Button Ranch is situated adjacent to an intensively farmed
settlement, and past use of OC pesticides in this area (on crops) as well as on the ranch
itself (in cattle feed and dip) is believed to have occurred (Martin Meadows, pers.
comm.). New River Watershed is comprised of the New River, New River Lagoon, and
associated tributaries in the Orange Walk and Corozal Districts (Figure 4.1). New River
Lagoon (17°42'N, 88°38'W; ca. 23 km long) and the southem-most 18 km of New River
constituted the New River Watershed study site for this project. This section of New
River Watershed is relatively remote, bordered by semi-evergreen seasonal forest
(Stafford, 2000) to the west and seasonally flooded savanna to the east. Both Gold
Button Lagoon and New River Watershed contain some of largest Morelet's crocodile
populations in Belize (Piatt, 1996; Rainwater et al., 1998; Piatt and Thorbjarnarson,
2000). During the 1995 pilot study, crocodile eggs from Gold Button Lagoon were found
to contain multiple environmental contaminants (Rainwater et al., 2002; Rainwater et al.,
unpublished data). Thus, when designing the present study. Gold Button Lagoon was
selected as the contaminated site. Although no samples had been collected from New
River Watershed, based on its remote location, surrounding topography limiting large-
scale agriculture, and logistical advantages, this area was selected as the reference site.
Animals, Blood Sampling, and Morphometries
Crocodiles from both sites were hand- or noose-captured at night from a boat
under permit from the Belize Ministry of Natural Resources from April through October,
1998-2000. Blood (volume not exceeding 1.6% body mass) was collected from the post-
cranial sinus, transferred to an EDTA-treated Vacutainer®, placed on ice in the field, and
later centrifuged at 2000 rpm for 10 minutes. The plasma supematant was then
104
transfen-ed to a collection tube and frozen at -25°C until shipment to Texas Tech
University. Samples were then stored at -80°C until assayed for T concentrations.
Following blood collection, measures of total length (TL; measured ventrally), ^nout-vent
length (SVL; measured ventrally from the tip of the snout to the anterior margin of the
cloaca), and mass were obtained. Animals were categorized as juveniles (females TL <
150 cm; males - TL < 180 cm) or adults (females - TL > 150 cm; males - TL > 180 cm)
(Table 4.1). The size at which male Morelet's crocodiles become reproductively active
(adults) in northem Belize is unknown. Thus, for this study the adult size class for males
was based on that reported for alligators (> 1.8 m; Ferguson, 1985), while the adult size
class for females was based on the smallest known nesting female Morelet's crocodile on
either of our study sites (150 cm TL; Piatt, 1996). Next, sex was determined for each
animal by cloacal examination of the genitalia (Lang and Andrews, 1994; Allsteadt and
Lang, 1995; Piatt, 1996; Guillette et al., 1996; Pickford et al., 2000). The genitaha of,
Morelet's crocodile is similar to that of alligators and other crocodilians examined to date
in that in both sexes the organ is similar in general shape but differs significantiy in size
and coloration between males and females (larger and redder in males) (Lang and
Andrews, 1994; Rainwater et al., unpublished data). If male, the length of the penis tip
and diameter of the penis cuff were measured to the nearest 0.1 mm using a dial caliper
with needle tips (see Guillette et al., 1996; Pickford et al., 2000). Tip length was
measured from the distal edge of the cuff to the distal edge of the tip on the lateral surface
of the everted phallus (Pickford et al., 2000) (Figure 4.2). Cuff diameter was measured
from the dorsal to ventral surface of the cuff at its midpoint (Figure 4.2). A single
researcher (TRR) took all measurements to minimize and standardize measurement error
(Guillette et al., 1996). Each measurement was taken in triplicate, and a mean value was
used in subsequent analyses. Following sample and data collection, each crocodile was
marked and released at its site of capture.
Steroid Hormone Radioimmunoassay
Crocodile plasma T was analyzed using radioimmunoassays previously validated
for aUigator plasma (Guillette et al , 1997a) and modified by Grain et al. (1997). Each
105
plasma sample (40 il) was extracted 2X with ethyl ether. Briefly, plasma was mixed
with 4 ml ether for 1 min. The aqueous layer was frozen in a dry ice-methanol bath (-
30°C) and the ether supernatant decanted into a borosilicate glass assay tube. The ether
extract (supernatant) was then dried using a vortex evaporator (Labconco, Lenexa, KS)
for 15 min. The aqueous pellet was re-extracted with ether and this second ether extract
added to the assay tube. The ether extract was then dried by vortex evaporation.
Extraction efficiency averaged 71%, and the assay had a linear range of at least two
orders of magnitude (1.56 - 800 pg hormone/100 il borate buffer) (see Chapter III).
Dried samples were resuspended with borate buffer (100 \i\; 0.5 M; pH 8.0). To reduce
nonspecific binding, 100 il of borate buffer with bovine semm albumin (BSA, fraction
V; Fisher Scientific) at a final assay concentration of 0.15% was added to each tube.
Antibody (Endocrine Sciences, Calabasas Hills, CA, USA) was then added to each tube
(200 fil; final concentration of 1:25,000). Cross reactivities of the T antibody to other
ligands are as fohows: dihydrotestosterone, 44%; A-1-testosterone, 41%; A-1-
dihydrotestosterone, 18%; 5 a-androstan-3p, 17p-diol, 3%; 4-androsten-3p, 17P-diol,
2.5%; A-4-androstenedione, 2%; 5p-androstan-3p, 17p-diol, 1.5%; estradiol, 0.5%; aU
other ligands <0.2%. Finally, 100 jxl of radiolabeled steroid was added (12,000 cpm per
100 [xl; [1,2,6,7- H]testosterone @ 1 mCi/ml; Amersham Intemational, Piscataway, NJ,
USA) to each tube. For standard tubes, T was added at 0, 1.56. 3.13, 6.25, 12.5, 25, 50,
100, 200, 400, and 800 pg/tube. Tubes were vortexed for 1 min and incubated ovemight
at 4°C. All standards and samples were prepared in duplicate. Separation of bound and
free hormone was accomplished by adding 500 \i\ 5% charcoal-0.5% dextran and
immediately centrifuging at 2500 rpm at 4°C for 30 min. Lastly, 500 il of the
supematant was added to 5 ml scintillation cocktail, and the tubes were counted on a
Beckman LS 6500 scintillation counter. Interassay variance was 22%, and intraassay
variance was 4.70%. Samples were analyzed by size group and were randomly selected
(both sexes) for each assay.
106
Statistical Analyses
The relationships between phallus size and body size, body size and T
concentrations, and phallus size and T concentrations were tested using linear regression
analysis. Although animals were separated into size classes based on TL, SVL was used
as the independent variable in regression analyses involving body size for ease of
comparison with previous studies (Guillette et al., 1996, 1999a,b, 2000; Pickford et al.,
2000). Values for TL and SVL for the crocodiles sampled in this study were highly
correlated (Figure 4.3), and statistical significance of results using TL as the independent
variable in each test were not different than those obtained when using SVL. No
differences in body size were observed for juveniles (X^ = 2.12; d.f = 1, 83; p = 0.15) or
adults (X^ - 0.00; d.f = 1,9; p = 1.00) from the two study sites. Values for phallus size
(tip length, cuff diameter), body size (SVL), and plasma T concentrations were log- or
square root-transformed to obtain normality and homogeneity of variance (the latter
checked using Levene's test). Non-parametric Wilcoxon rank sums test (two groups) or
the Kmskal-Wallis test (more than two groups) was used if an assumption of the
ANOVA was not met following transformation. The majority of data were not normally
distributed; so to maintain consistency, all data were analyzed using non-parametric
statistics. All statistical tests were considered significant when p < 0.05. All analyses
were performed using program JMPin statistical software (Version 3.2, SAS Institute,
Gary, NC, USA).
Results
Animals Captured and Sampled
Phallus measurements and blood samples were obtained from a total of 153
individual male crocodiles from 1 April 1998 to 13 October 2000 (Table 4.1). Due to
logistical constraints (e.g., no access to motor boats at Gold Button Lagoon; most animals
captured from a canoe, which reduced the area searched on a given night and increased
the difficulty of crocodile capture and sampling), considerably more animals were
captured at New River Watershed (n = 127) than Gold Button Lagoon (26).
Concentrations of T in crocodilians fluctuate seasonally (Lance, 1987, 1989; Kofron,
107
1990; Guillette et al., 1997b), but phallus size does not (Rooney, 1998 as cited in Guillete
et al., 2000). Thus, to minimize temporal effects of sampling, only animals captured in
April and May were included in analyses involving T (Table 4.2). This period
corresponds to the peak of the breeding season for Morelet's crocodile in northem Belize
(Piatt, 1996; see Chapter II). All animals were included in analyses of inter-site
differences in mean phallus morphometries and phallus size-body size relationships.
Phallus Morphometries
The length of the phallus tip and diameter of the phallus cuff were significantiy
correlated with body size (SVL) in juvenile crocodiles from both sites, and the strength of
these relationships was generally uniform between crocodile groups and habitats (Table
4.3, Figure 4.4). Snout-vent length explained 71% of the variation in tip length among
juveniles from New River Watershed and 80% of the variation in tip length among
juveniles from Gold Button Lagoon. Likewise, SVL explained 72% and 85% of the
variation in cuff diameter among animals from New River Watershed and Gold Button
Lagoon, respectively. Among adult crocodiles, phallus tip length was also significantiy
correlated with body size in animals from both sites (Table 4.3, Figure 4.5). Snout-vent
length explained 60% of the variation in tip length among animals from New River
Watershed and 90% of the variation in tip length among animals from Gold Button
Lagoon. Phallus cuff diameter was significantly correlated with body size in animals
from New River Watershed, with 68% of the variation in cuff diameter explained by SVL
(Table 4.3, Figure 4.5). Among adults from Gold Button Lagoon, SVL explained 64% of
the variation in cuff diameter, but the cuff diameter-SVL relationship was not significant
(Table 4.3, Figure 4.5).
Mean phallus size within groups was also similar between sites (Table 4.4). No
significant inter-site difference in phallus tip length was observed in juveniles (X^ = 1.92;
d.f = 1, 127; p = 0.17) or adults (X^ = 0.004; d.f = 1, 267; p = 0.95) from New River
Watershed and Gold Button Lagoon (Figure 4.6). Likewise, mean cuff diameter was not
significantiy different between sites in juveniles (X^ = 0.45; d.f = 1, 127; p = 0.50) or
adults (X^ = 004; d.f = 1, 26; p = 0.95) (Figure 4.6). Phallus tip lengths among juveniles
108
at New River Watershed were 12% larger than those of juveniles from Gold Button
Lagoon, while tip lengths from adults at Gold Button Lagoon were 8% larger than those
of adults from New River Watershed. Further, cuff diameter in both juveniles and adults
from Gold Button Lagoon was 1% larger than those in juveniles and adults from New
River Watershed. However, none of these differences were statistically significant.
Plasma Testosterone Concentrations
Regression analysis indicated a significant relationship between SVL and plasma
T in juveniles from New River Watershed (r = 0.11; p = 0.01), but not in juveniles from
Gold Button Lagoon or adults from either site (Table 4.5, Figure 4.7). When one juvenile
from New River Watershed with an exceptionally high plasma T concentration (27.25
ng/ml) was removed from the analysis, the SVL-T relationship remained significant (r =
0.09, p = 0.02).
Mean plasma T concentrations in adults were not significantly different between
sites (X^ = 2.4; d.f = 1, 9; p = 0.12) (Table 4.6, Figure 4.8). However, juveniles from
New River Watershed exhibited significantly higher T concentrations than juveniles from
Gold Button Lagoon (X^ = 6.10; d.f = 1, 69; p = 0.01) (Figure 4.8). Comparison of T
concentrations among all crocodile groups indicated that adults from New River
Watershed had significantly higher T concentrations than those in Gold Button Lagoon
juveniles but not Gold Button Lagoon adults or New River Watershed juveniles (X =
11.30; d.f = 3, 78; p = 0.01) (Figure 4.8).
Regression analysis further indicated a positive relationship between plasma T
concentration and both phallus morphometries in juveniles from New River Watershed
(tip length - r = 0.10, p = 0.02; cuff diameter - r = 0.16, p = 0.003) (Table 4.7, Figure
4.9). However, when the juvenile with the exceptionally high plasma T concentration
(see above) was removed from the analysis, these significant relationships disappeared.
No positive relationships between plasma T and phallus size were observed in juveniles
from Gold Button Lagoon or aduhs from either site (Table 4.7, Figure 4.10).
109
Discussion
Over the last decade, multiple studies have reported reproductive abnormalities in
juvenile alligators inhabiting polluted lakes in Florida, particularly Lake Apopka,
suggesting potential disruption of normal endocrine function by one or more
environmental contaminants (Guillette et al., 1994, 1996, 1997a, 1999a,b; Grain et al.,
1998a; Pickford et al., 2000; Milnes et al., 2002). One of the most consistentiy observed
anomalies has been reduced phallus size in juvenile males concurrent with reduced
circulating T concentrations (Guillette et al., 1994, 1996, 1999a,b). Guillette et al.
(1999a) proposed multiple hypotheses to explain the reduced phallus size in juvenile
male alligators from Lake Apopka. These include reduced circulating androgen
concentrations, alterations in the ratios of T to dihydrotestosterone and T to E2, reduced
numbers of androgen receptors on phallic tissue, and androgen receptor blockage by an
androgen antagonist (Guillette et al., 1999a). Currentiy, it remains unclear if the
depressed phallus size observed in juvenile males at Lake Apopka is directiy related to
reduced T concentrations or if other factors, singly or in combination, are involved
(Pickford et al., 2000).
Results of the present study are equivocal, as juvenile male crocodiles from New
River Watershed and Gold Button Lagoon exhibit characteristics consistent with those
seen in juvenile male alligators from reference and contaminated sites in Florida,
respectively. For example, crocodiles from New River Watershed, the reference site in
the present study, had the following characteristics in common with juvenile alligators
from Lake Woodmff, the reference site in many of the Florida studies: (1) higher mean
plasma T concentration than juveniles at the contaminated site, (2) a positive relationship
between body size and plasma T concentrations, and (3) a positive relationship between
phallus size and plasma T concentrations. Conversely, crocodiles from Gold Button
Lagoon, the contaminated site in the present study, had the following characteristics in
common with juvenile alligators from Lake Apopka, the contaminated site in the Florida
studies: (1) reduced mean plasma T concentration in juveniles compared to the reference
site, (2) no positive relationship between body size and plasma T concentrations, and (3)
no positive relationship between phallus size and plasma T concentrations. These inter-
110
study similarities suggest juvenile crocodiles living in more contaminated habitats in
Belize may be experiencing the same reductions in reproductive fitness observed in
juvenile alligators from contaminated habitats in Florida. However, one critical
observation confounds this notion: mean phallus size of crocodiles did not differ among
sites. This result contrasts markedly with data from Florida which have consistently
shown reduced phallus size in juvenile alligators from Lake Apopka compared to Lake
Woodmff (Guillette et al., 1994, 1996, 1999a,b). In addition to similar-sized phalli,
juvenile crocodiles from both study sites in Belize were also similar in that they both
exhibited significant body size-phallus size relationships. Body size-phallus size
relationships in Lake Woodruff alligators have been consistentiy strong, whereas
relationships have been weak or absent in animals from Lake Apopka (Guillette et al.,
1996, 1999a, 2000; Pickford et al., 2000).
Results for adult crocodiles examined in this study are also difficult to discem.
Animals from both sites exhibited no differences in phallus size, plasma T
concentrations, body size-T relationships, and phallus size-T relationships. In addition,
body size was significantly correlated with penis tip length in adults from both sites and
with cuff diameter in adults from New River Lagoon. Although the body size-cuff
diameter relationship in adults from Gold Button Lagoon was not significant (p = 0.06),
the significant relationship between body size and tip length suggests a stronger overall
relationship between phallus size and body size in these animals than would be inferred if
cuff diameter was considered alone. As was observed with juvenile crocodiles, adults
exhibited characteristics similar to those seen in alligators from both contaminated and
reference sites in Florida. Similar to alligators from Lake Apopka, no positive
relationships between T and body size or T and phallus size were observed in adult
crocodiles. Conversely, similar to alligators from Lake Woodmff adult crocodiles from
both sites exhibited significant body size-phallus size relationships. Although these
endpoints have not been specifically examined in aduh animals in the Florida studies,
Guillette et al. (1996) observed reduced plasma T concentrations in subaduh male
alligators from Lake Apopka, suggesting that this condition, and possibly others related
to endocrine dismption, are not transitory but persist in older, larger animals (Guillette et
111
al., 1995b). It should be noted that the statistical similarities (and one difference)
observed for adult crocodiles in this study may be due, in part, to the small number of
adults sampled at each site.
A significant finding late in this study may at least partially explain the variable
and conflicting nature (e.g., reduced T in Gold Button juveniles but no corresponding
inter-site difference in phallus size) of these results. Sample collection sites in Belize
were selected based on multiple factors including likelihood of contamination, logistical
considerations, and abundance of crocodiles. Gold Button Lagoon was designated as the
contaminated site after multiple OCs and mercury were detected in crocodile eggs from
the lagoon in 1995 (Rainwater et al., 2002; Rainwater et al., unpublished data). Although
New River Watershed had not been previously sampled, it was designated as the
reference site based on its remote location, distance from large-scale agriculture, and
various logistical advantages. However, approximately two years into the study,
comparable concentrations of multiple OCs were found in sediments, crocodile eggs, and
crocodile scutes from both sites (Wu, 2000; Wu et al., 2000a; DeBusk, 2001). Following
this discovery, considerable effort was made to locate non-contaminated crocodile habitat
to use as a reference site for comparisons of phallus size, steroid hormone concentrations,
and other ecotoxicological endpoints. However, crocodile eggs from six additional sites
in northem and southem Belize were also found to be contaminated (Wu et al., 2000b;
Rainwater et al., 2002). Similar contaminant concentrations were also found in American
crocodile (C. acutus) eggs from four localities within the coastal zone of Belize (Wu et
al., 2000b), further illustrating the widespread environmental contamination in the
country and the inherent difficulty of acquiring suitable reference sites for
ecotoxicological field studies (Matter et al., 1998b).
Researchers in Florida have experienced similar difficulties locating non-
contaminated reference sites for comparative purposes (Guillette et al., 1999a,b).
However, based on records and general knowledge of chemical inputs and sources,
researchers have been able to identify highly contaminated sites (i.e.. Lake Apopka) and
lesser contaminated sites (i.e.. Lake Woodmff Florida) from which to make comparisons
of various ecological and biological endpoints (Grain et al., 1997, 1998a; GuiUette et al.,
112
1994, 1995a, 1996, 1997a, 1999a,b; Milnes et al., 2001, 2002; Pickford et al., 2000). In
Belize, records pertaining to past and present pesticide use are largely non-existent. The
few reports that do exist indicate numerous OC pesticides including aldrin, chlordane,
dieldrin, endrin, lindane, and toxaphene were at one time commonly sold in Belize for
agricultural purposes (e.g., Cawich and Rhodes, 1981; Alegria et al., 2000), but the
amounts and locations in which these compounds were applied are unknown. DDT was
used to control agricultural pests in Belize (then British Honduras) since the 1940s, and
although banned for such use in 1988, its use in vector control continues today (Roberts
et al., 2002). Recent research, as well as our conversations with locals, indicate
agricultural use of DDT and other OCs also persists (Alegria et al., 2000). Numerous
OCs including DDTs, PAHs and PCBs have been found in sediments in Chetumal Bay,
Mexico (Norena-Barroso et al., 1998), approximately 100 km from New River Watershed
and Gold Button Lagoon, and multiple metals have been detected in sediments in Belize
City Harbor (Gibbs and Guerra, 1997). Alegria et al. (2000) recentiy found elevated
concentrations of p,p '-DDT, p,p '-DDE, p,p '-DDD, aldrin, dieldrin, and other OCs in air
samples from Belize, suggesting the potential for contamination of this region through
atmospheric deposition. Due to the remote location of New River Watershed and relative
absence of nearby agriculture, atmospheric deposition may be the most significant and
continual source of OC contamination at this site (Eisenreich et al., 1981; Rapaport et al.,
1985; Alegria et al , 2000).
Comparable contamination of both New River Watershed and Gold Button
Lagoon with multiple OCs, many of which are associated with endocrine disruption in
alhgators and other wildlife (Vonier et al., 1996; Grain and Guillette, 1997; Grain et al.,
1998b, 2000; Guillette et al., 2002), precludes the comparison of phallus size and plasma
T concentrations measured in this study to legitimate reference values. Consequently, the
biological significance of crocodile phallus sizes, plasma T concentrations, and
associated relationships observed at these sites remain unknown. It is unclear if the
phallus sizes and T concentrations observed in this study are within the normal range
exhibited by Morelet's crocodiles living in non-contaminated environments or if they are
abnormal (e.g., elevated, depressed). It is possible that the phallus sizes observed in this
113
study, although comparable between sites, may be abnormal, with this condition going
undetected due to the lack of suitable reference values. Conversely, it is also possible
that crocodile phalli at both sites are unaltered and fall within the normal range for
crocodiles residing in non-contaminated habitats. Likewise, the inter-site difference in
juvenile crocodile T concentrations may be representative of contaminant-induced
endocrine disruption as proposed for juvenile alligators living in Lake Apopka, or
concentrations at both areas may be normal, with inter-site differences due to natural
variability (Lance and Elsey, 1986; Guillette et al., 1997b) or other factors not related to
contaminant exposure.
Conclusions
The similarity in contaminant profiles between New River Watershed and Gold
Button Lagoon suggests similarities in crocodile exposure and response to these
compounds at both sites. If exposure to one or more of these chemicals can in fact
influence phallus growth, an inter-site similarity in crocodile phallus size might be
expected as a result of the similarity in crocodile exposure to contaminants at both sites.
However, in all studies which have reported reduced phallus size in alligators living in
contaminated habitats, animals exhibiting depressed phallus size have also exhibited
depressed plasma T concentrations (Guillette et al., 1994, 1996, 1999a,b). In the present
study, we observed reduced plasma T concentrations in juveniles from Gold Button
Lagoon compared to those from New River Watershed, but juvenile phallus size was not
different between sites. Based on these findings, we propose three primary hypotheses to
explain the resuhs obtained in this study: (1) environmental contaminants are not
influencing phallus size and other endpoints measured in this study, and all inter-site
differences observed are the result of natural variability, other factors unrelated to
contaminants, or both, (2) environmental contaminants are influencing phallus size and
other endpoints measured in this study, and all inter-site differences observed are the
resuh of contaminant-induced endocrine disruption, and (3) environmental contaminants
and natural factors both influence the endpoints measured in this study, and comparison
114
of these endpoints with those from a suitable reference site is needed to determine the
relative influence of each.
Despite contamination with numerous OC pesticides and other chemicals. New
River Watershed and Gold Button Lagoon exhibit two of the largest Morelet's crocodile
populations in Belize (Rainwater et al., 1998; Piatt and Thorbjarnarson, 2000). However,
many of the most serious effects of endocrine dismpting chemicals on reproduction are
organizational in nature, occurring during early development and inducing permanent
effects that may be subtie and unapparent until later in life (Guillette et al., 1995; Grain
and Guillette, 1997). Thus, contaminant-related effects at these higher levels of
organization will only be discerned through consistent and long-term monitoring.
Researchers conducting studies, ecotoxicological or otherwise, on tropical wildlife must
be aware of potential environmental contamination in remote, seemingly pristine areas
and account for possible influences of contaminants on research endpoints.
115
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121
Table 4.1. Sex, number, and size range (cm total length [TL]) of male crocodiles from New River Watershed and Gold Button Lagoon for which phallus size was measured in this study.
Site
Group
New River Watershed
Number Size range (cm TL)
Gold Button Lagoon
Number Size range (cm TL)
Adults 20 180,1-267,7 184,0-298,7
Juveniles 107 44.9-177.0 20 38,0- 161,0
Table 4.2. Sex, number, and size range (cm total length [TL]) of crocodiles from New River Watershed and Gold Button Lagoon sampled for plasma testosterone concentrations in this study.
site
Group
New River Watershed
Number Size range (cm TL)
Gold Button Lagoon
Number Size range (cm TL)
Adults 186.0-267,7 184.0-262.0
Juveniles 56 44,9-174.9 13 39.0-158.3
122
Table 4.3. Results of linear regression analysis comparing snout-vent length (SVL) and phallus size (tip length and cuff diameter) in male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize.
Group
Adults
Juveniles
n
r
r*
P
n
r
x"
P
New River Watershed
Tip Length
20
0.7777
0,6048
< 0,0001
107
0,8439
0,7121
< 0,0001
Cuff Diameter
20
0,8258
0,6820
< 0.0001
107
0,8476
0.7184
< 0.0001
Gold Button Lagoon
Tip Length
6
0,9472
0,8972
0,0041
20
0.8928
0,7971
< 0,0001
Cuff Diameter
6
0.8009
0,6415
0,0555
20
0.9218
0,8497
< 0,0001
Table 4.4. Mean (±SE) of phallus tip length (mm) and cuff diameter (mm) of juvenile and adult male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize.
Mean size (mm)
New River Watershed Gold Button Lagoon
Group Tip length Cuff diameter Tip length Cuff diameter
Adults
Juveniles
10.63 + 0.98(20) 23.21 ±2,06(20)
3,07 ±0.18 (107) 6,13 ±0.37 (107)
11,55 ±2.47 (6)
2.69 ± 0.45 (20)
23.47 ±2,87 (6)
6,20 ±1.12 (20)
123
Table 4.5. Results of linear regression analysis comparing body size (SVL) and plasma testosterone (T) concentrations of male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize.
Group
Adults
Juveniles
n
r
r'
P
n
r
r"
P
New River Watershed
6
0.0872
0.0076
0.8700
56
0.3268
0.1068
0.0140
Gold Button Lagoon
,3
0.9402 ,
0.8839
0.2214
13
0.2302
0.0530
0.4491
Table 4.6. Mean (±SE) plasma testosterone (T) concentrations (ng/ml) in juvenile and adult male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northem Belize.
T (ng/ml)
Group New River Watershed Gold Button Lagoon
Adults
Juveniles
6.07 ±1.64 (6)
2.97 ±0.51(56)
2.52 ±1.81 (3)
1.51 ±0.52(13)
124
Table 4.7. Results of linear regression analysis comparing phallus size (tip length and cuff diameter) and plasma testosterone concentrations of male Morelet's crocodiles from New River Watershed and Gold Button Lagoon, northern Belize.
Group
Adults
Juveniles
n
r
r'
P
n
r
^
P
New River
Tip Length
6
0.3216
0,0291
0.7467
56
0.3216
0.1034
0,0157
Watershed
Cuff Diameter
6
0,0141
0.0002
0.9786
56
0.3947
0.1558
0.0026
Gold Button Lagoon
Tip Length
3
0,9904
0,9808
0.0886
13
0,1153
0.0133
0,7072
Cuff Diameter
3
0.9725
0,9458
0,1495
13
0,1304
0,0170
0.6709
125
Mx/ .._/^ Caribbean
Sea
l
1
17°N-
16°N"
Figure 4.1. Map of Belize showing locations of the two study sites. GBL = Gold Button Lagoon, NRW = New River Watershed, NR = New River, NRL = New River Lagoon.
126
Ventral surface of crocodile ("belly")
Phallus cuff
Cloacal opening
Phallus tip
Adapted from Allsteadt and Lang, 1995
Tip Length
Cuff Diameter
Adapted from Guillette et al., 1996
Figure 4.2. Diagram of the crocodilian (alligator) phallus, showing its primary components (top) and points from which measurements were taken (bottom) from Morelet's crocodiles in this study.
127
160
140
120
^ 100
C 80 a>
Tota
l
60
40
20
0
160
140
•_ Gold Button Lagoon
-
* r^ = 0.9994
p < 0.0001
n = 26
50 100 150 200
120
>H- 100
o C 80 h 0) _l
•5 60 -
o »-
40
20 -
0
New River Watershed
250 300
r = 0.9977
p < 0.0001
n = 127
50 100 150 200
Snout-Vent Length (cm)
250
350
300
Figure 4.3. Relationship between total length (TL) and snout-vent length (SVL) in Morelet's crocodiles sampled in this study.
128
(mm
L
eng
th
Pen
is T
ip
14
12
10
8
6
4
2
- • O
J-
New River V\i Gold Button
Q*J*'
Lagoon
•
/ ^ ^
r° = 0.71, p< 0.0001
r = 0.80, p < 0.0001
-
•
30
E E
^ • ^
^ 0) * 0) E re O
%: 3 o <o c 0) Q.
?5
?n
15
10
20 40 60 80
r = 0.72, p < 0.0001
r = 0.85, p < 0.0001
20 40 60
Snout-Vent Length (cm)
100
100
Figure 4.4. Relationship between penis tip length (top) or penis cuff diameter (bottom) and snout-vent length (SVL) of juvenile male Morelet's crocodiles from two habitats in northern Behze. A significant relationship existed between SVL and both measures of phallus size at both sites.
129
E E
c 0)
w "E 0) Q.
30
25
20
15
10
.2 'E o 0.
Mew River Watershed Gold Button Lagoon
r' = 0.60, p < 0.0001
r' = 0.90, p = 0.004
60
« "S E re O
o
40
J= 30
20
10
^-O
_ i —
80 90 100 110 120 130 140 150 160
— r = 0.68, p < 0.0001
.— r = 0.64, p = 0.06
— - - O
••
80 90 100 110 120 130 140 150 160
Snout-Vent Length (cm)
Figure 4.5. Relationship between penis tip length (top) or penis cuff diameter (bottom) and snout-vent length (SVL) of adult male Morelet's crocodiles from two habitats in northern Belize. A significant relationship existed between SVL and tip length at both sites and between SVL and cuff diameter at New River Watershed but not Gold Button Lagoon.
130
10
E E
^ ^ New River Watershed 1 1 Gold Button Lagoon
E E
Tip Length Cuff Diameter
Phallus Characteristic
Figure 4.6. Mean (±SE) phallus size of male Morelet's crocodiles sampled from New River Watershed and Gold Button Lagoon in northern Belize. No significant (p < 0.05) difference in either morphometric was observed between sites.
131
30
25
O) 20 C
0) c o
(0
o 0)
15
20
:=> 15
c 0) c o
(A
o (A a>
10
• New River Watershed
O Gold Button Lagoon
80
Juveniles
r =0.11, p = 0.01
r' = 0.05, p = 0.45
10 20 30 40 50 60 70 80 90
r* = 0.01, p = 0.87
r' = 0.88, p = 0.22
•
•
•
o ^ ^ ^ o • • • • ^ 1 - ' - • — ' — ' — 1 — ^
^ ^ ^ •
Adults
•
•
• o
90 100 110 120 130 140
Snout-Vent Length (cm)
Figure 4.7 Relationship between snout-vent length (SVL) and plasma testosterone (T) concentration in juvenile (top) and adult (bottom) male Morelet's crocodiles from two habitats in northem Belize. A positive SVL-T relationship was observed only in juveniles from New River Watershed.
132
10
c o
V) 4 o 0)
^^M New River Watershed I I Gold Button Lagoon
Juveniles Adults
Size Class
Figure 4.8. Mean (±SE) plasma testosterone (T) concentrations in male Morelet's crocodiles sampled from New River Watershed and Gold Button Lagoon in northern Belize. Numbers above bars indicate the number of animals sampled per site within a group. Bars with different superscripts are significantly different. An asterisk above a bar indicates a significant difference within that pair.
133
14
—s 12
E E,
O) c 0)
1= 6 Q.
"E o>
Q.
10
8
2 -
30
0)
0> E re Q
3 o w 'E 0) Q.
20
iS 15
10
• New River Watershed O Gold Button Lagoon
r'= 0.10, p = 0.02
r'= 0.01, p = 0.71
10 15 20 25 30
r = 0.16, p = 0.002
> = 0.02, p = 0.67
0 5 10 15 20
Testosterone (ng/ml)
25 30
Figure 4.9. Relationship between plasma testosterone (T) concentration and penis tip length (top) and cuff diameter (bottom) in juvenile Morelet's crocodiles from two habitats in northern Belize. Significant relationships were observed between T and both measures of phallus size at New River Watershed but not Gold Button Lagoon. However, these significant relationships disappear if the one individual with the exceptionaUy high T concentration (27.25 ng/ml) is removed from the analysis.
134
E E
c 0)
40
30
20
.2 "E « 10
• New River Watershed O Gold Button Lagoon
i' = 0.03, p = 0.75 r = 0.98, p = 0.09
10 12 14
E E •
0) +rf 0) E re O ^ 3
o
nis
0) Q.
50
40
.10
20
10
1
" O ' -^"'
3
r =
r =
0.0002, p = 0.98
0.95, p = 0.15
• a
• ^, ,^ ^ ''^
, , 1 , , , . 1 ,
•
•
.
• •
•
•
4 6 8 10
Testosterone (ng/ml)
12 14
Figure 4.10. Relationship between plasma testosterone (T) concentration and penis tip length (top) and cuff diameter (bottom) in adult male Morelet's crocodiles from two habitats in northern Belize. No significant relationship was observed between T and either measure of phallus size.
135
CHAPTER V
CONCLUSIONS
Study Summary
Over the last 20 years, evidence of population declines and reproductive
impairment in American alligators (Alligator mississppiensis^ in Florida, USA has
increased concerns over the effects of endocrine-disrupting contaminants (EDCs) on
wildlife and stressed the importance and utility of reptiles, particulariy crocodilians, as
focal species in the field of ecotoxicology (Heinz et al., 1991; Jennings et al., 1988;
Woodward et al., 1993; Guillette et al., 1994, 1995a, 1996, 1997, 1999a,b, 2000, 2002;
Rice et al., 1996; Vonier et al, 1996; Grain and Guillette, 1997, 1998; Grain et al., 1997,
1998. 2000; Pickford et al., 2000; Milnes et al., 2001, 2002a; Hewitt et al., 2002).
Although various environmental contaminants have been found in eggs and tissues of
crocodilians worldwide, no studies have yet investigated endpoints of endocrine
disruption in wild crocodilians outside of Florida (see Rainwater et al, 2002; Chapter I).
The primary objective of this dissertation was to address this data gap by examining
ecotoxicological endpoints in another crocodilian species living in habitats contaminated
with EDCs, and where appropriate, compare results from this study with those observed
for alligators in Florida.
The focal species for this study was Morelet's crocodile (Crocodylus moreletii).
an endangered, freshwater crocodile found in Mexico, Guatemala, and Belize
(Groombridge, 1987; Lee, 1996; Ross, 1998). During a pilot study in 1995, multiple
organochlorine (OC) pesticides considered to be EDCs were found in eggs of Morelet's
crocodiles from three localities in northem Belize (Rainwater et al., 2002, Rainwater et
al., unpublished data). Based on these findings and previous data from Florida showing
egg contamination, population declines, and reproductive abnormalities in alligators
exposed to many of the same chemicals (see Guillette et al., 2000), a multi-year study
was initiated to examine various endpoints of contaminant exposure and response in
Morelet's crocodiles living on contaminated and reference sites in northern Belize. Gold
Button Lagoon, a man-made lagoon from which contaminated crocodile eggs were
136
collected in 1995, was selected as the contaminated site for this study, while New River
Watershed, a more remote site with fewer anthropogenic inputs than Gold Button
Lagoon, was selected as the reference site.
Specifically, this dissertation examined one endpoint of exposure (Chapter II) and
two endpoints of response (Chapters III and IV) to EDCs in Morelet's crocodile. First,
plasma vitellogenin induction was examined as a biomarker of exposure to xenobiotic
estrogens in male and immature crocodiles inhabiting contaminated sites in Belize.
Vitellogenin is an egg-yolk precursor protein expressed in all oviparous and
ovoviviparous vertebrates (Palmer and Palmer, 1995). Males and juveniles normally
have no detectable concentration of vitellogenin in their blood but can produce it
following stimulation by an exogenous estrogen, such as an EDC (Palmer and Palmer,
1995). Thus, the presence of vitellogenin in the blood of male and immature crocodiles
can serve as an indicator of exposure to estrogen-mimicking chemicals. Of 358 males
and juvenile females sampled in this study, no vitellogenin induction was observed,
suggesting these animals were likely not exposed to estrogenic xenobiotics. However,
many of the animals sampled were later found to contain OC pesticides in their caudal
scutes (DeBusk, 2001), confirming they had in fact been exposed to OCs (and EDCs).
Previous researchers have stressed that vitellogenin induction is a measure of a biological
effect, not merely the presence of a contaminant in the body of an animal (Palmer and
Palmer, 1995). Our results support this notion, and suggest plasma vitellogenin induction
may still serve as a reliable biomarker of estrogen exposure in crocodilians, but the lack
of a vitellogenic response should not necessarily be interpreted as an indication that no
exposure or other contaminant-induced biological response has occurred. Numerous
crocodiles sampled in this study contained OCs in their scutes but did not exhibh
vitellogenin induction. The lack of a vitellogenic response in these animals may be due
to several factors including insufficient contaminant concentrations to induce
vitellogenesis, no affinity of these particular compounds for the Morelet's crocodile
estrogen receptor, or antagonism among xenobiotics present in crocodile tissue.
Second, plasma steroid hormone concentrations were examined as an endpoint of
response to EDC exposure in crocodiles from the two study sites. The selection of this
137
endpoint was based on numerous studies reporting altered concentrations of estradiol-17p
(E2) and testosterone (T) in alligators from Lake Apopka and other contaminated lakes in
Florida (Guillette et al., 1994, 1996, 1997, 1999a; Grain et al., 1998; Milnes et al.,
2002a). In the present study, few inter-site differences in plasma hormone concentrations
were noted. No significant differences in plasma E2 concentrations were detected
between sites. However, large juvenile males and females from the contaminated site
exhibited significantly reduced plasma T concentrations compared to large juvenile males
and females from the reference site, respectively. This finding was consistent with
results from previous studies on alligators in Florida (Guillette et al., 1994, 1996, 1997,
1999a; Grain et al., 1998). No other inter-site differences in hormone concentrations
were observed. Relationships between body size and hormone concentrations were
variable and showed no clear pattern. It was discovered late in the study that New River
Watershed (reference site) exhibited a contaminant profile similar to that observed at
Gold Button Lagoon (contaminated site), with multiple OCs detected at similar
concentrations in sediments, crocodile eggs, and crocodile tail scutes at both sites (Wu et
al., 2000a; DeBusk, 2001). With the lack of a suitable reference site, it is thus unclear if
the steroid hormone concentrations observed in this study are within the normal range
exhibited by Morelet's crocodiles living in non-contaminated habitats or if they are
altered in some way (e.g., elevated, depressed). In addition, it is also unclear if inter-site
differences in plasma T are the result of exposure to EDCs, natural variation, one or more
undetermined factors (e.g., stress), or a combination of these factors.
Third, male phallus size was examined as a second endpoint of response to EDC
exposure in crocodiles from the two study sites. Concurrent with reductions in plasma T
concentrations, male alligators from Lake Apopka and other contaminated lakes in
Florida have exhibited smaller phallus size compared to animals from a reference lake
(Guillette et al., 1994, 1996, 1999a,b; Pickford et al., 2000). Researchers speculate that
abnormal hormone concentrations during critical eariy life stages may affect anatomical
structures dependent on these hormones for proper growth and development (i.e.,
genitalia) (Guillette et al., 2000). The DDT metabohte p,p '-DDE is one of the primary
contaminants of concern at Lake Apopka and has been shown to be anti-androgenic in
138
laboratory mammals, inhibiting normal androgen function in vivo (Kelce et al., 1995).
This persistent OC has been detected in alligator eggs and serum from Lake Apopka
(Heinz et al., 1991; Guillette et al., 1999b), suggesting its potential role in the •
reproductive anomalies observed in juvenile males (GuiUette et al., 1994, 1996, 1999a,b;
Pickford et al., 2000). p,p '-DDE has also been detected in Morelet's crocodile eggs and
scutes in Belize (Wu et al., 2000a,b), confirming EDC exposure in maternal females,
neonates, juveniles, and other adults. Thus, in the present study, male crocodile phallus
size and plasma T concentrations were examined as endpoints of response top,p'-DDE
exposure as well as exposure to other contaminants. No differences in mean phallus size
were observed between sites, whereas mean plasma T concentrations in juveniles from
Gold Button Lagoon were again (see Chapter III) significantly reduced compared with
those from New River Watershed. Juvenile males from both sites exhibited positive
relationships between body size and phallus size. However, while juvenile males from
New River Watershed also exhibited positive relationships between plasma T and body
size and plasma T and phallus size, no such relationships were observed for juveniles
from Gold Button Lagoon. For adults, no significant inter-site differences were observed
in phallus size, plasma T concentrations, or relationships between plasma T and body size
or phallus size. Due to the similarity in contaminant profiles between New River
Watershed and Gold Button Lagoon, it is unclear whether phallus size and plasma T
concentrations observed in crocodiles from these two sites are normal or altered by some
stressor (e.g., endocrine-dismpting chemicals). Thus, the biological significance of the
few site differences observed in this study is difficult to interpret.
Comparison of this Study with Studies on Florida Alligators
Comparisons of data obtained in this study with those reported for Florida
aUigators reveal both similarities and differences in results. This is largely due to the fact
that multiple studies examining the same endpoints have been conducted at Lake
Apopka, and in many cases conflicting results have been observed (Table 5.1, Table 5.2).
The Florida studies have primarily examined juvenile animals, so comparisons of data
from those studies to data from crocodiles in Belize are confined to that size group (Table
139
5.1, Table 5.2). For juvenile males, seven of the ten endpoint measures at Lake Apopka
exhibit conflicting results (Table 5.1). Thus, by virtue of there being more than one
observation of the same endpoint at Lake Apopka, the result of a particular measured
endpoint from Gold Button Lagoon has a high likelihood of agreeing with one of the two
or three possible results for that endpoint measured at Lake Apopka. For example, two
studies have shown elevated E2 concentrations in juvenile male alligators from Lake
Apopka compared to Lake Woodruff (GuiUette et al., 1999a; Milnes et al., 2002a), while
three other studies have shown no difference in E2 concentrations between the two lakes
(Guillette et al.. 1997, 1999b; Grain et al., 1998). For Morelet's crocodiles in Belize, no
difference in plasma E2 concentrations was observed between Gold Button Lagoon and
New River Watershed, thereby agreeing with the resuhs of three studies on Lake Apopka
but disagreeing with two. This pattern follows for six of the remaining nine endpoints
(Table 5.1). For each of the remaining three endpoints (phallus tip length, phallus cuff
diameter, body size-cuff diameter relationship), only one result was obtained in all
studies that examined that endpoint at Lake Apopka. In all three cases, the result of the
corresponding endpoint examined at Gold Button Lagoon is different (e.g., reduced
phallus tip length at Lake Apopka, no reduction at Gold Button Lagoon). Overall, the
primary difference between males at these two sites is that Apopka juveniles exhibit
reduced phallus size while Gold Button juveniles do not. In addition, males at Gold
Button Lagoon exhibit a positive body size-phallus cuff diameter relationship, while no
such relationship is observed at Lake Apopka. For juvenile females, the only difference
in endpoint measurements between contaminated sites is that females at Gold Button
Lagoon exhibit reduced plasma T concentrations, while females at Lake Apopka do not
(Table 5.2).
In contrast to the contaminated sites, few conflicting results exist for endpoints
measured at the reference sites. For male juveniles, animals at New River Watershed and
Lake Woodmff exhibit similar responses in all endpoints measured. However, when only
larger juvenile males from New River Watershed are included in the analysis (Chapter
III), no positive T-body size relationships are observed (Table 5.1). This represents the
single difference in endpoint measures between the two reference sites and is the only
140
instance in the present study in which conflicting results exist for the same endpoint. For
female juveniles, no differences in endpoint measurements are observed between
reference sites.
Uncertainties
Multiple uncertainties associated with the present study make many of the results
difficult to interpret and threaten the validity of inter-study (Belize to Florida)
comparisons. The primary confounding factor in this study is the lack of a reference site.
Similarity in contamination profiles between New River Watershed and Gold Button
Lagoon precludes the comparison of endpoint measurements to legitimate reference
values. Thus, it is difficult to determine the toxicological significance of any inter-site
differences observed in this study. In addition, in instances where no inter-site difference
was observed for a particular endpoint (e.g., plasma E2 concentration; Table 5.1), it is
difficult to discem whether the lack of a difference indicates that that particular endpoint
measurement is normal (i.e., unaltered) at both sites or if the endpoint is altered in some
way, but to the same degree, at both sites. In tum, uncertainty as to whether endpoint
measures in this study are altered or unaltered may render comparisons of these data to
ecotoxicological data on other crocodilians (e.g., Florida alligators) less meaningful.
The toxicological significance of OC concentrations in sediments at Gold Button
Lagoon and New River Watershed is unknown. Although OC concentrations detected at
both sites are well below protective levels established for humans in Texas, USA, they
exceed those considered to be ecological benchmarks protective of benthic organisms
(Table 5.3). Due to the paucity of toxicity data pertaining to OC effects on crocodilians
and other reptiles, it is unknown what concentrations pose a risk to these animals, and
extrapolations based on data from other organisms may be inappropriate. Mean
concentrations of p,/?'-DDE in crocodile eggs from Belize are among the lowest reported
for any crocodilian species (Table 5.4). However, despite low levels of OC
contamination at New River Watershed and Gold Button Lagoon compared to other areas
of the worid, potential chemical-induced effects on Morelet's crocodiles should not be
ignored. Currently it is unknown at what OC concentrations endocrine disruption may
141
occur in crocodilians, but recent research suggests low concentrations (e.g., 100 ppb p,p '-
DDE) similar to those detected at the Belize study sites may cause reproductive
impairment in alligators (Milnes et al., 2002b).
Another confounding factor in the present study is the lack of previous research
on Morelet's crocodiles focusing on the response endpoints examined in this study. In
the absence of a legitimate reference site, basic information on vitellogenin induction,
plasma steroid hormone concentrations, and male phallus size in other Morelet's
crocodile populations would be particulariy useful for comparative purposes. Although
such data on actual concentrations or morphometries may be limited in their applicability
due to multiple sources of inter-study variation, information on seasonal and age- and
sex-specific pattems in these endpoints in other populations might aid in interpreting the
results observed in this study. However, apart from this study, data on the endocrinology
of Morelet's crocodile is non-existent.
A third uncertainty associated with this study is the relationship between
contaminant exposure and the magnitude of response, if any, in crocodiles at the two
study sites. Due to the current endangered status of Morelet's crocodile, collection of
intemal tissues for contaminant residue analysis is not feasible. Caudal scutes, collected
from crocodiles as a by-product of the marking procedure (Jennings et al., 1991), have
been analyzed for OC pesticides and have confirmed OC exposure in animals at both
sites (DeBusk, 2001). While these samples provide valuable qualitative data on crocodile
exposure, their utility as indicators of OC concentrations in internal tissues is unknown.
Jagoe et al. (1998) found that caudal scutes from alligators were relatively poor predictors
of mercury in intemal tissues, but added that these samples may provide a rough
estimation of contamination in populations without sacrificing animals. Our observations
in this study indicate an allometric relationship between crocodile size and fat content in
caudal scutes. No fat is visibly present in the scutes of hatchlings and smaU juveniles, but
fat content increases with size such that substantial fat cores are present in the scutes of
large aduhs (Rainwater, personal observation). Due to the size-specific variation in
crocodile scute fat content and the inability to investigate the relationships between OC
142
concentrations in scutes and intemal tissues, the relationship between OC exposure and
biomarker response in these animals remains unclear.
Natural variability between sites and site-specific influences of other stressors
(e.g., disease, injury, malnutrition) may have also influenced the endpoints measured in
this study. In addition, capture stress and inter-site variability in the timing of blood
collection may have affected the steroid hormone concentrations in crocodiles sampled in
at each locality. Lastiy, slight differences in size class designations, possible differences
in species sensitivity to the endpoints measured, and temporal differences in sample
collection (e.g., different months) may reduce the validity of comparisons between the
present study and similar studies on Florida alligators.
Future Research Directions
In addition to those examined in the present study, other endpoints of endocrine
disruption including gonadal morphology and gonadal aromatase activity have been
examined in Florida alligators, and animals from Lake Apopka have exhibited alterations
in these endpoints (Grain and Guillete, 1998). Ideally, future research on Morelet's
crocodiles would closely examine these endpoints as well. However, due to the
endangered status of this species, invasive or lethal endpoints are not feasible. Most
importantly, future studies on the ecotoxicology of Morelet's crocodiles in Belize should
attempt to locate legitimate reference populations for comparative purposes. If a non-
contaminated site cannot be found, efforts should shift to finding a highly contaminated
site to compare to Gold Button Lagoon, New River Watershed, or other lesser-
contaminated sites. Once such sites are located, consistent long-term monitoring of
individual-level endpoints examined in this study as well as population-level endpoints
including egg viability, sex-ratios, neonatal survival, and population density should be
employed. Data on many of these population-level endpoints were collected from 1997
to 2000 and are currentiy being examined. These data will provide valuable insight into
the overall status of crocodiles at both sites. However, conclusions drawn from these
data must acknowledge the uncertainties discussed above, particularly the lack of a
reference site.
143
Conclusions
In general, the results of this dissertation indicate few or no effects of EDC
exposure on Morelet's crocodiles inhabiting contaminated wetlands in northern Belize.
The only major difference observed between crocodiles from New River Watershed and
Gold Button Lagoon in this study was that juveniles from Gold Button Lagoon exhibited
lower plasma T concentrations than juveniles at New River Watershed. This same
difference has been consistently observed in juvenile male alligators from Lake Apopka
compared to Lake Woodruff. However, in the Florida studies, reduced plasma T was
also concurrent with reduced phallus size, suggesting a potential link between reduced T
and alterations in anatomical stmctures (e.g., male phallus) dependent on androgens for
proper growth and development (Guillette et al., 2000). In the present study, no inter-site
differences in phallus size were observed. This suggests the lower T concentrations in
Gold Button Lagoon juveniles, whether contaminant-induced or not, may not be
biologically significant. In addition, no inter-site differences in plasma E2 concentrations
were observed in this study, and vitellogenin induction was not observed in any of the
358 male or juvenile female crocodiles examined. These results suggest no or minimal
alteration of E2 concentrations in crocodiles from the two sites, despite the fact that many
of these animals have been exposed to environmental contaminants considered to be
xenobiotic estrogens. It is possible that the concentrations of EDCs to which Morelet's
crocodiles in northem Belize are exposed are insufficient to influence these endpoints,
while EDC concentrations at Lake Apopka are sufficiently high to induce an effect.
Indeed, mean p,p '-DDE concentrations in alligator eggs from Lake Apopka are 27- to 45-
fold higher than those observed in eggs from Gold Button Lagoon or New River
Watershed (Heinz et al., 1991; Wu, 2000).
Currentiy, Morelet's crocodile populations in northem Behze appear to have
recovered from past over-harvesting, and threats related to habitat loss and human
exploitation appear minimal (Piatt and Thorbjamarson, 2000). Researchers have recently
speculated that although Morelet's crocodiles in northem Belize seemingly face no
immediate threats, exposure to environmental contaminants may present a subtie yet
significant long-term threat to populations in certain areas (Rainwater et al., 1998; Piatt
144
and Thorbjarnarson, 2000). Results of the present study provide little evidence of
contaminant-induced effects on crocodiles from two polluted habitats in northem Belize.
However, multiple confounding factors and uncertainties encountered in this study make
inter-site and inter-study (crocodile to alligator) comparisons difficult and some results
equivocal. Thus, the potential effects of EDCs and other contaminants on crocodiles
inhabiting these sites should not be assumed to be negligible. Long-term studies are
essential to adequately assess the effects of EDCs on crocodilian populations, as many of
the contaminant-induced effects are organizational in nature, occurring during embryonic
development but not appearing until later in life (Guillette et al., 1995b).
145
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150
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