BIOLOGY AND CONSERVATION OF THE DIAMONDBACK TERRAPIN,
MALACLEMYS TERRAPIN PILEATA, IN ALABAMA






by
ANDREW THOMAS COLEMAN

THANE WIBBELS, COMMITTEE CHAIR
KEN MARION
DAVID NELSON
WILLEM ROOSENBURG
ROBERT THACKER













A DISSERTATION

Submitted to the graduate faculty of The University of Alabama at Birmingham,
in partial fulfillment of the requirements for the degree of
Doctor of Philosophy

BIRMINGHAM, ALABAMA

2011






































Copyright by
Andrew Thomas Coleman
2011
iii

BIOLOGY AND CONSERVATION OF THE DIAMONDBACK TERRAPIN,
MALACLEMYS TERRAPIN PILEATA, IN ALABAMA
ANDREW THOMAS COLEMAN
BIOLOGY
ABSTRACT
The diamondback terrapin is the only obligate estuarine turtle in North America and is
considered an integral member of the salt marsh ecosystem. Unfortunately, many
populations throughout the terrapin’s range have experienced declines due to past
overexploitation and have been unable to rebound due to current threats, including crab
trap mortality, habitat degradation, nest predation, and road mortality. The current study
was the first comprehensive study examining various population and conservation
parameters of Mississippi diamondback terrapins in Alabama. Through various field
survey methods conducted in numerous salt marshes along the Gulf Coast of Alabama, it
was concluded that Cedar Point Marsh supported the largest aggregation of terrapins in
Alabama, and the beach bordering Cedar Point Marsh represented the most important
nesting habitat. However, population estimates indicated a significant size reduction
from historical levels, and crab trap mortality and nest predation were identified as major
threats currently impacting this population. The major population decline in Alabama
was also reflected in the terrapins’ genetic diversity, whose low diversity was similar to
other sampled terrapin populations. By-catch reduction devices were shown to be an
effective management tool to prevent terrapin entry into crab traps, although decreases in
crab capture were observed. Obtaining eggs from nesting females to help offset nest
predation allowed investigations of female allocation strategies and post-emergence
orientation behavior of hatchlings. Larger and older females produced larger eggs and
iv

hatchlings, but the advantage of larger hatchling sizes was not detected with the potential
fitness indicators examined in this study. The consequences of high levels of road
mortality, which would theoretically result in removing older females, were examined,
and the Alabama population, which does not experience high road mortality, produced
larger eggs than the Georgia population that does suffer from this threat. Terrapin
hatchlings utilized the same orientation cues as sea turtle hatchlings but moved toward
the higher marsh areas rather than open water. This underscored the necessity of healthy
marsh habitat adjacent to nesting beaches. The initiation of this long-term dataset is
crucial in developing optimal management strategies for ensuring the future survival of
diamondback terrapins in Alabama.

conservation—bycatch—nest predation—habitat loss—microsatellites—turtles

v
DEDICATION



To my mother and father, Connie and Mark Coleman, whose love and support made this
dissertation possible. They never wavered when their youngest son decided to study
salamanders and turtles when other sons and daughters pursued normal 9-5 careers.

To the residents of the Gulf Coast who truly care for their surrounding natural beauty and
whose resilience and determination are unsurpassed.
vi

ACKNOWLEDGEMENTS
I first would like to thank the members of my committee: Thane Wibbels, Ken
Marion, Bob Thacker, Willem Roosenburg, and David Nelson. The insight and advice
they provided with the research and dissertation have been greatly appreciated. I would
also like to acknowledge the major funding resources for this project: Alabama Center
for Estuarine Studies and Alabama Department of Conservation and Natural Resources. I
also received funding from University of Alabama at Birmingham Biology Department,
Alabama Academy of Science, Birmingham Audubon Society, and University of
Alabama Graduate Student Association. I would like to thank our collaborators: John
Dindo, Kristen Hart, and Jordan Gray. The support I received from the Department
Chair, Bud Fischer, and Graduate Program Director, Steve Watts, was extremely helpful.
I have also been the recipient of great help from the Biology Department office staff and
lab coordinators. Additionally, I would like to thank my friends who endured countless
turtle anecdotes and enjoyed sharing their favorite turtle jokes. And, I would be remiss if
I did not thank Mrs. Kathy Wibbels for all of the time and effort she provided to me and
the rest of the Wibbels’ lab. I consider myself lucky to have had the great opportunity to
study under Thane Wibbels and Ken Marion whose guidance made me a better scientist
and whose stories in the field made my time here just that much more enjoyable. Finally,
this degree would have been impossible without the love, compassion, and care given
freely by my parents, Mark and Connie Coleman, and the rest of my family including
Jeremy, Julie, Julian, and Rebecca Coleman.
vii

TABLE OF CONTENTS
Page
ABSTRACT ....................................................................................................................... iii
DEDICATION .....................................................................................................................v
ACKNOWLEDGEMENTS ............................................................................................... vi
LIST OF TABLES ............................................................................................................. ix
LIST OF FIGURES .......................................................................................................... xii
DEVELOPMENT OF RECOVERY PROGRAMS FOR THE DIAMONDBACK
TERRAPIN: EVALUATION OF THREATS AND CONSERVATION STRATEGIES
..............................................................................................................................................1

POPULATIN ECOLOGY OF THE DIAMONDBACK TERRAPIN (MALACLEMYS
TERRAPIN PILEATA) IN ALABAMA .............................................................................37

POPULATION GENETIC EVALUATION OF THE DIAMONDBACK TERRAPIN
(MALACLEMYS TERRAPIN PILEATA) IN ALABAMA .................................................72

EFFECT OF BYCATCH REDUCTION DEVICES (BRDS) ON THE CAPTURE OF
DIAMONDBACK TERRAPINS (MALACLEMYS TERRAPIN) IN CRAB TRAPS IN
AN ALABAMA SALT MARSH ......................................................................................96

EVALUATION OF MATERNAL INVESTIMENT, EGG SIZE AND HATCHLING
FITNESS IN THE DIAMONDBACK TERRAPIN (MALACLEMYS TERRAPIN) .......118

ORIENTATION OF DIAMONDBACK TERRAPIN (MALACLEMYS TERRAPIN)
HATCHLINGS ON A NATURAL NESTING BEACH.................................................147

FINAL DISCUSSION .....................................................................................................163

IACUC APPROVAL FORM ...........................................................................................177

HEAD SURVEY RESULTS ...........................................................................................178
viii

Page
DEPREDATED NEST SURVEY RESULTS .................................................................184
CAPTURE DATA RESULTS .........................................................................................203
ix

LIST OF TABLES
Table Page
INTRODUCTION
1 State regulations regarding terrapin status, bycatch and take ......................................36

POPULATIN ECOLOGY OF THE DIAMONDBACK TERRAPIN (MALACLEMYS
TERRAPIN PILEATA) IN ALABAMA
1 Morphological measurements taken from each captured terrapin ...............................44
2 Heads per unit effort by location in 2004-2005 ...........................................................46
3 Heads per unit effort for various salt marshes from 2006-2010 ..................................47
4 Heads per unit effort for Cedar Point Marsh from 2006-2010 ....................................47
5 Catch per unit effort for each trapping method by year ...............................................48
6 Number of juveniles, females, and males captured in Cedar Point Marsh by year .....48
7 Average morphological measurements by sex ............................................................49

POPULATION GENETIC EVALUATION OF THE DIAMONDBACK TERRAPIN
(MALACLEMYS TERRAPIN PILEATA) IN ALABAMA
1 Sampled populations included in genetic analysis and sample sizes ...........................77
2 Microsatellite primers developed by King and Julian (2004) and used in Hart (2005)
as well as the current study .........................................................................................78
3 Observed and expected heterzygosities for the 12 sampled loci in the Alabama
population ....................................................................................................................81

x

Table Page
4 Analyses of Molecular Variance results for the three different groupings ..................81
5 Pairwise population Fst comparisons..........................................................................81

EFFECT OF BYCATCH REDUCTION DEVICES (BRDS) ON THE CAPTURE OF
DIAMONDBACK TERRAPINS (MALACLEMYS TERRAPIN) IN CRAB TRAPS IN
AN ALABAMA SALT MARSH
1 Morphological measurements obtained from each captured terrapin ........................102
2 Average morphological measurements of terrapin bycatch .......................................102

EVALUATION OF MATERNAL INVESTIMENT, EGG SIZE AND HATCHLING
FITNESS IN THE DIAMONDBACK TERRAPIN (MALACLEMYS TERRAPIN)
1 Averages of various egg and hatchling parameters obtained from 22 clutches .........126
2 Results from Pearson’s correlations of adult female weight with both adult female age
and plastron length .....................................................................................................126
3 Results from Pearson’s correlations between the adult female parameters and egg
parameters ..................................................................................................................127

4 Results from Pearson’s correlations between adult female parameters and initial
hatchling parameters ..................................................................................................128

5 Results from Spearman’s correlations between adult female parameters and potential
hatchling fitness indicators ........................................................................................128

6 Results from Pearson’s and Spearman’s correlations between incubation temperature
and initial hatchling size and hatchling growth parameters .......................................129

7 Results of two-tailed t tests of comparisons of egg and initial hatchling parameters
between Cedar Point Marsh, AL, and Tybee Island, GA ..........................................129


xi

Table Page
ORIENTATION OF DIAMONDBACK TERRAPIN (MALACLEMYS TERRAPIN)
HATCHLINGS ON A NATURAL NESTING BEACH
1 Results of Chi-square goodness of fit analyses examining orientation preferences of
diamondback terrapin hatchlings ...............................................................................152

2 Results of Chi-square goodness of fit analysis of orientation behavior on the natural
nesting beach surrounding Airport Marsh .................................................................152








xii

LIST OF FIGURES
Figure Page
POPULATIN ECOLOGY OF THE DIAMONDBACK TERRAPIN (MALACLEMYS
TERRAPIN PILEATA) IN ALABAMA
1 Salt marshes along the Gulf Coast of Alabama that were sampled from 2004-2010 ..61
2 Number of depredated nests by location per year ........................................................62
3 Number of captured terrapins in 2005 with modified crab traps by location ..............63
4 Number of captured terrapins with pitfall traps by location per year ..........................64
5 Number of depredated nests found at various locations from 2006-2010 ..................65
6 Summary of depredated nest surveys completed at Cedar Point Marsh from 2006-
2010..............................................................................................................................66
7 Number of captures with modified crab traps by location per year .............................67
8 Number of captures with pitfall traps in Cedar Point Marsh per year .........................68
9 Number of captures in Cedar Point Marsh with trawling by year ...............................69
10 Captures of age classes by sex from 2006-2010 in Cedar Point Marsh .....................70
11 Population estimates for the population not including hatchlings as well as nesting
females .......................................................................................................................71

POPULATION GENETIC EVALUATION OF THE DIAMONDBACK TERRAPIN
(MALACLEMYS TERRAPIN PILEATA) IN ALABAMA
1 Mean allelic diversity of the sampled populations across 12 loci ...............................91
2 Mean allelic range of the sampled populations across 12 loci .....................................92
xiii

Figure Page
3 Observed heterozyogsities of the sampled locations across 12 loci ............................93
4 Mean M ratio for the sampled populations across 12 loci ...........................................94
5 Evaluation of K (2-7) populations using Bayesian clustering method ........................95

EFFECT OF BYCATCH REDUCTION DEVICES (BRDS) ON THE CAPTURE OF
DIAMONDBACK TERRAPINS (MALACLEMYS TERRAPIN) IN CRAB TRAPS IN
AN ALABAMA SALT MARSH
1 Average weekly capture of terrapin bycatch by traps fitted with BRDs and non-BRD
traps ............................................................................................................................114
2 Catch per unit effort (C.P.U.E.) of terrapins caught in traps fitted with BRDs and non-
BRD traps...................................................................................................................115
3 Catch per unit effort (C.P.U.E.) of marketable-sized crabs caught with traps fitted
with BRDs and non-BRD traps..................................................................................116

4 Catch per unit effort (C.P.U.E.) for total crabs caught with traps fitted with BRDs and
non-BRD traps ...........................................................................................................117

EVALUATION OF MATERNAL INVESTIMENT, EGG SIZE AND HATCHLING
FITNESS IN THE DIAMONDBACK TERRAPIN (MALACLEMYS TERRAPIN)
1 Pictorial representation of the predictions of the optimal egg size theory (Smith and
Fretwell, 1974) ...........................................................................................................142
2 Scatter plot of clutch sizes by adult female weight....................................................143
3 Scatter plot of average egg weight by clutch by adult female weight ......................144
4 Scatter plot of average hatchling weight by clutch by adult female weight ..............145
5 Pictorial representation of the overall results from the current study ........................146

xiv

Figure Page
ORIENTATION OF DIAMONDBACK TERRAPIN (MALACLEMYS TERRAPIN)
HATCHLINGS ON A NATURAL NESTING BEACH
1 Diagram of orientation arena on natural nesting beach ..............................................158
2 2008 results of diamondback terrapin hatchling orientation behavior on a natural
nesting beach surrounding Cedar Point Marsh ..........................................................159

3 2009 results of diamondback terrapin hatchling orientation behavior on a natural
nesting beach surrounding Cedar Point Marsh ..........................................................160

4 2009 results of diamondback terrapin hatchling orientation behavior on a natural
nesting beach surrounding Airport Marsh ..................................................................161

5 2010 results of diamondback terrapin hatchling orientation behavior on a natural
nesting beach surrounding Airport Marsh .................................................................162


1

CHAPTER 1
INTRODUCTION

DEVELOPMENT OF RECOVERY PROGRAMS
FOR THE DIAMONDBACK TERRAPIN:
EVALUATION OF THREATS AND CONSERVATION STRATEGIES

Andrew Coleman, Thane Wibbels, Ken Marion
Department of Biology, University of Alabama at Birmingham
Birmingham, AL, 35242









Submitted to Chelonian Conservation and Biology
Format adapted for dissertation
2

Abstract
The survival status of the diamondback terrapin, the only North American estuarine
turtle, is being impacted by a wide variety of threats. Terrapins were once abundant in
the salt marshes lining the Atlantic and Gulf coasts of the U.S. They represented a
valuable economic resource that was exploited as a culinary delicacy. Although, they are
no longer subject to range wide commercial exploitation, terrapin populations have not
rebounded to historic levels. To the contrary, many populations have declined to the point
at which their survival requires protection and management through conservation
programs. Current threats, including habitat loss, crab trap and road mortality, and nest
predation, are significantly affecting terrapins throughout their range. The following
review discusses the wide variety of threats that are impacting terrapins and explores
conservation strategies that could mitigate these threats, and thus enhance the recovery of
terrapin populations.
Key Words
diamondback terrapins—conservation—habitat loss—bycatch—head-starting—
depredation—road mortality—crap trap
Introduction
The diamondback terrapin (Malaclemys terrapin) is of distinct ecological interest
because it is the only turtle, and one of the few reptiles, to exclusively inhabit bays,
estuaries, and salt marshes in North America (Brennessel, 2006). While it is an obligate
inhabitant of a brackish water environment, it belongs to the freshwater turtle family
Emydidae, and its closest relatives are the map turtles of the genus Graptemys (Lamb and
Osentoski, 1997). Despite their freshwater ancestry, terrapins have evolved certain
3

physiological and behavioral adaptations that allow them to thrive in these harsh brackish
water environments (Dunson, 1970; Gilles-Baillen, 1970; Gilles-Baillen, 1973; Robinson
and Dunson, 1976; Davenport and Macedo, 1990; Davenport and Magill, 1996;
Brennessel, 2006). It is also of ecological interest because it is an integral part of the salt
marsh ecosystem representing a top level predator and potential keystone species (Tucker
et al., 1995; Silliman and Zieman 2001; Silliman et al., 2005; Gustafson et al., 2006). It
is known to feed on a variety of invertebrates including periwinkle snails of the genus
Littorina, blue crabs, Callinectes sapidus, and a variety of other mollusks, crustaceans,
and fishes (Tucker et al., 1995).
Diamondback terrapins have experienced a rich and diverse cultural history over the
past three centuries in North America (Carr, 1952; Brennessel, 2006; Hart and Lee, 2006;
Schaffer et al., 2008). This history indicates that they were once numerous throughout
most of their range and were used as an abundant food source as well as an important
economic resource. Interestingly, terrapins made a transition from a nuisance species to a
top culinary delicacy (Carr, 1952; Brennessel, 2006; Hart and Lee, 2006; Schaffer et al.,
2008). During the 1700’s terrapins were often considered nuisance by-catch and
represented an inexpensive food source. For example terrapins were fed to servants and
slaves in some coastal plantations as well as the Continental Army (Brennessel, 2006;
Hart and Lee, 2006; Schaffer et al., 2008). During the early 1800’s, terrapins made a
remarkable transition in popularity as stew and soup recipes proliferated and turned the
diamondback terrapin into a culinary delicacy. This popularity remained high and by the
mid 1800’s there was commercial harvesting in areas ranging from Maryland to Texas
(Brennessel, 2006). By the end of nineteenth and beginning of the twentieth centuries,
4

terrapin stew became almost a mandatory dish at upscale restaurants and social
gatherings. The demand for terrapin was so great during the mid 1800’s through the early
1900’s that natural stocks declined due to decades of overharvesting, and by the early
1900’s stocks in many areas were too low to sustain commercial harvest (Coker, 1920;
Babcock, 1926; Finneran, 1948; McCauley, 1945; Schaffer et al., 2008). In an effort to
meet the demand for terrapin, commercial aquaculture was initiated in the late 1800’s and
was enticing enough that a stocking program was initiated by the federal fisheries at
Beaufort, North Carolina (Hildebrand and Hatsel, 1926).
A decline in the demand for terrapins began in early 1900’s for a variety of reasons.
In addition to declining terrapin populations, the social and economic atmosphere
associated with World War I and then the Great Depression decreased the demand for
this high priced reptile, and Prohibition prevented the purchase of the sherry wine,
popular in terrapin soups and stews (Brennessel, 2006; Hart and Lee, 2006; Schaffer et
al., 2008). Collectively, these factors resulted in a distinct decline in the demand for
terrapins during the first few decades of the 20
th
century. Although some local
populations may have been wiped out by the decades of overharvesting, some began to
slowly recover due to the decreased harvesting (Coker, 1951; Hurd et al., 1979;
Brennessel, 2006; Hart and Lee, 2006; Schaffer et al., 2008). However, during the
1940’s and 1950’s the crab fisheries along the Atlantic and Gulf Coasts of U.S. began
large scale utilization of crab traps as an efficient method for harvesting crabs (Kennedy
et al., 2007). As the crab fishery flourished, the impact on terrapin populations
significantly increased due to incidental capture and mortality of terrapin in crab traps
(Roosenburg et al., 1997; Wood 1997; Dorcas et al., 2007).
5

As reviewed below, crab traps represent a major threat to terrapin populations and
have caused significant declines in populations throughout the Atlantic and Gulf coastal
waters of the U.S. Further, other threats including increased coastal development along
with heavy predator load on nesting beaches have also impacted terrapins in recent years
(Feinburg and Burke, 2003; Butler et al., 2004; Seigel, 1980). Collectively, such threats
have impacted terrapin populations throughout their range and have prevented their
recovery. The purpose of the current manuscript is to provide a comprehensive overview
of the major threats affecting the diamondback terrapin and review potential strategies for
enhancing their recovery, including the review of strategies implemented for other reptile
recovery programs when appropriate.
Habitat Loss
Although the range of terrapins extends along the majority of the Atlantic and Gulf
coasts of the United States, it exists as a relatively thin strip of estuarine habitat, (Carr,
1952). Further, the total amount of this estuarine habitat has been decreasing yearly.
According to a 2008 study, over 350,000 acres were lost in coastal watersheds in the
eastern U.S. between 1998-2004, and losses in the Gulf of Mexico region were 25 times
higher in that same time period (Stedman and Dahl, 2008). Of the designated wetland
types, salt marshes sustained the heaviest losses. Many of these losses can be attributed to
anthropogenic effects associated with coastal development (Stedman and Dahl, 2008;
Hartig et al., 2002). Dredging, filling, shoreline hardening, and a variety of other
activities alter natural processes and contribute to the loss of salt marsh habitat (Stedman
and Dahl, 2008) and diamondback terrapin nesting habitat (Roosenburg, 1991). As an
example, it has been estimated that 38% to 78% of salt marsh habitat associated with
6

islands in Jamaica Bay (New York City, NY) has been lost since 1974 (Hartig et al.,
2002). Compounding the loss of habitat from anthropogenic effects, loss of salt marsh
habitat can also be accelerated episodically by natural events as exemplified by Hurricane
Katrina in 2005 (Stedman and Dahl, 2008).
Impact of Global Climate Change on Terrapin Habitat
Another threat which could significantly impact the terrapin’s habitat is global
climate change (IPCC, 2007). Of particular importance, models project significant
increases in sea level and temperature over the 21
st
Century (IPCC, 2007). The potential
effects of global climate change on salt marshes have been discussed by a variety of
authors (Donnelly and Bertness, 2001; Simas et al., 2001; Hartig et al., 2002; Scavia et
al., 2002; Hughes, 2004; Najjar, et al., 2000). Sea level rise has the potential of
drowning salt marshes, but it has been suggested that is some cases, salt marshes may be
able to survive by accreting sedimentation vertically (Simas et al., 2001). Salt marshes
with large tidal ranges and high sediment transport would be predicted to more capable of
compensating for sea level increases, whereas those with small tidal ranges and low
sediment transport would be more susceptible to drowning (Simas et al., 2001).
Terrapins inhabit a relatively thin strip of estuarine environment lining the Atlantic
and Gulf coasts of the U.S., and in many cases, those habitats are now bordered by
coastal development. Therefore, if sea level rise negatively impacts salt marsh, this
would result in the narrowing or disappearance of terrapin habitats in some locations
(Donnelly and Bertness, 2001; Hartig et al., 2002; Hughes, 2004).
In addition to sea level rise, temperature change could also have a variety of impacts
on the salt marshes, including primary productivity, eutrophication, and dissolved oxygen
7

content (Najjar et al., 2000; Scavia et al., 2002), all of which could affect the ecology of
the diamondback terrapin. Global temperature change could also affect the reproductive
ecology of terrapins, including the timing of reproduction (Marion, 1982; Ganzhorn and
Licht, 1983; Mendonça, 1987) as well as sex determination. In the case of the terrapin’s
temperature-dependent sex determination, a 1.0 ºC increase temperature could result in a
shift from an approximate 1:1 sex ratio to the production of all female hatchlings
(Jeyasuria and Place, 1997; Wibbels, 2003), and current models suggest a temperature
increase ranging from 1.8 to 4.0 C during the 21
st
Century (IPCC, 2007). Thus, although
there are many immediate threats that are typically addressed when generating a recovery
strategy, global climate change could emerge as a major factor threatening the survival of
terrapin over the next century.
Habitat Quality and Habitat Pollution
In addition to the loss of salt marsh habitat, factors such as pollution, invasive
species, freshwater diversion, shoreline development, and loss of natural species can
adversely affect the structure and quality of the salt marsh ecosystem (Bertness et al.,
2002; Stedman and Dahl, 2008). For instance, shoreline development has been shown to
increase invasive species and decrease species diversity in adjacent salt marshes
(Bertness et al., 2002; Silliman and Bertness, 2004). The preservation of salt marsh
quality is of paramount importance to maintaining the diversity and stability of the salt
marsh (Pennings et al., 2002). A variety of ecosystem functions have been attributed to
salt marshes, some of which extend to adjacent ecosystems (Richardson, 1994). For
example, the salt marsh represents a nursery ground for many fishes and invertebrates
which may play vital roles in the food web of the salt marsh as well as adjacent
8

ecosystems (e.g. adjacent marine and freshwater habitats). In the case of the
diamondback terrapin, its survival is dependent upon a healthy salt marsh ecosystem.
Unfortunately, certain pollutants such as heavy metals have been shown to accumulate in
salt marshes (Giblin et al., 1980; Williams et al., 2003), and terrapins have been shown to
accumulate heavy metals in salt marshes that are exposed to this form of pollution
(Burger, 2002; Blanvillian et al., 2007).
The Deepwater Horizon oil spill in the northern Gulf of Mexico during 2010 and the
Chalk Point oil spill in the Patuxent River in Maryland during 2000 (Michel et al., 2001)
have highlighted the potential impact of oil pollution on estuaries inhabited by terrapin.
In the case of the Chalk Point oil spill, it was estimated to cause the mortality of
approximately 826 hatchlings and 122 adults and juveniles (Michel et al., 2001),
however, these predictions were based on a model. Holliday et al. (2008) measured
polycyclic aromatic hydrocarbon (PAH) levels in terrapin eggs a year after the Chalk
Point oil spill. The observed levels of PAH’s were low and were not strongly correlated
with shoreline oiling. A nest from a low impacted site displayed the highest PAH
concentrations. The authors attributed the oil exposure to maternal transfer reflecting
background levels instead of direct effects of the oil spill (Holliday et al., 2008).
Toxic effects of PAH’s have also been examined in freshwater turtle species. Van Meter
et al. (2006) observed snapping turtle (Chelydra serpentine) embryo mortality as well as
developmental abnormalities after exposure to oil. Bell et al. (2006) also detected
deformities in adult snapping turtles from the same population in addition to painted
turtles (Chrysemys picta). The authors credited these effects to PAH exposure because
9

high levels of PAH’s were observed in the fat reserves of the snapping turtles (Bell et al.,
2006).
The toxic effects of oil pollution have also been examined in sea turtles (Milton et
al., 2003). Sea turtles may have a higher risk oil exposure due to the wide range of
habitats that sea turtles utilize during various stages in their life history (Milton et al.,
2003). That study noted direct effects of oil pollution on the physiology and behavior of
sea turtles. The authors suggested that oil pollution could also have a variety of indirect
effects such as masking odors that may be important as foraging or orientation cues. In
addition, exposure during the smaller life history stages (i.e. hatchlings) could be more
harmful due to smaller overall size, immature metabolic physiology that is not able to
properly detoxify, and higher concentration of lipids to which contaminants can attach
(Milton, et al., 2003). The authors indicated that oiled nesting beaches could represent
another negative impact for sea turtles. Both females and hatchlings have to travel across
oil that is present on the beach and oil can also have damaging effects on the nests. It can
significantly modify gas exchange, the surrounding hydric environment, and the
incubation temperatures of the nests. Altered incubation temperatures can affect
embryonic development and sex determination (Milton et al., 2003). However, these
impacts appear to dissipate over time because sand measured the year after an oil spill did
not have measureable effects on hatchling survival or morphology (Milton et al., 2003).
Therefore, loss of salt marsh habitat, salt marsh pollution, and decrease in salt marsh
quality represent significant threats to terrapin conservation. While terrapins are
dependent upon salt marsh for their survival, they also enhance the stability and health of
the salt marsh. Terrapins are an important component in the salt marsh food web feeding
10

on a wide variety of prey (Tucker et al., 1995). Their predation on Littorina periwinkle
snails can play a pivotal role in maintaining overall salt marsh health. One of the major
vegetation types in southeastern U.S. salt marshes is Spartina alterniflora, and Silliman
et al. (2005) and Gustafson et al. (2006) both observed the deleterious effects of
unchecked grazing of Littorina snails on Spartina vegetation. Plant biomass decreased to
such an extent that all that was left were exposed mudflats (Silliman et al., 2005).
Gustafson et al. (2006) suggested that predator control of Littorina snails (such as
diamondback terrapins) is important in preserving ecosystem integrity and function of
salt marshes. The connection between diamondback terrapin abundance and salt marsh
health provides an ecosystem-wide rationale for terrapin conservation. If threats result in
the loss of a terrapin population from a salt marsh, it may represent a long-term or
permanent loss for the salt marsh. Though adult males disperse during mating season
(Hauswaldt and Glenn, 2005), overall, terrapins display high site fidelity. So if a local
aggregation is extirpated, it may take years if ever for it to become re-established
(Gibbons et al., 2001).
Road Mortality and Habitat Fragmentation
The construction of roadways in environmentally sensitive areas has led to habitat
fragmentation and road mortality in a variety of organisms (Gloyne and Clevenger, 2000;
Aresco, 2003; Szerlag and McRoberts, 2006). For terrapins, road mortality represents a
significant threat to adults during reproductive migration and can significantly decrease a
population’s reproductive potential (Wood and Herlands, 1997). In particular, adult
females are vulnerable as they migrate to suitable nesting sites through habitats that have
been fragmented by roadways. For example, on the Cape May Peninsula of New Jersey,
11

over four thousand adult females were killed between 1989 and 1995 on 11.5 km of
roadways that dissect the terrapins’ habitats (Wood and Herlands, 1997). Because natural
sand dunes that were once used for nesting had been removed during coastal
development, increased numbers of females nest along those roads’ embankments (Wood
and Herlands, 1997). Road mortality may have resulted in a change in this local terrapin
population structure with a decrease in the number of adult females in comparison to the
same population in the late 1980s (Avissar, 2006). Further, the adult females that were
captured were smaller, and because size can correlate with age, the author suggested that
the average age of adult females in the population had decreased (Avissar, 2006).
Mitigating Road Mortality
The threat of road mortality has been addressed in other turtle species. Aresco (2003)
provided a detailed account of how high amounts of traffic on U.S. Highway 27 adjacent
to Lake Jackson, FL affect the local herpetological fauna. Before fencing was inserted,
over 350 turtles were found on the short stretch of highway in a 44-day monitoring effort.
It was estimated that a turtle had a 98% chance of being struck if attempted a crossing.
As observed in the Cape May population of terrapins, a male-biased sex ratio was
detected in three species of turtles, Pseydemys floridana, Trachemys scripta, and
Sternotherus odoratus, due to the road mortality (Aresco, 2003). Steen and Gibbs (2004)
also observed male-biased sex ratios in populations of painted turtles (Chrysemys picta)
and snapping turtles (Chelydra serpentina) that inhabited high road density freshwater
wetlands. The authors noted that because turtles’ are a longed live species, populations
can endure for years before the consequences from a threat is noticeable (Steen and
Gibbs, 2004). However, insertion of fencing around the busy roadway intersecting Lake
12

Jackson led to a decrease in the amounts of dead turtles, from 9.7 individuals/km/day to
0.08 individuals/km/day (Aresco, 2003).
Road fencing has been one strategy for lessening road mortality associated with
habitat fragmentation. Fencing along U.S. Highway 90 north of Mobile Bay in AL was
constructed to significantly decrease the high levels of both female and hatchling
mortality of the federally endangered Alabama red-bellied turtle, Pseudemys alabamensis
(Nelson and Turner, 2004; David Nelson, University of South Alabama, pers. comm.).
Aresco (2003) noted the success of the inserted fences to preventing turtle mortality on
U.S. Highway 27 in FL, but the fencing was not nearly as successful in preventing road
mortality in other herpetofauna species. Less than one-half of all individuals (other than
turtles) were prevented from being struck by cars. Even though the fencing prevented
most turtles from accessing the road, many were killed due to predation behind the
fences. Only one culvert, allowing animals to access the other side of the highway, exists
in the area, so chances for predation for turtles as well as other herpetofauna increases as
they travel farther distances to find the culvert (Aresco, 2003).
The findings by Aresco (2003) suggested that some form of tunnel structure might be
an effective strategy for facilitating movement of turtles and other herpetofauna under
roadways separating habitats. In fact, Jackson and Griffin (2000) argued that fencing
without any crossing structure will only increase the fragmentation of habitats. They
listed the various factors that should be considered when constructing movement tunnels.
They suggest that the placement of such structures could be the most important factor
because, if they are too widely spaced, the tunnels may not properly serve their purpose.
Other factors include size, light, moisture, temperature, noise, substrate, approaches, and
13

species interactions, although different species will have different requirements for these
factors. The authors suggested that more cost efficient strategy would include a mixture
of widely spaced large structures and more frequently spaced small structures (Jackson
and Griffin, 2000). However, Simberloff et al. (1992) contended that despite possible
rationales for the construction of migration corridors (to maintain overall number of
species, to decrease demographic stochasticity, to allow for gene flow to prevent
inbreeding depression, and to provide for inherent need of certain species to move) data
is needed to evaluate the cost effectiveness of this strategy. Until such data exist, other
management options must not be blindly discarded (Simberloff et al., 1992).
The development or preservation of wildlife migration corridors has also been
suggested as a method of alleviating problems associated with habitat fragmentation and
road mortality. Simberloff et al. (1992) discussed the various rationales for the
construction of migration corridors: to maintain overall number of species, to decrease
demographic stochasticity, to allow for gene flow to prevent inbreeding depression, and
to provide for inherent need of certain species to move. However, they contend that data
is needed to evaluate the cost effectiveness of migration corridors, and until such data
exits, other management options must not be blindly discarded (Simberloff et al., 1992).
Crab Trap-Induced Mortality
Another major threat to terrapin populations is crab trap-induced mortality (reviewed
below). The crab fishery along the Atlantic and Gulf coasts of the U.S. primarily targets
the blue crab (Callinectes sapidus) (Kennedy et al. 2007). The perishability and
marketing of the blue crab initially limited the proliferation of this fishery, but by the late
1800’s and early 1900’s commercial crab fisheries were established in various locations
14

along the Atlantic and Gulf coasts of the U.S (Stagg and Whilden, 1997; Guillory et al.,
2001b; Kennedy et al., 2007). The modern crab trap was introduced in 1927 in
Chesapeake Bay (Kennedy et al., 2007). The design was improved in 1938 and became
the “gear of choice” for crab capture by the 1950’s as the crab fisheries proliferated
throughout the Atlantic and Gulf coast regions (Stagg and Whilden, 1997; Guillory et al.,
2001b; Kennedy et al., 2007). Yearly landings have fluctuated, but the total amount of
blue crab landed along the Atlantic and Gulf coast of the U.S. gradually increased from
the 1950’s through 1970’s (Stagg and Whilden, 1997; Guillory et al., 2001b; Kennedy et
al., 2007). In general, this was concomitant with an increase in the number of crab traps
and crab fisherman (Hill et al., 1989; Guillory et al., 2001b). Since the 1970’s, fishing
effort has remained relatively high in the blue crab fisheries, although significant declines
in landings have occurred in recent years in many areas due to factors such as
overharvesting, pollution, and loss of habitat (Murphy et al., 2007; Sutton and Wagner,
2007; CBSAC Report, 2010).
The expansion of blue crab fishery along the Atlantic and Gulf coasts of the U.S.
from the middle of the 20
th
century to present day has significantly impacted terrapin
populations. The crab trap is the preferred capture method in most areas and a numerous
studies over the past several decades have documented their impact on terrapin
populations. An initial study by Bishop (1983) examined the incidence of terrapin by-
catch in crab traps in South Carolina estuaries. Crab traps were observed to capture high
numbers of terrapins, and the catch was male biased (2.3:1). Based on his sampling
results and the size of the crab fishery in South Carolina, he estimated that 2,853 terrapin
would be captured daily in that fishery from May through April (Bishop, 1983).
15

However, he indicated that the mortality rate was lower than the capture rate depending
on factors such as the frequency at which traps were checked and water temperature
(ranging to as low as 10% of the capture rate) (Bishop, 1983).
Subsequent to the study by Bishop (1983), the widespread impact of crab traps on
terrapin populations has been documented in numerous studies. Crab trap-induced
mortality of terrapins has been documented in New Jersey (Wood, 1997), Maryland
(Roosenburg, 1991; Roosenburg et al., 1997; Roosenburg and Green, 2000), South
Carolina (Dorcas et al., 2007), Georgia (Grosse et al., 2009), Florida (Siegel, 1993), and
Alabama (Coleman et al., unpublished data). The study in Maryland estimated that more
than 2000 terrapins were caught annually by crab traps (Roosenburg et al., 1997).
Roosenburg et al. (1997) indicate that if all those terrapins died, this would represent 15
to 78% of local population being removed annually. A male-biased capture ratio (3:2),
was observed in that study, and it probably contributed to the female biased sex ratio
(2:1) that was observed. Terrapins exhibit sexually dimorphic body size with females
growing significantly larger than males, and Roosenburg et al. (1997) indicate that once
females reach a certain size, they are too large to enter the crab trap funnels. However,
males never reach that size, so they are vulnerable to crab trap mortality throughout their
lives (Roosenburg et al., 1997). Both commercial and recreational crab trapping
disproportionally remove juveniles and adult males creating populations with female
biased sex ratios. However, the elimination of juvenile females might have the greatest
impact because these individuals have not yet contributed reproductively (Roosenburg et
al., 1997). In fact, Tucker et al. (2001) calculated the mean life span for females in the
16

sampled population in South Carolina and found that the average female did not survive
to reproductive maturity.
A long-term study by Dorcas et al. (2007) indicated that crab trap-induced mortality
was a major factor causing declines and demographic changes in the terrapin population
inhabiting the salt marshes of Kiawah Island, GA. That study was based on more than
twenty years of mark-recapture data and found a significant decrease in population size
based on long-term sampling from multiple locations and found a demographic shift to a
female-biased population comprised of larger and older individuals.
Roosenburg et al., (1997) suggested that the recreational and commercial crab
fisheries could have varying impacts on terrapin populations. Recreational crab traps are
often in more near-shore salt marsh habitats (Roosenburg et al., 1997) where the terrapins
(juveniles and adult males) that are more susceptible to crab trap mortality are more apt
to be found (Roosenburg et al., 1999). In South Carolina, recreational crab traps could
possibly outnumber commercial crab traps, suggesting a relatively high impact of the
recreational crab fishery on terrapin mortality (Hoyle and Gibbons, 2000).
Data from these studies also suggested that abandoned or lost crab traps (often
referred to as derelict or “ghost” crab traps) may pose a greater risk of mortality for
terrapins than those that are checked regularly. Bishop (1983) initially indicated that
derelict crab traps could pose an increased risk due to numerous records suggesting that
terrapins captured in crap traps tend to attract other terrapins. There have been several
published anecdotes of derelict traps containing large number of carcasses: 29 terrapins in
one trap in South Carolina (Bishop, 1983), 49 in one trap in Maryland (Roosenburg,
1991), and 94 in one trap in Georgia (Grosse et al., 2009). This problem is amplified by
17

the relatively large number of derelict crab traps that are added yearly to estuaries along
the Atlantic and Gulf coasts of the U.S. As examples, it has been estimated that 25% to
30% of crab traps are lost or abandoned yearly resulting in estimates of 250,000 traps per
year added to the Gulf of Mexico (Guillory and Perret, 1998; Guillory et al., 2001a), and
100,000 traps per year added to the Chesapeake Bay in the State of Virginia (Havens et
al., 2008).
Mitigating Crab Trap-Induced Mortality
Two mitigation measures have been developed to modify crab traps in order to
decrease terrapin mortality; (1) modified crab traps (Roosenburg et al., 1997) and (2) by-
catch reduction devices or BRD’s (Wood, 1997) Modified crab traps are taller than
standard crab traps and allow terrapins to enter the upper tier of the trap, to surface and
breath. In a study examining the efficacy of modified crab traps, no significant
differences in crab captures were observed between modified and standard traps, and the
modified crab traps did not display the terrapin mortality. However, due to their
increased costs in building and awkwardness in handling, modified crab traps are better
suited as recreational crab traps (Roosenburg et al., 1997).
By-catch reduction devices are metal wire or plastic rectangles initially developed by
Wood (1997), which fit into the crab trap funnel openings and are intended to prevent
terrapin entry into the traps while still allowing crabs to enter. Wood (1997) found that
traps outfitted with 5 x 10 cm BRDs performed the best in preventing terrapin capture
and permitting crab capture. In fact, capture of marketable sized crabs was significantly
enhanced in these traps. This scenario was also observed by Guillory and Prejean (1998)
and Roosenburg and Green (2000). Further, no significant differences in crab captures
18

between traps fitted with BRD’s and non-fitted traps were detected in other studies (Cole
and Helser, 2001; Butler and Heinrich, 2007; Morris et al., 2010; Rook et al., 2010).
Guillory and Prejean (1998) suggested that the presence of BRD’s may prohibit the
egress of crabs by preventing manipulation of the funnel openings by the crabs.
Both Roosenburg and Green (2000) and Dorcas et al. (2007) contended that the size of
BRD’s should not be uniform across the terrapin’s range but instead be adjusted based on
regional size differences. While Wood (1997) found that 5 x 10 cm BRD performed best
in New Jersey, the 4.5 x 12 cm performed best in the Chesapeake Bay (Roosenburg and
Green, 2000) and Delaware Bay (Cole and Helser, 2001). The orientation of the BRD in
funnel opening could also affect its performance. Crabs would presumably still enter
traps fitted with BRD’s in the vertical position, but terrapins could not (Hoyle and
Gibbons, 2000; Dorcas et al., 2007).
Programs for mitigating the impact of derelict crab trap-induced mortality have been
initiated in a variety of states. Trap removal programs have been initiated in both Gulf
and Atlantic coast states in the past (Guillory et al., 2001a), and now many have yearly
derelict trap removal programs (Perry et al., 2008).
Finally, restricting the use of crab traps in specific terrapin habitats and/or during
specific times of the year could represent a powerful mitigation tool for initiating and
facilitating the recovery of populations which are in danger of extirpation (Butler, 2002).
Nest and Hatchling Predation
The predation of nests and hatchlings also represents a major threat to terrapin
populations. While some predation can be considered a natural threat which has long
been an integral part of terrapin ecology, anthropogenic factors such as coastal
19

development can increase the natural predator load and the access of predators to terrapin
nesting areas. Further, the decline of terrapin populations associated with other threats
such as habitat loss and crab trap-induced mortality could leave a population in a
precarious state in which nest depredation could have a much greater effect on the
survival status of a population.
High levels of egg and hatchling predation have been observed on numerous terrapin
nesting beaches. Butler et al., (2004) reviewed a variety of terrapin studies that reported
nest predation rates from 41 to 88 %. Feinburg and Burke (2003) observed over 3,000
depredated nests over two nesting periods and attributed over 98% of the depredations
were due to raccoons. In Maryland, 94% of nests were taken from a sampled nesting
location with raccoons as the main predator (Roosenburg and Place, 1995). Butler et al.
(2004) detected over 80% of nests were depredated over two nesting seasons in a Florida
terrapin population, and raccoons were the primary nest predator. A study of a terrapin
population in Rode Island found that 87% of the monitored nests were depredated and
raccoons were the primary predator (Goodwin, 1994). Raccoons appear to use cues such
as scent and habitat disturbance when locating nests. In a study by Burke et al. (2005) at
a nesting habitat at Jamaica Bay Wildlife Refuge, cues that seemed to be most important
for raccoons were disturbance of nesting soils and as well as disturbance of the natural
beach scents. Further, the majority of nests were preyed upon within 24 hours of
oviposition (Burke et al., 2005).
In addition to raccoons, other reported predators include foxes (Burger, 1977;
Roosenburg and Place, 1995), otters (Roosenburg and Place, 1995), gulls (Burger, 1977),
crows, boat-tailed grackles, armadillos, and ghost crabs (Butler et al., 2004). Beach
20

vegetation can also contribute to egg and hatchling mortality. The roots of various beach
grasses have also been observed to invade terrapin nests (Lazell and Auger, 1981; Butler
et al., 2004). Beach grasses have been planted for erosion control as alternative to other
measures such as bulkheading (Roosenburg, 1991). The roots of beach grasses can
impact eggs and hatchlings, including the direct penetration and mortality of eggs. Nests
can also become root bound, which can result in embryonic death, prevent successful
pipping and hatching, and/or prevent the emergence of hatchlings (Lazell and Auger,
1981; Butler et al., 2004).
A variety of human activities can increase the population density of predators for
terrapin eggs and hatchlings. For example, nuisance raccoons were illegally released on
an island in the Jamaica Bay Wildlife Refuge whose terrapin nesting beach subsequently
experienced high levels of nest predation (Feinburg and Burke, 2003). Human-induced
reduction of natural predators of raccoons and foxes (e.g. wolves, coyotes, etc.) combined
with the reduction of the fur market has also contributed to increases in terrapin nest
depredation (Congdon, et al., 1993; Roosenburg and Place, 1994). Additionally,
raccoons can thrive in suburban environments, which are increasing in many coastal
areas (Roosenburg and Place, 1994). Finally, the increase in roads and bridges associated
with coastal development can increase predator access to salt marshes inhabited by
terrapin.
Mitigating Nest and Hatchling Predation
Different methods of raccoon control have been attempted in order to decrease nest
predation in terrapin populations as well as in other turtle species. Munscher (2007)
found that removal of raccoons significantly decreased nest predation in a terrapin
21

population in northeastern Florida. However, after the control efforts were ended,
raccoons quickly repopulated the area (Munscher, 2007). A study by Christiansen and
Galloway (1984) documented a significant decrease in both nest and hatchling predation
of freshwater turtles at Big Sand Mound, Iowa, when raccoons were removed from the
area. A study by Ratnaswamy et al. (1997) found that lethal removal of raccoons did not
significantly affect sea turtle nest predation, even though the authors noted that this
method may need to occur over a longer time period than it did in this study to have a
significant effect. At Ten Thousand Islands, FL, raccoons were preying on 76-100% of
sea turtle nests on the four sampled islands over four years. Sixteen raccoons were
removed over two years on one of these islands (Panther Key), and nest depredation fell
to 0% (Garmestani and Percival, 2005). However, it was suggested that the repopulation
of raccoons in that area might occur more quickly if a given site were better connected to
surrounding habitats, thus overcoming the initial benefit of predator removal (Garmestani
and Percival, 2005).
Ratnaswamy et al. (1997) compared the efficacy of three different control methods
for raccoons on a sea turtle nesting beach: lethal removal, non-lethal conditioned taste
aversion, and screening of nests. The most labor intensive and costly method was nest
screening, but it was also the most effective at preventing raccoon predationIn the case of
a terrapin population. Nest screening has also been used on terrapin nesting beaches as a
method for decreasing predation (Goodwin, 1994; Giambanco, 2002; Butler et al., 2006).
However, the strategy requires close monitoring of nesting beaches in order to identify
and screen nests soon after they are laid.
22

Ratnaswamy and Warren (1998) contended that it is pivotal to truly understand the
complete role a predator has in an ecosystem before it is removed. Raccoons may
provide benefits to the ecology of a specific habitat. For example, raccoons could serve
as seed dispersers as well as important predators on invertebrate and other vertebrate
species in coastal systems, and altering these relationships could have unforeseen
detrimental effects (Ratnaswamy and Warren, 1998). Barton (2005) observed that ghost
crab abundance increased when raccoons were removed, and this shift resulted in overall
increases in sea turtle egg mortality due to ghost crab predation. However, if predator
removal is undertaken, efforts have to be consistent at least in the short term. Engeman et
al. (2006) detected an almost immediate increase to high levels of sea turtle nest
predation in the middle of the 2004 nesting season after several years because predator
removal efforts were abandoned due to budgetary constraints.
As an alternative to predator removal, efforts have been undertaken to exclude
predators from nesting beaches. Bennett et al. (2009) constructed electric fences on
terrapin nesting beaches to keep out raccoons and foxes. Although no significant
differences in predator exclusion were observed between control and treatment plots,
predation was less in the treatment plots, so absence of statistical significance was
attributed to low sample size. The authors noted that this method could have
complications based on the conductivity of the nesting beach substrate, amount of
vegetation on nesting beach, and relative size of nesting beach (Bennett et al., 2009).
Head-Starting as a Strategy for Decreasing Egg and Hatchling Mortality
Another potential alternative for decreasing predation in terrapin populations is
through head-starting programs. This method involves the artificial incubation of eggs
23

followed by the rearing of hatchlings to a certain size before releasing them into the
environment. This circumvents the high mortality associated with early life history
stages of terrapins (i.e. egg incubation, hatchling, and post hatchling stages). Head-start
projects have been incorporated into ongoing terrapin conservation programs in New
Jersey (Herlands et al., 2004), Maryland (Smeenk, 2010) and Alabama (Coleman et al.,
2010). A terrapin head-starting project was initiated in southern New Jersey in 1989 in
response to high levels of road mortality. Annually, 200-300 hatchlings are released back
into the salt marshes, although 400-600 adult females are killed each year due to road
mortality (Herlands et al., 2004), thus the possible benefits of head-starting may not be
able to compensate for the high level of road mortality. Smeenk (2010) studied the
effectiveness of head-starting as a conservation strategy in a terrapin population in
Maryland. That project released 664 terrapins that were 2 -5 years old. The results
suggest that potential benefits from head-starting project were not sufficient to
compensate for the high mortality from crab traps in that area (Smeenk, 2010).
A factor that should be considered when head-starting terrapins is the threat of
introducing pathogens into the wild populations (Cunningham, 1996). Werner et al.
(2002) screened both captive-reared diamondback hatchlings as well as wild hatchlings
for pathogens and found both groups relatively healthy. It was concluded that
introducing head-started individuals did not pose any health dangers to the wild terrapins
(Werner et al., 2002).
The effectiveness of head-starting has also received attention in the Kemp’s ridley
sea turtle (Dodd and Siegel, 1991; Frazer, 1992; Heppell et al., 1996). Heppell et al.
(1996) created a model that examined the effects of head-starting on a population of
24

Kemp’s ridley sea turtle. That model indicted that the beneficial effects of head-starting
would be negated if there were high levels of threat-induced mortality of the sub-adult
and adult portions of the population. This conclusion aligned with those of Frazer
(1992). He described head-starting as a “half-way technology”, because it did nothing to
address the threats faced by juvenile and adult portions of the population. However, data
from that project indicate that head start Kemp’s ridleys can survive and reproduce in the
wild (Shaver and Caillouet, Jr., 1998; Shaver and Wibbels, 2007). Head-started female
western pond turtles (Actinemys marmorata) have also been documented nesting (Vander
Haegen et al., 2009). To properly assess the effectiveness of any head-start program,
long-term monitoring on both the nesting beaches and foraging habitats will be required
(Shaver and Wibbels, 2007), an idea also argued by Dodd and Siegel (1991).
Protection and Legal Take of Terrapins
Data indicate that terrapin populations throughout their range are being impacted by
the various threats discussed above (Butler et al., 2006). However, the survival status
varies between populations that have been studied, and the survival status of many
populations has not been well documented (Butler et al., 2006). Because of the
variability in the survival status and the paucity of data from most locations, the laws
protecting terrapins and regulating their legal take vary greatly. The U.S. Fish and
Wildlife Service (USFWS) has listed the diamondback terrapin as a “status review
species” for several decades (Hart and Lee, 2006). However, range-wide population
surveys data are lacking, so the USFWS has not been able to determine if federal
protection is necessary. Therefore, the current protection and regulation of legal harvest
of terrapins is under state control. The state laws regulating the protection and legal take
25

of terrapins vary widely (Reviewed in Figure 1). Of the 16 states along the Atlantic and
Gulf coasts of the U.S. with terrapin populations, seven have regulations that prevent the
legal take of terrapins (AL, CT, GA, MA, RI, TX and VA). Six states (DE, FL, MD, MS,
NC and SC) have laws that limit daily or yearly take of terrapin (Figure 1), and three
states (LA, NJ and NY have open harvest during a specific season (Figure 1). The laws
regulating the protection and harvesting of terrapin can obviously impact the survival
status and recovery of a given population. However, to fully understand that impact, the
overall magnitude of the legal take must be taken into account relative to the size and
stability of the population. It seems intuitive that decreasing legal harvest would enhance
the stability and recovery in a population, but this needs to be evaluated on a population
by population basis.
Summary
Diamondback terrapins endured historical exploitation and their current survival is
being impacted by a spectrum of new anthropogenic threats. The purpose of the current
review was to highlight and evaluate the primary threats to terrapin populations and
potential strategies for their recovery. Surveys of terrapin researchers from along the
Atlantic and Gulf coasts indi