B I O L O G I C A L C O N S E RVAT I O N
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Long-term trends in the demography of the Allen Cays
Rock Iguana (Cyclura cychlura inornata): Human
disturbance and density-dependent effects
John B. Iversona,*, Sarah J. Converseb,1, Geoffrey R. Smithc, Jennifer M. Valiulisd
a
Department of Biology, Earlham College, Richmond, IN 47374-4095, USA
Department of Fishery and Wildlife Biology, Colorado State University, Fort Collins, CO 80523-1474, USA
c
Department of Biology, Denison University, Granville, OH 43023, USA
d
Department of Biology, Colorado State University, Fort Collins, CO 80523-1474, USA
b
A R T I C L E I N F O
A B S T R A C T
Article history:
Allen Cays Rock Iguanas (Cyclura cychlura inornata) are native to two small islets (Leaf and U
Received 23 December 2005
Cay) in the north-central Bahamas. These populations were nearly extirpated in the early
Received in revised form
1900s because of heavy hunting pressure (for food), but increased to a total of ca. 150 lizards
16 April 2006
in 1970, and now number over 500 (not including juveniles). Over the past several decades
Accepted 18 April 2006
poaching has declined, but tourist visitation (including nearly daily supplemental feeding
Available online 21 June 2006
of iguanas) has increased. To examine human impacts on the demography of these iguanas, survival, population growth rates, and population sizes for subadult and adult
Keywords:
(>25 cm snout-vent length) males and females on the two cays were estimated based on
Population growth
mark–recapture data collected over a 25-year period (1980–2004). As predicted, annual sur-
Survival
vival probability was higher on U Cay (with less human visitation) than on Leaf Cay, was
Tourist feeding
higher in females than in males (which are bolder), and exhibited a declining trend. Both
Carrying capacity
populations more than doubled during this study, but population growth rates declined
Relocation
to near zero in recent years. These data reflect the importance of human impacts, but also
Lizard
suggest that the populations may be nearing carrying capacity. The rapid population
growth observed on these cays, and that seen for several other translocated iguana populations, suggest that if unnatural causes of mortality are reduced or eliminated, island populations of iguanas are capable of rapid recovery. The inexpensive establishment of
assurance colonies on undisturbed ‘‘islands’’ should be considered for any comprehensive
management plan for endangered species of iguanas.
2006 Elsevier Ltd. All rights reserved.
1.
Introduction
As remote locations become more accessible, and as rates of
human visitation to remote locales increase, populations of
threatened or endangered animals and plants (as well as
some healthy populations) face new threats and challenges.
For example, increased interactions with humans, often as a
result of increased tourism, have been shown to affect a range
* Corresponding author: Tel.: +1 765 983 1405; fax: +1 765 983 1497.
E-mail addresses: [email protected] (J.B. Iverson), [email protected] (S.J. Converse), [email protected] (G.R. Smith),
[email protected] (J.M. Valiulis).
1
Present address: Colorado Cooperative Fish and Wildlife Research Unit, Patuxent Wildlife Research Center, 12100 Beech Forest Road,
Laurel, MD 20708, USA.
0006-3207/$ - see front matter 2006 Elsevier Ltd. All rights reserved.
doi:10.1016/j.biocon.2006.04.022
B I O L O G I C A L C O N S E RVAT I O N
of vertebrates, including birds (Steidl and Anthony, 2000;
McClung et al., 2004; Müllner et al., 2004; Blumstein et al.,
2005), mammals (Boydston et al., 2003; Manor and Saltz,
2003), amphibians (Rodrı́guez-Prieto and Fernández-Juricic,
2005), and reptiles (Garber and Burger, 1995; Labra and Leonard, 1999; Romero and Wikelski, 2002; Kerr et al., 2004). Such
disturbance can have negative effects on the long-term viability of populations, either through decreased reproductive success, increased mortality, or reduced access to resources
(Garber and Burger, 1995; Steidl and Anthony, 2000; Yorio
et al., 2001; McClung et al., 2004; Rodrı́guez-Prieto and Fernández-Juricic, 2005).
Long-lived lizards, especially those that live on islands, are
especially vulnerable to human disturbance (Alberts, 2000;
Romero and Wikelski, 2002; Lacy and Martins, 2003; Knapp,
2004; Pregill and Steadman, 2004). However, evaluating the
real and potential impacts of increasing human disturbance
requires good basic life history data, as well as information
about changes to demographic parameters in populations exposed to humans. Unfortunately, although considerable research has been done on the life history of short-lived
lizards (Dunham and Miles, 1985; Shine and Charnov, 1992;
Niewiarowski et al., 2004), much less work has focused on
long-lived lizards (i.e., those that mature at >5 years; e.g., reviews in Huey et al. (1983) and Vitt and Pianka (1994)). This is
particularly troublesome because long-lived lizards are disproportionately under threat of extinction, especially those
on islands (Alberts, 2000; IUCN, 2000), at least in part due to
their particular life histories (delayed maturity, low reproductive rates, sensitivity to adult survival). But because of their
long lives, obtaining multi-generational data to quantify life
history traits and to permit population modeling with the goal
of developing sound management strategies is often timeconsuming and expensive. This is further complicated because so many populations of long-lived lizards are in decline
and/or are too rare to make field work feasible even over the
short term (Alberts, 2000), and thus not likely to provide accurate measures of demographic parameters or their normal
variation under natural (i.e., undisturbed) conditions. This
places a high premium on the immediate study of the least
disturbed long-lived species in order to provide baseline
demographic data for managing those species in the future,
for identifying factors that may contribute to population increases and declines, and for extrapolation to similar but
imperiled species with the goal of reversing their declines.
To date, long-term studies of long-lived lizards under conditions of minimal human disturbance have been confined to
the Galapagos Islands (Snell and Tracy, 1985; Laurie and
Brown, 1990a,b; Cayot et al., 1994; Wikelski and Thom, 2000;
Wikelski and Romero, 2003), the Turks and Caicos Islands
(only four years; Iverson, 1979), and the Bahamas (Iverson
et al., 2004a,b). The latter site has been studied since 1980
(only a few months shorter than the longest study in the Galapagos; Snell and Tracy, 1985), and has provided the most detailed life history data available for a long-lived lizard
species (the Allen Cays Rock Iguana, Cyclura cychlura inornata)
outside the Galapagos, including data on growth (Iverson
et al., 2004a), and reproductive output and nesting ecology
(Iverson et al., 2004b). In addition, Iverson et al. (2004b) reported estimates of survival in the nest, but only preliminary
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301
estimates of adult survival have previously been generated
(Iverson, 2000).
The Allen Cays Rock iguana, the most northerly member
of its genus (24.75 N latitude), is known to occur naturally
on only two cays in the Allen Cays of the northern Exuma Islands in the north-central Bahamas: Leaf Cay (4 ha; 4 m maximum elevation) and U Cay (3 ha; 9.5 m). In 1892 C.J. Maynard
found Allen Cay iguanas to be ‘‘not uncommon’’ on U Cay, but
when he was storm-bound off the island in early March of
1915, he could find only two (Barbour and Noble, 1916). Both
of those were shot; one became the holotype and the other
escaped. Hence, Barbour and Noble (1916) believed that the
Allen Cays iguana was ‘‘beyond doubt extinct’’ (see also Barbour, 1937) because of human removal for food. However,
although Maynard apparently had ‘‘ample opportunity to
cover [U Cay] completely’’ (Barbour and Noble, 1916), the
weather was inclement during his visit, and most of the population likely remained out of sight, within retreats. Indeed,
the density of the vegetation on some parts of the islands,
the small size of juveniles, and the abundance of inaccessible
cavities in the honeycomb limestone into which the iguanas
can retreat make it doubtful that humans could directly remove every iguana from either Leaf or U Cay. Nevertheless,
early editions of the ‘‘Yachtsman’s Guide to the Bahamas’’,
published annually since 1956, also reported that iguanas
were extinct in the Allen Cays (Carey, 1976).
When Wayne King (personal communication) visited the
Allen Cays in August of 1964, he reported that iguanas ‘‘while
still extant, [were] not thriving’’. Six years later, in late March
of 1970, Carey (1976) used two (undefined) survey techniques
to estimate the population of iguanas on Leaf Cay at 65 and
73. He did not systematically survey U Cay, but speculated
that the population there was ‘‘at least equivalent to that of
Leaf Cay’’. Field observations by the authors suggested that
the populations have increased substantially since then, perhaps indirectly due to the passage of the Wild Animals Protection Act by the Bahamas government in 1968 (Sandra
Buckner, personal communication). That law provided legal
protection for the Allen Cays Rock Iguana for the first time,
and hunting pressure has apparently declined since then.
However, this iguana is still considered Endangered by the
IUCN because of its restricted range, its small population size,
and its vulnerability to hurricanes (Alberts, 2000). In addition,
some poaching still occurs (for consumption and for the illegal pet trade) and tourist visitation has increased since 1970,
with nearly daily supplemental feeding today (Fig. 1). Up to
150 visitors/day now visit Leaf Cay (most on powerboat tours
that arrive at 1000 h and 1600 h, feed fruits to the iguanas, and
leave after 30–45 min), and only up to 25/day visit U Cay (inaccessible to powerboats). Feeding by tourists seems not to have
affected average growth rates of individual iguanas over the
past two decades (Iverson et al., 2004a). However, while these
lizards are almost completely herbivorous by nature (e.g., Alberts, 2000), they readily accept and consume almost any item
thrown to them by tourists (from brownies to bread to ground
beef to styrofoam; personal observations). The direct impact
of this activity on individual lizards is not known, but it is suspected that it might affect their survival.
Because the two inhabited islands have similar areas, have
no invasive vertebrate species, supported iguana populations
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Fig. 1 – Tourist-acclimated Allen Cay Iguanas on the ‘‘feeding beach’’ at Leaf Cay. As many as 50% of the captures of iguanas
during a sample period are made on this beach, which represents only 2% of the island area.
of similar size in 1970 (Carey, 1976), and the average growth
rate of individual iguanas between 1980 and 2000 was similar
(Iverson et al., 2004a), these islands are an ideal system in
which to examine the effects of tourist visitation, as well as
the reduction in poaching, on iguanas. This paper analyzes
25 years of mark–recapture data from Allen Cays rock iguanas
to estimate survival rates, as well as population sizes and
population growth rates.
The authors predicted that because males appear to be
more bold than females (JBI, GRS, JMV; personal observations),
they should be more strongly impacted by humans (via misfeeding, poaching, or simple mischief). Hence, survival in
males should be lower than in females. It was also predicted
that because Leaf Cay is visited by humans significantly more
than U Cay, survival rates should be higher on U Cay than on
Leaf Cay. Since tourist visitation has increased over the last 25
years, the data were also examined for evidence that these effects might have increased over that time. Finally, because of
the reduction in poaching during the last 25 years, and the
apparent increases in population sizes during that time, the
possibility that these populations are nearing carrying capacity is discussed.
2.
Materials and methods
2.1.
Field methods
Populations of Allen Cays rock iguanas (C. c. inornata) were
sampled on Leaf Cay and U Cay in the Allen Cays, northern
Exuma Islands, Bahamas in 17 of 25 years from 1980 to
2004. A detailed study area description is provided by Iverson
et al. (2004b). These are believed to be the only natural populations of this subspecies, because their presence there has
been known from at least 1892 (Barbour and Noble, 1916).
Introduced populations are known from at least four other
cays in the northern Exumas (Knapp, 2001; Iverson, unpublished). Sampling effort on Leaf and U Cays was variable, with
a general increase over time. In 1980, one researcher (JBI)
spent only a few hours in mid-March on each island. From
1982 through 2000, at least 6–12 researchers spent five days
in mid-March capturing and marking as many iguanas as
possible. During mid-May in 2001 through 2004, at least 13
researchers spent 5–6 days capturing and processing iguanas.
From mid-June to mid-July in 2001 and 2002, three researchers also captured and marked iguanas while studying the
nesting ecology of females (Iverson et al., 2004b).
Captured iguanas were uniquely marked by toe clips (three
toes maximum) following Ferner (1979), since toes are sometimes lost naturally in this genus (Carter and Hayes, 2004;
Iverson, unpublished), and toe-clipping has been shown to
have little effect on cursorial lizards (Dodd, 1993; BorgesLandáez and Shine, 2003; Langkilde and Shine, 2006). Between
1993 and 2004, 572 iguanas were also injected with PIT tags
(Info-PetTM; passive integrated transponder). Iguanas were
sexed by cloacal probing (Dellinger and von Hegel, 1990) and
released as soon after capture as possible (usually within a
few hours; always within 24 h). Full details of capture and
processing methods were presented by Iverson et al. (2004b).
Juvenile Allen Cays iguanas (particularly young of the year)
are difficult to capture as reliably as subadults and adults.
However, by about 6.5 years of age, and about 20 cm snoutvent length (SVL), they are easily detected and can then
usually be captured when located (personal observations). Because of this, analyses were limited to captures of subadult
and adult iguanas >25 cm SVL (nearly all >8.5 years of age
Iverson et al., 2004a).
2.2.
Demographic analyses
A three-step analytic process was used for the demographic
analyses, which consisted of: (1) modeling capture probability (p), (2) modeling survival probability (/), and (3) modeling
population growth rate (k). The Cormack–Jolly–Seber model
(Cormack, 1964; Jolly, 1965; Seber, 1965), generalized by Lebr-
B I O L O G I C A L C O N S E RVAT I O N
eton et al. (1992), was used to estimate, from a joint likelihood function, capture probability (p) and apparent survival
(/). Apparent survival is the probability of surviving and
remaining on the study area, which is here equivalent to
true survival because natural, successful emigration from
the islands is essentially impossible. Pradel’s (1996) reversetime model was used to estimate population growth rate
(k). The Pradel model is an extension of the Cormack–Jolly–
Seber model, where k may also be estimated in a joint
likelihood function with capture (p) and apparent survival
probabilities (/). Analyses were conducted in Program MARK
(Version 3.2; White and Burnham, 1999). Although sampling
occurred on a less-than-annual basis, annual survival rates
and population growth rates were computed by scaling rates
computed over longer periods by taking the appropriate root
of the rate.
Prior to the analyses, goodness-of-fit (GOF) of the Cormack–Jolly–Seber model was examined using Program RELEASE (Test 2 + Test 3; Burnham et al., 1987), which resulted
in v2 = 377.4 (df = 184), indicating that overdispersion existed
in the data, most likely due to heterogeneity in capture and/
or survival probabilities among individuals. Therefore, an
information-theoretic model selection and multi-model inference approach was used based on AIC (Akaike’s Information
Criterion; Akaike (1973)) corrected for both small sample size
and overdispersion (QAICc; Hurvich and Tsai, 1989; Lebreton
et al., 1992). To account for overdispersion further, variance
estimation was adjusted by ^c ¼ 2:05 (Lebreton et al., 1992;
Burnham and Anderson, 2002).
After goodness-of-fit testing, the top-ranked model structure (i.e., the model with the lowest QAICc value and hence,
the most supported structure) on capture probabilities was
determined by analyzing 10 possible capture probability
structures with the fully parameterized / model, /(sex ·
cay · t). This notation follows Lebreton et al. (1992), and indicates full time dependence in survival probabilities, with
interactions between sex, cay, and year. The 10 capture probability structures were p(sex · cay · t), p(sex · cay + t), p(sex ·
t + cay), p(sex + cay · t), p(sex + cay + t), p(sex · t), p(sex + t),
p(cay · t), p(cay + t), and p(t), where the symbol ‘‘+’’ indicates
main effects only (i.e., no interactions). Thus, for example,
the model p(sex · cay + t) includes the main effects of sex,
cay and time, and the interactions between sex and cay. Only
time-dependent structures on capture probability were considered, so as to retain the existing time structure in the data
and thus avoid biasing the estimation of time trends in the
data (described below).
The mark–recapture data were next submitted to the Cormack–Jolly–Seber model survival analysis, where survival was
modeled using the AICc-based top-ranked parameterization
of capture probabilities, and parameterizations of / that included three variables: sex, cay, and time trend (T). All combinations of these effects with and without interactions were
considered, for a total of 15 models.
Based on this suite of models survival rates and capture
probabilities were computed using multi-model inference
procedures described by Burnham and Anderson (2002); i.e.,
model-averaged estimates of survival and capture probabilities were calculated by averaging estimates for all of the models, weighted by Akaike weights (i.e., based on relative AICc
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303
values), and unconditional standard errors were calculated
as described by Burnham and Anderson (2002).
To estimate population growth rate, the top-ranked model
structure from the survival analysis on p and / was used, and
then the same set of structures on k within the Pradel model
was considered for / in the Cormack–Jolly–Seber model analysis; that is, all 15 combinations of sex, cay, and T, with and
without interactions. Estimates of population growth rate
were computed by model-averaging estimates from this set
of models (Burnham and Anderson, 2002), as described above.
Finally, estimates of abundance for each group in the analysis (i.e., two sexes and two cays) and associated standard errors were calculated based on the Williams et al. (2001)
estimator of abundance, where abundance in year i is calculated as
^ i ¼ ni
N
^i
p
with approximate standard error, calculated using a deltamethod approximation (Seber, 2002), equal to
^i Þ 1=2
n
ni ð1 p
^ i Þ varðp
^i Þ 2i þ
SEðN
^i
^i
p
p
where ni is the number of individuals captured in year i, and ^pi
is the estimated capture probability in year i, computed based
on model-averaged capture probability estimates from the
analysis of survival rates.
3.
Results
3.1.
Capture probabilities
Through 2004, 1074 individual iguanas had been marked on
Leaf Cay (with 2162 additional recaptures; 52.9% of those of
determinable sex being male) and 482 on U Cay (plus 1325
recaptures; 57.1% male). The analysis of capture probabilities
indicated that the p(sex · cay + t) structure was the best model
of capture probability, with 75% of the Akaike weight (wi; Burnham and Anderson, 2002). This structure was used in the subsequent Cormack–Jolly–Seber model of survival probability.
Capture probabilities from the survival analysis were variable
over years, cays, and sexes, possibly due to variation in the
size and skill of the field crews as well as conditions on the islands and variation in individual lizard behavior. The lowest
estimated capture probability was 0.23 (SE = 0.10) for females
at U Cay in 1982, whereas the highest estimated capture probability was 0.84 (0.03) for males at U Cay in 2001 (Table 1).
3.2.
Survival probabilities
The two best model structures for survival probability were
k(sex + cay) with 18.7% of the Akaike weight, and k(sex) with
18.2% of the Akaike weight, indicating strong sex and cay effects on survival (Table 2). In addition, the presence of a time
trend (T) in the structure of the next five best models, with
Akaike weights from 12.8% to 7.2%, suggests an apparent,
though weak, time trend in survival.
As predicted, estimated survival was lower for males than
for females (logit-scale effect estimate from the top-ranked
model in which the effect appeared = 0.45; 95% CI = 0.72
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Table 1 – Average recapture probabilities (p) and estimates of abundance (N̂) for Allen Cays iguanas on U Cay and Leaf Cay
Year
Leaf Cay
U Cay
Males
^
p ðSEÞ
1982
1983
1986
1988
1990
1992
1993
1994
1995
1996
1998
2000
2001
2002
2003
2004
0.33
0.53
0.58
0.71
0.80
0.78
0.43
0.69
0.62
0.56
0.58
0.72
0.82
0.80
0.53
0.67
(0.13)
(0.09)
(0.09)
(0.06)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.04)
(0.03)
(0.03)
(0.04)
(0.05)
Females
^ ðSEÞ
N
72
51
78
67
79
86
113
121
119
121
135
134
174
178
163
145
(29)
(10)
(14)
(7)
(6)
(7)
(16)
(11)
(11)
(12)
(13)
(10)
(9)
(10)
(16)
(13)
^p ðSEÞ
0.31
0.50
0.55
0.69
0.78
0.76
0.40
0.67
0.59
0.53
0.55
0.70
0.80
0.79
0.50
0.65
(0.13)
(0.09)
(0.09)
(0.07)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.04)
(0.03)
(0.04)
(0.04)
(0.05)
Males
^ ðSEÞ
N
32
22
38
48
55
66
72
93
84
109
151
145
183
182
142
188
^p ðSEÞ
(14)
(5)
(8)
(6)
(5)
(6)
(11)
(9)
(9)
(12)
(15)
(11)
(10)
(10)
(15)
(17)
0.37
0.57
0.61
0.74
0.82
0.80
0.47
0.72
0.66
0.60
0.61
0.75
0.84
0.83
0.57
0.71
(0.14)
(0.09)
(0.09)
(0.06)
(0.04)
(0.05)
(0.06)
(0.05)
(0.05)
(0.05)
(0.05)
(0.04)
(0.03)
(0.03)
(0.05)
(0.05)
Females
^ ðSEÞ
N
71
46
72
65
57
65
57
94
85
97
80
82
79
80
69
58
(27)
(8)
(12)
(7)
(4)
(5)
(9)
(8)
(8)
(10)
(8)
(6)
(5)
(5)
(8)
(6)
^p ðSEÞ
0.23
0.40
0.44
0.59
0.70
0.67
0.31
0.57
0.49
0.43
0.44
0.60
0.73
0.71
0.40
0.55
(0.10)
(0.09)
(0.09)
(0.08)
(0.06)
(0.06)
(0.05)
(0.06)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.05)
(0.06)
^ ðSEÞ
N
57 (27)
25 (7)
45 (11)
57 (9)
46 (6)
55 (7)
16 (4)
83 (10)
98 (13)
79 (11)
68 (10)
109 (11)
95 (8)
86 (8)
104 (14)
84 (11)
Capture probabilities were estimated based on model-averaged estimates from the fixed-effects Cormack–Jolly–Seber models described in
Section 2.
Table 2 – Model selection results from the CJS model analysis of apparent survival (/) for Allen Cays iguanas on Leaf and U
Cays, Bahamas
Model
/
/
/
/
/
/
/
/
/
/
/
/
/
/
/
(sex + cay) p(t + sex · cay)
(sex) p(t + sex · cay)
(sex + T) p(t + sex · cay)
(sex + cay + T) p(t + sex · cay)
(sex + cay · T) p(t + sex · cay)
(sex · T) p(t + sex · cay)
(sex · T + cay) p(t + sex · cay)
(sex · cay) p(t + sex · cay)
(sex · cay + T) p(t + sex · cay)
(sex · cay · T) p(t + sex · cay)
(.) p(t + sex · cay)
(cay) p(t + sex · cay)
(T) p(t + sex · cay)
(cay + T) p(t + sex · cay)
(cay · T) p(t + sex · cay)
QAICc
DQAICc
Weight
K
3594.23
3594.28
3594.99
3595.20
3595.52
3595.89
3596.13
3596.22
3597.19
3599.08
3602.69
3602.78
3604.00
3604.28
3604.66
0.00
0.05
0.76
0.97
1.29
1.66
1.90
1.99
2.96
4.85
8.46
8.55
9.77
10.05
10.43
0.187
0.182
0.128
0.115
0.098
0.082
0.072
0.069
0.042
0.016
0.003
0.003
0.001
0.001
0.001
23
22
23
24
25
24
25
24
25
28
21
22
22
23
24
Parameters include apparent survival (/) and capture probability (p) as a function of year (t) or time trend (T), sex, and cay. Results include
QAICc, relative QAICc (DQAICc), model weight, and number of parameters (K).
to 0.17), and lower on Leaf Cay than on U Cay (effect estimate = 0.29; 95% CI = 0.74 to 0.07). In addition, the estimated time trend in survival was negative (effect
estimate = 0.02, 95% CI = 0.06, 0.02). Overall, annual survival estimates ranged from 0.866 (SE = 0.022) to 0.933
(SE = 0.016; Table 3).
3.3.
Population growth rate and size
The best model structure for population growth rate (k) in the
Pradel analysis was sex + cay + T (36% of the Akaike weight;
Table 4). In addition, all of the top four structures (Akaike
weights 16–36%, combined weight, 97%) included sex, cay,
and time trend effects, either with or without interactions.
Estimates of k were consistently lower on U Cay than on Leaf
Cay, lower in males than females, and decreased over time,
from estimates over 1.1 early in the study to estimates near
or below 1.0 late in the study (Table 5 and Fig. 2). Despite considerable variation in the population estimates early in the
study, estimates of abundance (Table 1 and Fig. 3) indicated
that subadult/adult males and (particularly) females of both
populations increased over the course of the study, and that
the population on Leaf Cay increased faster than on U Cay.
4.
Discussion
4.1.
Capture probabilities
Estimated capture probabilities supported the impression
that males are more bold than females; i.e., they had a greater
B I O L O G I C A L C O N S E RVAT I O N
305
1 3 2 ( 2 0 0 6 ) 3 0 0 –3 1 0
Table 3 – Annual apparent survival estimates (/), for male and female iguanas on Leaf Cay and U Cay over 16 intervals
based on the Cormack–Jolly–Seber model
Interval
Leaf Cay
1980–1982
1982–1983
1983–1986
1986–1988
1988–1990
1990–1992
1992–1993
1993–1994
1994–1995
1995–1996
1996–1998
1998–2000
2000–2001
2001–2002
2002–2003
2003–2004
U Cay
Males
Females
Males
Females
/ (SE)
/ (SE)
/ (SE)
/ (SE)
0.887
0.885
0.884
0.883
0.882
0.880
0.879
0.878
0.876
0.875
0.873
0.872
0.871
0.869
0.867
0.866
(0.020)
(0.018)
(0.017)
(0.015)
(0.014)
(0.013)
(0.012)
(0.012)
(0.011)
(0.012)
(0.013)
(0.014)
(0.016)
(0.018)
(0.020)
(0.022)
0.928
0.927
0.926
0.925
0.924
0.922
0.921
0.920
0.918
0.917
0.915
0.914
0.912
0.910
0.909
0.907
(0.017)
(0.016)
(0.015)
(0.014)
(0.013)
(0.012)
(0.011)
(0.011)
(0.010)
(0.010)
(0.011)
(0.012)
(0.013)
(0.015)
(0.017)
(0.020)
0.893
0.893
0.892
0.892
0.891
0.890
0.890
0.889
0.888
0.888
0.887
0.886
0.886
0.885
0.884
0.883
(0.019)
(0.017)
(0.016)
(0.015)
(0.013)
(0.013)
(0.012)
(0.012)
(0.012)
(0.013)
(0.014)
(0.015)
(0.016)
(0.018)
(0.019)
(0.021)
0.933
0.932
0.931
0.931
0.930
0.929
0.928
0.927
0.926
0.925
0.924
0.923
0.922
0.921
0.919
0.918
(0.016)
(0.015)
(0.014)
(0.013)
(0.012)
(0.012)
(0.011)
(0.011)
(0.010)
(0.011)
(0.011)
(0.012)
(0.014)
(0.016)
(0.018)
(0.020)
Estimates are model-averaged from models described in Section 2.
Table 4 – Model selection results from the Pradel model analysis of population growth rate (k) for Allen Cays iguanas on
Leaf and U Cays, Bahamas
Model
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
k(sex + cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
p(t + sex · cay)
/
/
/
/
/
/
/
/
/
/
/
/
/
/
/
(sex + cay + T)
(sex + cay · T)
(sex · cay + T)
(sex · T + cay)
(sex · cay · T)
(cay + T)
(cay · T)
(sex + T)
(sex · T)
(T)
(sex + cay)
(sex · cay)
(cay)
(sex)
(.)
QAICc
DQAICc
Weight
K
6482.78
6483.36
6484.10
6484.42
6488.02
6499.37
6500.06
6504.80
6506.44
6520.20
6539.56
6540.92
6550.90
6554.57
6565.29
0.00
0.58
1.32
1.64
5.24
16.59
17.28
22.02
23.66
37.43
56.78
58.15
68.12
71.79
82.52
0.360
0.269
0.186
0.159
0.026
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
0.000
28
29
29
29
32
27
28
27
28
26
27
28
26
26
25
Parameters include k, apparent survival (/), and capture probability (p) as a function of year (t) or time trend (T), sex, and cay. Results include
QAICc, relative QAICc (DQAICc), model weight, and number of parameters (K).
tendency to approach humans. However, it can only be speculated as to why males on the less-visited U Cay would be
easier to capture compared to those on Leaf Cay. It is possible
that the intensity of human interactions on Leaf Cay has
actually made some males more skittish there (Ron et al.,
1998; Kerr et al., 2004; but see Labra and Leonard, 1999 for
the opposite effect). In addition, because there is less feeding
of iguanas on U Cay, greater competition among males on
that island for the more limited supplemental food may make
males more aggressive, and hence, more easily captured. In
addition, the behavioral difference may be related to the geophysical differences between the islands. Most of Leaf Cay is
low and sandy, and habitats there have predominately open
substrates (even in the forest understory). Apart from its cen-
tral open, sandy isthmus, U Cay is mostly rugged honeycomb
limestone (>90%) with nearly impenetrable vegetation. The
physiognomy of Leaf Cay is therefore more conducive to thorough sampling than U Cay. It may be that the bolder males on
U Cay present themselves on the open, sandy isthmus,
whereas the more shy females are able to seek refuge easily
in microhabitats more difficult to sample.
4.2.
Survival probabilities
There was strong support for the predicted differences in survival between the two cays, with higher rates on U Cay than
Leaf Cay, presumably because Leaf Cay experienced much
higher tourist visitation. Thus, these analyses support the
306
B I O L O G I C A L C O N S E RVAT I O N
1 3 2 ( 2 0 0 6 ) 3 0 0 –3 1 0
Table 5 – Estimates of annual population growth rate (k) for male and female iguanas on Leaf Cay and U Cay
Interval
1980–1982
1982–1983
1983–1986
1986–1988
1988–1990
1990–1992
1992–1993
1993–1994
1994–1995
1995–1996
1996–1998
1998-2000
2000–2001
2001–2002
2002–2003
2003–2004
Leaf Cay
U Cay
Males
Females
Males
Females
k (SE)
k (SE)
k (SE)
k (SE)
1.175
1.162
1.148
1.135
1.121
1.108
1.095
1.083
1.070
1.058
1.045
1.033
1.021
1.009
0.997
0.986
(0.024)
(0.022)
(0.019)
(0.017)
(0.015)
(0.013)
(0.012)
(0.010)
(0.010)
(0.009)
(0.010)
(0.010)
(0.011)
(0.013)
(0.014)
(0.016)
1.233
1.218
1.204
1.189
1.175
1.161
1.147
1.133
1.119
1.106
1.092
1.079
1.066
1.053
1.041
1.028
(0.029)
(0.026)
(0.024)
(0.021)
(0.019)
(0.017)
(0.015)
(0.013)
(0.012)
(0.011)
(0.011)
(0.012)
(0.013)
(0.014)
(0.016)
(0.017)
1.124
1.110
1.096
1.082
1.068
1.054
1.041
1.027
1.014
1.001
0.989
0.976
0.963
0.951
0.939
0.927
(0.022)
(0.020)
(0.018)
(0.016)
(0.014)
(0.013)
(0.011)
(0.010)
(0.010)
(0.010)
(0.010)
(0.011)
(0.012)
(0.013)
(0.014)
(0.016)
1.183
1.168
1.152
1.137
1.122
1.108
1.093
1.079
1.065
1.051
1.037
1.023
1.010
0.997
0.984
0.971
(0.026)
(0.024)
(0.021)
(0.019)
(0.017)
(0.015)
(0.014)
(0.012)
(0.012)
(0.011)
(0.011)
(0.012)
(0.013)
(0.014)
(0.015)
(0.016)
Fig. 2 – Annual population growth rate estimates (lambda) and 95% confidence intervals for male (closed circles) and female
(open circles) Allen Cays Iguanas on Leaf and U Cays over the last 25 years.
Fig. 3 – Population size estimates (means ± SE) for male (open circles) and female (closed circles) Allen Cays Iguanas on Leaf
and U Cays for individuals >25 cm snout-vent length over the last 25 years.
B I O L O G I C A L C O N S E RVAT I O N
hypothesis of a negative impact of tourists on survival in
these iguanas, but do not permit the identification of the precise mechanism for that increased mortality (e.g., via improper feeding, introduction of pathogens, poaching, or direct
killing). The survival analysis also indicated a weak, but notable decline in survival over the study period, particularly over
the last five intervals (Table 3). Part of this decline could be explained by the continuing increase in visitation, but it is more
likely related to density-dependent feedbacks in the populations as they reach carrying capacity (see discussion below).
The annual survival estimates of approximately 87–93%
for Cyclura cychlura are similar to those for other long-lived
lizards. For example, Laurie and Brown (in Table 2, 1990a) reported estimates for Galapagos marine iguanas (Amblyrhynchus cristatus) of 83% for females and 79% for males (91%
and 88%, respectively, when an El Niño year was excluded).
In addition, Iverson (1979) estimated annual survival at 90%
(females) and 95% (males) for Cyclura carinata in the Turks
and Caicos Islands. Finally, desert iguanas (Dipsosaurus dorsalis) and chuckwallas (Sauromalus obesus) in southeastern California had estimated annual survival rates of 61% (Krekorian,
1984) and 72–74% (Abts, 1987), respectively.
The finding that mortality in males is higher than in females was corroborated for the Galapagos marine iguana by
Laurie and Brown (1990a,b). However, males are larger than
females in both these species, and Laurie and Brown (1990b)
found that mortality was size-related, such that after controlling for body size, females actually had lower survival probabilities. The higher mortality in males observed in this study
may be related to their boldness and hence their greater exposure to predation (e.g., by osprey, herons, or owls; Iverson,
1979; Hayes et al., 2004), harassment by tourists, and/or
poaching (Smith and Iverson, 2006). An alternative explanation for this pattern might be the result of vigorous aggressive
encounters between males during the breeding season in
May. However, females aggressively defend their nest sites
against all conspecifics, and for a longer period than the mating season (Iverson et al., 2004b), so the basis for the sexual
differences in mortality deserves further investigation.
Preliminary data on survival in adult Caicos Island rock
iguanas (Iverson, 1979) suggested that females experienced
higher mortality rates than males; however, those data were
collected on islands onto which non-native predators had recently been introduced (Iverson, 1979), and which nearly
extirpated the population in about 4 years (Iverson, 1978).
Those results may reflect the increased vulnerability of
nest-guarding female rock iguanas to predation (e.g., see
Hayes et al., 2004) by cats and dogs.
4.3.
Population growth rate and size
Population growth rates were highest for Leaf Cay females,
followed by Leaf Cay males, U Cay females, and U Cay males.
The higher population growth rates on Leaf Cay may be
attributable to the higher levels of supplemental feeding there
(Iverson et al., 2004a,b). However, this result at first seems
inconsistent with the higher survival rates on U Cay, and
the fact that measures of reproductive output (e.g., clutch
size, egg size, and frequency) apparently do not differ between the cays (Iverson et al., 2004b). Furthermore, although
1 3 2 ( 2 0 0 6 ) 3 0 0 –3 1 0
307
primary productivity is certainly higher on Leaf than U Cay
(compare color aerial photographs in Iverson et al. (2004b)),
and rates of food supplementation on Leaf Cay are also
higher, individual growth rates have not been higher on Leaf
Cay than U Cay (Iverson et al., 2004a). The slower rate of population growth on U Cay is therefore likely due to lower rates
of survivorship in juveniles on that cay. Although sufficient
recapture data of juveniles on U Cay to test this assumption
are lacking, it is known that nest mortality on U Cay is higher
than on Leaf Cay in some years, primarily because the nesting
areas are all low and dangerously moist (Iverson et al., 2004b).
Based on estimates of abundance, the population on Leaf
Cay has increased from about 100 subadults and adults
(>25 cm SVL; see Section 2) in 1982 to perhaps 360 in 2002 (Table 1 and Fig. 3). Indeed, 290 and 286 subadults and adults (453
and 487 total individuals) were captured in 2001 and 2002,
respectively. In 2002, based on the proportion of iguanas still
lacking paint marks at the end of the field season, the senior
author subjectively estimated the entire Leaf Cay population
(including juveniles) at about 600, excluding young of the
year. The total mass of 484 of the iguanas actually captured
on Leaf Cay in 2002 was 376.3 kg (a minimum standing crop
biomass of 94.1 kg/ha), which is among the highest values
yet recorded for an iguana (Table 6) or any other lizard (Iverson, 1982; Rodda and Dean-Bradley, 2002).
Similarly, the population on U Cay is estimated to have increased from about 127 subadults and adults in 1982 to about
167 in 2002. For comparison, 136 and 127 subadult and adult
iguanas (166 and 179 total) were actually captured on U Cay
in 2001 and 2002, respectively, and the total population in
2002 was subjectively estimated at about 300, excluding
young of the year. The total mass of 169 of the iguanas captured on U Cay in 2002 was 212.9 (a minimum of 71.0 kg/ha).
It is important to note that estimates of abundance as
computed here are sensitive to the assumption of equal capture probability over individuals (Williams et al., 2001). However, the entire island was searched during each survey in
order to minimize this bias. In addition, there was no trend
toward lower estimates of capture probability early in the
study. Finally, the estimates of population sizes on the two
cays are reasonable given the estimates in 1970 (Carey,
1976), and the actual captures in 2001 and 2002.
These data suggest that the combined Leaf and U Cay subadult and adult populations have increased from only a few
individuals during the first half of the 1900s to perhaps 150
in 1970, to about 230 in 1982, and to about 529 in 2002, a rate
of increase of about 4% per annum since 1970. Although this
is a reassuring recovery rate given the plight of West Indian
iguanas, it is far less than the maximum potential for this taxon. Indeed, a population of eight subadult Allen Cays iguanas
introduced to Alligator Cay in the central Exuma Islands (6 in
1988; 2 in 1990) grew to between 75 and 90 individuals by 1998
(Knapp, 2001), an annual increase of at least 32% (8–75 in 8
years).
Allen Cays rock iguanas are apparently not the only iguanas with such a great potential for population growth. A few
Cuban iguanas (Cyclura nubila) escaped in the mid-1960s to
Isla Magueyes off Puerto Rico (Rivero, 1978), and increased
to perhaps 157 adults and 10 juveniles in the mid-1980s
(Christian, 1986). If it is conjectured that 6–12 individuals
308
B I O L O G I C A L C O N S E RVAT I O N
1 3 2 ( 2 0 0 6 ) 3 0 0 –3 1 0
Table 6 – Standing crop biomass (kg/ha) of populations of West Indian iguanas (genus Cyclura)
Taxon
Location
Age class
Biomass
Source
Cyclura carinata
Caicos Islands
C. cychlura
inornata
Leaf Cay
Leaf Cay
U Cay
Alligator Caya
Guana Cay
Anegada
Adult
Juvenile
All
All
All
All
All
Adult
Juvenile
All
All
All
All
All
17.0
5.2
26
94.1
71.0
61.6b
26
11.6
21.1
3.1–22.5
3.7
23.7–58.9
104.4
109
Iverson (1979)
Iverson (1979)
Carey (1976)
Iverson, unpublished
Iverson, unpublished
Knapp (2001)
Carey (1976)
Carey (1976)
Carey (1976)
Hayes et al. (2004)
Hayes et al. (2004)
Hayes et al. (2004)
Hayes et al. (2004)
Iverson et al., unpublished
C. c. figginsi
C. pinguis
C. r. rileyi
C. r. cristata
C. r. nuchalis
San Salvador
White Cay
Acklin Bight
Bush Hill Caya
Bush Hill Caya
a Introduced populations.
b Minimum estimate based on density (47/ha) and mean body mass of non-founders (1.288 kg).
originally escaped, annual estimates of increase would have
been at least 14–18% over those 20 years. In addition, five Acklins Bight iguanas (Cyclura rileyi nuchalis) were introduced to
Bush Hill Cay in the (current) Exumas Cays Land and Sea Park
in 1973 (Sandra Buckner, unpublished). In 1997 Hayes et al.
(2004) captured 53 individuals on Bush Hill and estimated that
the population had grown to about 300 individuals (a rate of
increase of 19% per year). Based on 250 captures (58 recaptures) between 2002 and 2004, simple Lincoln-Peterson Index
estimates for the island were 315–325 (excluding juveniles) in
May 2004 (Iverson, unpublished). If the total population was
conservatively estimated to be 350 in 2004, an annual rate
of increase of at least 15% is suggested for the 31-year period.
Finally, a population of 8 individuals of the Anegada Iguana
(Cyclura pinguis) introduced to Guana Cay in the British Virgin
Islands in 1984 increased to over 300 individuals by 2004
(Anonymous, 2004), an annual increase of at least 19.9% per
year.
These recovery rates instill optimism in the potential to
bring West Indian iguanas back from the brink of extinction,
provided that direct and indirect human effects on iguana
mortality are eliminated (or at least reduce significantly).
While the populations on these two cays have increased substantially over the course of the study, there is also evidence
of the impact of human disturbance. Of course, reducing human effects on mortality is more easily said than done (Burton, 2004; Wilson et al., 2004; Bradley and Gerber, 2005).
However, this study suggests that establishing assurance colonies of endangered iguanas on small islands near to natural
(presumably disturbed) populations may be extremely costeffective and would likely have a high probability of success
(Lazell, 2000; Towns and Ferreira, 2001; Knapp and Hudson,
2004) if they are sheltered from human disturbance. In addition, this strategy would dilute the impact of a devastating
hurricane. Hence, despite the arguments of some workers
against translocations (Dodd and Seigel, 1991; but see debate
in Burke (1991); and Reinert (1991) , and the closing paragraph
of Dodd and Seigel (1991)), relocations should at least be considered during the development of any recovery plan for an
endangered iguana.
In many respects the West Indian iguanas of the genus
Cyclura would seem to be a classic example of a densityregulated species (Roff, 2002), with the most delayed maturity
of any known lizard, the production of relatively few large
eggs, and parental care in the form of nest guarding (Iverson
et al., 2004b). However, the species is also capable of extremely rapid population growth in island environments with
low competition and predation. Hence, its colonization (or
recolonization) ability mirrors that of species experiencing
much more density-independent selection. The ability of
these lizards to exhibit such exponential population growth
is likely adaptive for these large lizards in island environments influenced by stochastic events like hurricanes in the
short term, and sea level changes in the longer term.
The high values for standing crop biomass and the decline
in lambda over the last 25 years suggest that the Allen Cays
iguana populations may be near carrying capacity (K). This
possibility begs the questions as to what K is for Allen Cays
iguanas on Leaf and U Cays, and how population growth will
be mediated as it is reached. Whether density-dependent effects will be greatest on growth, reproduction, or survival remains to be determined (see Massot et al., 1992).
Comparisons of current life history parameters in the Allen Cays (Iverson et al., 2004a,b; this study) with those from
the recently introduced population on Alligator Cay (Knapp,
2000, 2001), as well as those from future study in the Allen
Cays should address these questions. The effects of the high
levels of food supplementation on Leaf Cay during this period
should also be illuminating.
Acknowledgements
This paper is dedicated to Mrs. Sandra Buckner, Past President
of the Bahamas National Trust, who has supported our work
with iguanas in nearly every imaginable way, from providing
logistical and historical information, obscure literature, aerial
and ground photographs, gear storage, transportation,
accommodations, field assistance, and pleasurable company
in the field and in Nassau for so many years. She has unselfishly supported nearly every herpetological matter in the
B I O L O G I C A L C O N S E RVAT I O N
Bahamas for over 14 years, and has facilitated the pursuit of
knowledge and conservation of reptiles and amphibians in
that country more than any other person.
Financial support for this study was provided by the Center for Field Research (Earthwatch), the Charles and Sally Test
Fund of Earlham College, the Professional Development Fund
of Earlham College, the James Cope Fund of the Joseph Moore
Museum of Natural History, the Lincoln Park Zoo Neotropic
Fund, the Bahamas National Trust, and participating students
and faculty. From 1980 through 2004, 75 students and five faculty at Earlham College, as well as 21 other volunteers provided support and field assistance. A complete list of those
participants as well as the boats and crews who provided
transportation appeared in Iverson et al. (2004a,b). Permits
for field work were issued through the Bahamas Ministry of
Agriculture. W. King provided unpublished field notes. S.
Buckner, K. Hines, and J. Nichols offered constructive criticism of early drafts.
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