|Scientific Name:||Dermochelys coriacea|
|Species Authority:||(Vandelli, 1761)|
|Infra-specific Taxa Assessed:|
Testudo coriacea Vandelli, 1761
|Red List Category & Criteria:||Vulnerable A2bd ver 3.1|
|Assessor(s):||Wallace, B.P., Tiwari, M. & Girondot, M.|
|Reviewer(s):||Chaloupka, M.Y., Dutton, P.H., van Dijk, P.P., Mortimer, J.A., Casale, P., Bolten, A.B., Eckert, K.L., Nel, R., Musick, J.A., Pritchard, P.C.H., Dobbs, K., Miller, J. & Limpus, C.|
|Contributor(s):||Barragan, A., Amorocho, D., Beheret, N., Chacon, D., Chan, E., DeFreitas, R., Diez, C., Eckert, S., Entraygues, M., Felix, M., Formia, A., Garner, J., Girard, A., Guada, H., Hamann, M., Harrison, E., Honarvar, S., Hughes, G., Kalamandeen, M., Kapurusinghe, T., Livingstone, S., Lloyd, C., Lopez-Mendilaharsu, M., Marcovaldi, M., Meylan, A., Nel, R., Patino-Martinez, J., Pilcher, N.J., Reina, R., Sarti, L., Shanker, K., Stewart, K., Tomas, J., Turny, A. & Urteaga, J.|
The global population of Leatherback turtles (Dermochelys coriacea) comprises seven subpopulations (see Figure 2 in attached PDF) that vary widely in population size, geographic range, and population trends, and are the appropriate units for assessment of global conservation status for this species (Wallace et al. 2010, 2011). As such, we performed assessments for each of the seven subpopulations, in addition to the combined global population assessment required by the IUCN (see Table 1 in attached PDF). Due to this species’ geographically widespread distribution, Criterion A was the only appropriate criterion for assessment that could be applied to the global Leatherback population. Generation length was estimated as 30 years. Estimation of global population change based on subpopulation trends (weighted by subpopulation sizes relative to the global population size) produced an estimate of -40.1% decline over the past three generations (see Table 2 in attached PDF), corresponding to a category of Vulnerable based on Criterion A2 (i.e., threats are not reversible nor have they ceased), subcriteria b (i.e., an index of abundance appropriate to the taxon—annual nest counts—was used) and d (i.e., decline was due to actual or potential levels of exploitation). In contrast, assessment of the data under Criterion A4—past, present, and future abundance—revealed that the global Leatherback population trends over three generations will no longer meet thresholds for threat categories by 2020 (-29.4%), and will be increasing by 2030 (3%) and beyond (104% by 2040) (see Table 3 in attached PDF). Therefore, within the next ten years, the global Leatherback population might no longer qualify as “Threatened”—i.e. a category listing of Vulnerable, Endangered, or Critically Endangered—according to the IUCN Red List Criteria.
However, future population increases will be dependent on the success of conservation actions mitigating current and future threats to this species throughout its range, especially in breeding and foraging areas, and on no new threats arising (e.g. climate change) that could cause population declines. Moreover, nearly 99% of the global population (based on data available currently) will be contained within the Northwest Atlantic, which obscures the declines of the Pacific subpopulations and threatened status of other relatively small subpopulations (e.g. Southwest Atlantic, Southwest Indian) (see Figure 3 in attached PDF), not to mention data deficient subpopulations (Northeast Indian, Southeast Atlantic). While these results demonstrate that Leatherbacks, as a single taxonomic entity, will not go extinct globally in the next generation, the global listing is not an appropriate representation of the conservation status of the biologically relevant subpopulations that make up the global Leatherback population. For this reason, the subpopulation-level Red List assessments for Leatherbacks should be given priority in evaluating the true global extinction risk for this species.
Comprehensive analyses of 39 existing datasets—including 28 time series datasets with ≥10 years of data—of abundance of nesting females or their nesting activities on beaches revealed different population trends among subpopulations, but a global decline overall based on subpopulation trends weighted by subpopulation size three generations ago relative to combined global population size (see Table 2 in attached PDF for all datasets used). Overall, considering only datasets with ≥10 yr of abundance data, the total global abundance across Leatherback subpopulations had declined from 90,599 nests yr-1 to 54,262 nests yr-1 over three generations until 2010. Using average conversion factors from different subpopulations to provide bracketed estimates of nesting female abundance (i.e. 5 and 7 clutches per female, three years for re-migration intervals, intermediate between subpopulation averages; TEWG 2007, Reina et al. 2002), these annual nesting abundance values correspond to approximately 12,943-18,120 nesting females yr-1 (or 38,828-54,359 adult females) three generations ago and 7,752-10,852 nesting females yr-1 (or 23,255-32,557 adult females) in 2010. Because we did not include abundance data from several rookeries that had <10 yr of data—including Gabon, which is the largest Leatherback rookery in the world (Witt et al. 2009), as well as Grand Riviere and Fishing Pond, Trinidad, Armila, Panama, and Gulf of Urubá, Colombia (TEWG 2007, Patiño-Martínez et al. 2008)—in our annual and total abundance subpopulation and global abundance estimates, these values should be considered conservative estimates.
Overall, although the global listing for Leatherbacks was derived by abundance-weighted analyses of three-generation trends for all subpopulations with available data, it is not an appropriate representation of conservation status of biologically relevant population units (i.e. subpopulations) that make up the global Leatherback population. Subpopulation assessments demonstrated wide variation not only in status of individual subpopulations (as indicated by the Red List Categories), but also in how categories were derived for each subpopulation (as indicated by the Red List Criteria used) (see Table 1 in attached PDF). Specifically, final threat categories were triggered by thresholds under different criteria, depending on whether a subpopulation has declined significantly over time (e.g. East Pacific, West Pacific), was geographically constrained (e.g. Southwest Indian Ocean), or characterized by small (e.g. Southwest Indian Ocean) or very small population sizes (e.g. Southwest Atlantic Ocean). Therefore, the variation in conservation status among subpopulations warrants preference for subpopulation assessments over the global assessment when evaluating and describing the global conservation status of Leatherbacks.
Presently, the Northwest Atlantic subpopulation—i.e. from Florida, USA throughout the Wider Caribbean—is large and increasing (TEWG 2007) (Table 2 in attached PDF). Furthermore, the Southeast Atlantic subpopulation—i.e. West Africa, especially Gabon—is the largest in the world (Witt et al. 2009), but long-term trend data are not available for this assessment (TEWG 2007). The existence of these large (and increasing in at least one case) subpopulations makes it extremely unlikely that Leatherbacks globally will go extinct in the near future. If current trends in the Northwest Atlantic subpopulation continue, the global population trend might no longer meet thresholds for IUCN Threatened categories (i.e., Vulnerable, Endangered, Critically Endangered) within ten years (Table 3 in attached PDF). In fact, if current trends continue, future global population sizes are projected to increase to 184,662 nests yr-1 (approximately 26,380-36,932 females yr-1, 79,141-110,797 adult females total) within one generation (i.e. by 2040). The projected abundance of the Northwest Atlantic subpopulation alone will account for nearly 99% of the global Leatherback population abundance by that time (Figure 3 and Table 3 in attached PDF), and increase from 46% of historical global population abundance three generations ago.
In spite of the large and increasing Northwest Atlantic subpopulation, the magnitudes of declines for the East Pacific subpopulation (i.e., which nests along the Pacific coast of the Americas) and, to a slightly lesser extent, the West Pacific subpopulation (i.e., Malaysia, Indonesia, Papua New Guinea, Solomon Islands) over three generations were the main driver the global decline result. Specifically, the East Pacific subpopulation has declined by > 97% over three generations (from >35,000 nests yr-1 to <1,000 nests yr-1, or >5,000 nesting females yr-1 to < 140 nesting females yr-1) (Eckert 1993, Santidrián Tomillo et al. 2007, Sarti Martínez et al. 2007), and its historic abundance accounted for roughly 39% of the estimated global abundance three generations ago (Table 1 in attached PDF). Similarly, the West Pacific subpopulation has declined >80% over three generations, from >12,000 nests yr-1 (2,600 females yr-1) to <2,200 nests yr-1 (<500 females yr-1) (Eckert 1993, Chan and Liew 1996, Dutton et al. 2007, Hitipeuw et al. 2007, Tapilatu et al. 2013), and its historic abundance accounted for 14% of the estimated global abundance three generations ago. Both of these subpopulations are projected to decline further in coming decades (Table 3 in attached PDF). Population declines in these subpopulations (and others) have been attributed to extensive egg harvest and mortality due to incidental capture in fishing gear (Eckert 1993, Wallace and Saba 2009, Tapilatu et al. 2013). Considering the precedent of the collapse of the historically large Pacific subpopulations, the persistence of significant threats in all regions (see Wallace et al. 2011 and Eckert et al. 2012 for review) warrants concern for the future viability of even the largest subpopulations. Current efforts to protect Leatherbacks, their offspring, and their habitats must be maintained—or even augmented, where possible—to reverse declines in Pacific and Indian Ocean subpopulations and to sustain population growth in the Northwest Atlantic.
To explore future projections of Leatherback subpopulation and global abundance and trends, we also applied Criterion A4, which analyzes the global population trend within time intervals from the past, present, and future (Table 3 in attached PDF). Because the length of three generations (~90 yr) was greater than the length of our available datasets, we used the historical abundance values shown in Tables 3 and 4 (see attached PDF) as the baseline population sizes for “moving window” analyses. According to our assessment of the data under Criterion A4, it would no longer qualify as “Threatened” according to IUCN Red List Criteria by 2020 (three-generation decline of 29.4%), and would qualify as Least Concern by 2030 (3% increase) under IUCN Guidelines (IUCN 2011) (Table 3 in attached PDF). This discrepancy between A4 and A2 illustrates a dichotomy in how extinction risk is assessed by these two criteria, despite the same empirical data being used for assessments. Whereas A2 assesses percent decline over time, using the historical abundance as the baseline, A4 uses historical as well as current abundance to project future abundance, and as such accounts for recent growth in subpopulations in the global population estimate. Because the IUCN Guidelines stipulate that multiple criteria be evaluated, the criterion that triggers the highest threat category must be selected for the official assessment. Therefore, Leatherbacks globally are considered Vulnerable (A2b,d) based on IUCN Guidelines (IUCN 2011).
Nonetheless, these results further illustrate the inability of the global assessment to adequately characterize variation among subpopulations. The likelihood of global extinction of this species is extremely low (Tables 2 and 3 in attached PDF); however, the Northwest Atlantic subpopulation alone is projected to account for nearly 99% of total global Leatherback abundance by 2040, with four other subpopulations for which data were available for this assessment accounting for the remaining 1% (Figure 3 and Table 3 in attached PDF). Such a global assessment produces a result indicating no risk of species-level extinction based on the mere existence of any Leatherback subpopulation, while obscuring the dire conservation situation of declining or extinct subpopulations (e.g., the Pacific subpopulations). This type of assessment and its result are contrary to the IUCN SSC Marine Turtle Specialist Group’s (MTSG) mission to guide conservation of marine turtles and their ecological roles. For these reasons, the MTSG defined subpopulations (i.e., regional management units; Wallace et al. 2010) for all marine turtle species, including Leatherbacks, to provide a biologically described framework for evaluating conservation status of appropriate population segments. A global marine turtle population is an amalgam of its subpopulations, all of which are valuable for maintaining the health of the “species” in terms of evolutionary history, genetic diversity, and life history variations. For these reasons, the inability of marine turtle global assessments to appropriately account for variation among subpopulations (see Seminoff and Shanker 2008 for review) further supports the primacy of subpopulation assessments when describing conservation status of marine turtles globally.
Previous global Leatherback assessments (Pritchard 1982, Spotila et al. 1996), including the previous Leatherback Red List assessment (Sarti Martinez 2000), also described global declines in Leatherbacks. However, these previous assessments lacked the geographic breadth and rigour of time series datasets accessed for the current assessment. Spotila et al. (1996) estimated the global Leatherback population in 1995 to be approximately 34,500 adult females (range: 26,200-42,900 females), and stated that this value was roughly a third of the only existing global estimate at the time (115,000 adult females; Pritchard 1982). However, there were several significant assumptions in the Pritchard (1982) estimate that make its use as a reliable global population estimate tenuous, not the least of which was the lack of robust, nesting beach-based estimates of the East Pacific subpopulation abundance on which the regional and global estimates were constructed (see Mrosovsky 2003 for review). Sarti Martinez (2000) assessed the global Leatherback population as Critically Endangered largely on the evidence of significant declines in the East Pacific subpopulation (based on the Pritchard  and Spotila et al.  estimates) and Malaysian rookery, despite some indications of large populations in other regions, such as the Northwest and Southeast Atlantic.
In addition, abundance of several rookeries has increased substantially in the time since these assessments were conducted. For example, estimates for Trinidad, St. Croix, Puerto Rico, and Florida used by Spotila et al. (1996) were all substantially lower—by an order of magnitude in some cases—than current estimates (Table 3 in attached PDF). Furthermore, the current assessment only included time series datasets of a decade or longer in abundance estimates, whereas Spotila et al. (1996) included abundance estimates based on variable numbers of nesting seasons. The 2010 abundance estimates presented in the current assessment are generally lower, but overlap with the 1995 range in estimated abundance reported by Spotila et al. (1996). If abundance estimates from all rookeries with available nesting data (e.g. Panama [~8,000 nests yr-1], Colombia [2,300 nests yr-1], Trinidad [>40,000 nests yr-1, and especially Gabon [36,185-126,480 nests yr-1]; Table 3 in attached PDF) were included in the 2010 assessment, the total global abundance of Leatherbacks in 2010 would be much higher than the estimate shown in Table 3, and would exceed the average estimate of Spotila et al. (1996). However, this is not necessarily indicative of a global population increase, but rather an increase in available information. Therefore, given the availability of new datasets from all Leatherback subpopulations globally, use of previous estimates (Pritchard 1982, Spotila et al. 1996) to characterize global Leatherback population abundance is no longer appropriate.
Interpreting the current assessment (2010) in the context of previous assessments and available data at present, the global Leatherback population has not collapsed, and, in fact, appears to be larger than estimated in previous global assessments (Spotila et al. 1996). This result appears counter-intuitive with the Vulnerable category to which this subpopulation has been assigned, but it can be traced to the nature of the procedure for applying Criterion A to long-lived, widely distributed species like marine turtles (see Seminoff and Shanker 2008 for review). The change in IUCN Red List Category from Critically Endangered (in 2000) to Vulnerable (assessment year 2010) is mostly due to new datasets made available for assessment, and associated detection of increasing trends that were not reported previously, as well as a rigorous application of the IUCN 3.1 Criteria. The continued severe declines of the Pacific subpopulations require urgent and effective conservation interventions to prevent complete collapse (Spotila et al. 2000, Santidrián Tomillo et al. 2007, Sarti Martínez et al. 2007, Tapilatu et al. 2013). Furthermore, the projected increase in the global population is predicated entirely on continued growth of the Northwest Atlantic subpopulation alone, which itself depends on sustained conservation efforts to protect Leatherbacks, their offspring, and their habitats.
Given the widespread, long-lived nature of Leatherback turtles, Criterion A (i.e., decline in population of mature individuals over time) is the only appropriate criterion that could be used for this assessment; the restricted geographic range and small population size criteria (Criterion B, C, and D) did not apply, and no population viability analysis was available (Criterion E). However, Criteria A-D were applied in subpopulation assessments (see Leatherback subpopulation assessments for details; Table 1 in attached PDF).
We obtained time series datasets of abundance of nesting females or their nesting activities collected on Leatherback rookeries (i.e., nesting beaches) around the world, and organized data by subpopulations. For marine turtles, annual counts of nesting females and their nesting activities (more often the latter) are the most frequently recorded and reported abundance metric across index monitoring sites, species, and geographic regions (NRC 2010). Conversion from number of nests to number of nesting females requires estimates of clutch frequency (i.e. number of clutches per nesting female per breeding season) (e.g. Reina et al. 2002), which are not available for all rookeries and subpopulations. Therefore, we presented and analysed all abundance data in numbers of confirmed nests yr-1, as this metric was the most commonly available (Table 2 in attached PDF). See Sources of Uncertainty for discussion of caveats associated with these conversion factors.
We calculated annual and overall population trends for each rookery within a subpopulation, and then calculated average subpopulation trends by weighting rookery population trends by rookery abundance 3 generations ago relative to subpopulation abundance three generations ago. We then repeated this step to derive the global population trend by weighting subpopulation trends by subpopulation abundance three generations ago relative to global abundance three generations ago (Table 2 in attached PDF). We only included time series datasets of ≥10 yr in trend estimations, although we included all rookeries for which we obtained abundance values in the overall summary tables (Table 2 in attached PDF). Because rookeries represent varying proportions of total subpopulation sizes, we ensured that time series from major rookeries within each subpopulation were included in analyses such that the majority of subpopulation abundance was represented. In cases where that was not possible (e.g. Southeast Atlantic, Northeast Indian), we did not derive a subpopulation trend, and such cases were excluded from calculation of global trends.
The most recent year for available abundance data across all rookeries and subpopulations was 2010. Where time series ended prior to 2010, we estimated population sizes for each rookery through 2010 based on the population trend for existing years (e.g. French Guiana, Trinidad). Furthermore, if a longer time series for a rookery within a subpopulation was available that reflected a trend not captured by shorter time series, we estimated historical abundance to calculate overall declines for that subpopulation. For example, abundance data for three of five index sites in the Mexican Pacific—the East Pacific subpopulation—begin in the early 1980s, while the remaining sites (i.e., Barra de la Cruz and Cahuitán, Oaxaca) begin in the early 1990s (Table 2 in attached PDF). All other sites in Mexico, as well as other sites within the same subpopulation (i.e., those in Costa Rica), showed a decline of >97%, whereas the Barra de la Cruz and Cahuitán showed much less dramatic declines, because the time series began after the broader population decline had already begun to occur. Given the synchrony in inter-annual abundance fluctuations and historical reports of high abundance among these rookeries (Eckert 1993), we assumed that the abundance at Barra de la Cruz and Cahuitán was similar to that of other Mexican rookeries at the beginning of those time series, i.e., 1982 (L. Sarti Martínez pers. comm.). This allowed us to standardize trend and abundance estimates within the Mexican rookeries.
To apply Criterion A, three generations (or a minimum of ten years, whichever is longer) of abundance data are required (IUCN 2011). In the case of the Leatherback, we conservatively estimate its generation time as 30 years (see below). For criterion A2, data from three generations ago (~100 yr) are necessary to estimate population declines beginning three generations ago up to the present (i.e. assessment) year. The challenges of this requirement on long-lived species like turtles—with generation lengths of 30 yr or more—are obvious (see Seminoff and Shanker 2008 for review). Abundance data from ~100 yr ago are not available for Leatherbacks anywhere in the world. Extrapolating backward using population trends based on current datasets was considered inappropriate because estimates produced would be biologically unrealistic and unsubstantiated, given what is currently known about sea turtle nesting densities on beaches and other factors (Mrosovsky 2003). In the absence of better information, we assumed that population abundance three generations (~100 years, one generation estimated 30 yr; see below) ago was similar to the first observed abundance rather than to assume that the population has always been in a decline (or increase) of the same magnitude as in the current generation (Table 2 in attached PDF). A similar approach was used in the Red List assessment of another long-lived, geographically widespread taxon, the African Elephant (Blanc 2008). Thus, to apply Criterion A to subpopulations (see separate subpopulation assessments) and the global population, we assumed that the abundance at the beginning of an available time series dataset had not changed significantly in the preceding three generations, and therefore used the same abundance value in trend calculations.
We also evaluated to the global population against Criterion A4 (Table 3 in attached PDF), using the same overall scheme as described above. Criterion A4 permits for analysis of population trend during a “moving window” of time, i.e. over three generations, but where the time window must include the past, present, and future. Furthermore, multiple time-frames are to be examined, and the maximum decline calculated for a given time-frame is to be compared to the thresholds (IUCN 2011). Therefore, we made the same assumption about earliest available historical abundance being equivalent to the population abundance for generations past, and estimated future population abundance in 2020, 2030, and 2040, which all fall within the next generation (i.e., 30 yr). These future projections assume that the derived population trend will continue without deviation during the next generation. Implicit in this assumption is that no changes to degree of threats or conservation efforts impacting rookeries, subpopulations, or the global population will occur during that time. Based on available information, threats to Leatherbacks globally that have caused observed declines have not ceased and are not reversible (see Wallace et al. 2011 and Eckert et al. 2012 for review), making this a reasonable assumption in the absence of better information. The global population will no longer be “Threatened” according to IUCN Red List thresholds by 2020 (-29.4% decline over three generations), and will be increasing by 2030 (3% increase) and beyond (104% increase by 2040)—i.e. IUCN category of Near Threatened or Least Concern—according to Criterion A4, depending on the “time window” applied (Table 3 in attached PDF). This result is due mainly to the currently large and growing Northwest Atlantic subpopulation, and in spite of the lack of sufficient information for the Southeast Atlantic subpopulation and the continued declines in the East Pacific and West Pacific subpopulations.
Estimating Generation Length:
Leatherback age at maturity is uncertain, and estimates range widely (see Jones et al. 2011 for review). Reported estimates fall between 9-15 yr, based on skeletochronology (Zug and Parham 1996), and inferences from mark-recapture studies (Dutton et al. 2005). Furthermore, updated skeletochronological analyses estimated Leatherback age at maturity to be between 26-32 yr (mean 29 yr) (Avens et al. 2009). Extrapolations of captive growth curves under controlled thermal and trophic conditions suggested that size at maturity could be reached in 7-16 yr (Jones et al. 2011). Thus, a high degree of variance and uncertainty remains about Leatherback age at maturity in the wild. Likewise, Leatherback lifespan is unknown. Long-term monitoring studies of Leatherback nesting populations have tracked individual adult females over multiple decades (e.g. Santidrián Tomillo et al. unpublished data, Nel and Hughes unpublished data), but precise estimates of reproductive lifespan and longevity for Leatherbacks are currently unavailable.
The IUCN Red List Criteria define generation length to be the average age of parents in a population; older than the age at maturity and younger than the oldest mature individual (IUCN 2011). Thus, for the purposes of this assessment, we estimated generation length to be 30 yr, or equal to the age at maturity (estimated to be 20 yr on average), plus a conservative estimate of reproductive half-life of 10 yr, as assumed by Spotila et al. (1996).
Sources of Uncertainty
Although monitoring of nesting activities by adult female sea turtles is the most common metric recorded and reported across sites and species, globally, there are several disadvantages to using it as a proxy for overall population dynamics, some methodological, some interpretive (NRC 2010). First, because nesting females are a very small proportion of a sea turtle population, using abundance of nesting females and their activities as proxies for overall population abundance and trends requires knowledge of other key demographic parameters (several mentioned below) to allow proper interpretation of cryptic trends in nesting abundance (NRC 2010). However, there remains great uncertainty about most of these fundamental demographic parameters for Leatherbacks, including age at maturity (see Jones et al. 2011 for review), generation length, survivorship across life stages, adult and hatchling sex ratios, and conversion factors among reproductive parameters (e.g., clutch frequency, nesting success, re-migration intervals, etc.). These values can vary among subpopulations, further complicating the process of combining subpopulation abundance and trend estimates to obtain global population abundance and trend estimates, and contributing to the uncertainty in these estimates. Second, despite the prevalence of nesting abundance data for marine turtles, monitoring effort and methodologies can vary widely within and across study sites, complicating comparison of nesting count data across years within sites and across different sites as well as robust estimation of population size and trends (SWOT Scientific Advisory Board 2011). For example, monitoring effort on Matura beach, Trinidad, has changed multiple times since the early 1990s, which necessitated a modelling exercise to estimate a complete time series for years with reliable monitoring levels (Table 2 in attached PDF). Furthermore, there was a general lack of measures of variance around annual counts provided for the assessment, which could be erroneously interpreted as equally high confidence in all estimates. Measures of variance around annual counts would provide information about relative levels of monitoring effort within and among rookeries, and thus reliability of resulting estimates. For all of these reasons, results of this assessment of global population decline should be considered with caution. For further reading on sources of uncertainty in marine turtle Red List assessments, see Seminoff and Shanker (2008).
Leatherbacks are distributed circumglobally, with nesting sites on tropical sandy beaches and foraging ranges that extend into temperate and sub-polar latitudes (see Figure 1 in attached PDF and global distribution map). See Eckert et al. (2012) for review of Leatherback geographic range.
Native:Albania; American Samoa (American Samoa); Angola (Angola); Anguilla; Antigua and Barbuda; Argentina; Aruba; Australia; Bahamas; Bahrain; Bangladesh; Barbados; Belize; Benin; Bermuda; Bonaire, Sint Eustatius and Saba (Saba, Sint Eustatius); Bosnia and Herzegovina; Brazil; Brunei Darussalam; Cambodia; Cameroon; Canada; Chile; China; Colombia; Comoros; Congo; Congo, The Democratic Republic of the; Costa Rica; Côte d'Ivoire; Croatia; Cuba; Curaçao; Cyprus; Dominica; Dominican Republic; Ecuador; Egypt; El Salvador; Equatorial Guinea; Eritrea; Fiji; France (Clipperton I., France (mainland)); French Guiana; French Polynesia; French Southern Territories (Mozambique Channel Is.); Gabon; Gambia; Ghana; Greece; Grenada; Guadeloupe; Guam; Guatemala; Guinea; Guinea-Bissau; Guyana; Haiti; Honduras; India; Indonesia; Ireland; Italy; Jamaica; Japan; Kenya; Kiribati; Korea, Democratic People's Republic of; Korea, Republic of; Lebanon; Liberia; Libya; Madagascar; Malaysia; Marshall Islands; Martinique; Mauritania; Mauritius; Mayotte; Mexico; Micronesia, Federated States of ; Montenegro; Montserrat; Morocco; Mozambique; Myanmar; Namibia; Netherlands Antilles (Bonaire); New Caledonia; New Zealand; Nicaragua; Nigeria; Northern Mariana Islands; Palau; Panama; Papua New Guinea; Peru; Philippines; Portugal; Puerto Rico; Russian Federation; Saint Helena, Ascension and Tristan da Cunha; Saint Kitts and Nevis; Saint Lucia; Saint Martin (French part); Saint Vincent and the Grenadines; Samoa; Sao Tomé and Principe; Senegal; Seychelles; Sierra Leone; Sint Maarten (Dutch part); Slovenia; Solomon Islands; South Africa; Spain; Sri Lanka; Suriname; Syrian Arab Republic; Taiwan, Province of China; Tanzania, United Republic of; Thailand; Togo; Tonga; Trinidad and Tobago; Tunisia; Turks and Caicos Islands; Tuvalu; United Kingdom; United States; Uruguay; Venezuela, Bolivarian Republic of; Virgin Islands, British; Virgin Islands, U.S.
|FAO Marine Fishing Areas:||
Atlantic – western central; Atlantic – southwest; Atlantic – eastern central; Atlantic – northeast; Atlantic – northwest; Atlantic – southeast; Indian Ocean – western; Indian Ocean – eastern; Mediterranean and Black Sea; Pacific – southwest; Pacific – western central; Pacific – northeast; Pacific – eastern central; Pacific – northwest; Pacific – southeast
|Range Map:||Click here to open the map viewer and explore range.|
Leatherbacks are a single species globally comprising seven biologically described regional management units (RMUs; Wallace et al. 2010), which describe biologically and geographically explicit population segments by integrating information from nesting sites, mitochondrial and nuclear DNA studies, movements and habitat use by all life stages. RMUs are functionally equivalent to IUCN subpopulations, thus providing the appropriate demographic unit for Red List assessments. There are seven Leatherback RMUs (hereafter subpopulations): Northwest Atlantic Ocean, Southeast Atlantic Ocean, Southwest Atlantic Ocean, Northeast Indian Ocean, Southwest Indian Ocean, East Pacific Ocean, and West Pacific Ocean (Figure 2 in attached PDF). Multiple genetic stocks have been defined according to geographically disparate nesting areas around the world (Dutton et al. 1999, 2013), and are included within RMU delineations (Wallace et al. 2010; shapefiles can be viewed and downloaded at: http://seamap.env.duke.edu/swot).
|Habitat and Ecology:||
D. coriacea is an oceanic, deep-diving marine turtle inhabiting tropical, subtropical, and subpolar seas. Leatherbacks make extensive migrations between different feeding areas at different seasons, and to and from nesting areas. Leatherbacks feed predominantly on jellyfishes, salps and siphonophores. Females usually produce several (3-10) clutches of 60-90 eggs in a reproductive season, and typically have a re-migration interval of multiple years (2+) between subsequent reproductive seasons. For a thorough review of Leatherback biology, please see Eckert et al. (2012).
|Use and Trade:||Leatherback eggs and animals are taken for human use (i.e. consumption and commercial products), eggs are also eaten by domestic animals (e.g. dogs).|
Threats to Leatherbacks vary in time and space, and in relative impact to populations. Threat categories affecting marine turtles, including Leatherbacks, were described by Wallace et al. (2011) as:
1) Fisheries bycatch: incidental capture of marine turtles in fishing gear targeting other species;
2) Take: direct utilization of turtles or eggs for human use (i.e. consumption, commercial products);
3) Coastal Development affecting critical turtle habitat: human-induced alteration of coastal environments due to construction, dredging, beach modification, etc.;
4) Pollution and Pathogens: marine pollution and debris that affect marine turtles (i.e. through ingestion or entanglement, disorientation caused by artificial lights), as well as impacts of pervasive pathogens (e.g. fibropapilloma virus) on turtle health;
5) Climate change: current and future impacts from climate change on marine turtles and their habitats (e.g. increasing sand temperatures on nesting beaches affecting hatchling sex ratios, sea level rise, storm frequency and intensity affecting nesting habitats, etc.).
The relative impacts of individual threats to all Leatherback subpopulations were assessed by Wallace et al. (2011). Fisheries bycatch was classified as the highest threat to Leatherbacks globally, followed by human consumption of Leatherback eggs, meat, or other products, and coastal development. Due to lack of information, pollution and pathogens was only scored as affecting three subpopulations and climate change was only scored for two subpopulations. Enhanced efforts to assess and reduce the impacts of these threats on Leatherbacks—and other marine turtle species—should be a high priority for future conservation efforts.
Leatherbacks are protected under various Conventions, national and international laws, treaties, agreements, and memoranda of understanding. A partial list of international conservation instruments that provide legislative protection for Leatherbacks are: Annex II of the SPAW Protocol to the Cartagena Convention (a protocol concerning specially protected areas and wildlife); the Leatherback’s inclusion in Appendix I of CITES (Convention on International Trade in Endangered Species of Wild Fauna and Flora); and Appendices I and II of the Convention on Migratory Species (CMS); the Inter-American Convention for the Protection and Conservation of Sea Turtles (IAC); the Memorandum of Understanding on the Conservation and Management of Marine Turtles and their Habitats of the Indian Ocean and South-East Asia (IOSEA); the Memorandum of Understanding on ASEAN Sea Turtle Conservation and Protection; and the Memorandum of Understanding Concerning Conservation Measures for Marine Turtles of the Atlantic Coast of Africa.
Long-term efforts to reduce or eliminate threats to Leatherbacks on nesting beaches have been successful in many places (e.g. Dutton et al. 2005, Chacón-Chaverri and Eckert 2007, Santidrián Tomillo et al. 2007, Sarti Martínez et al. 2007) but not all places (e.g. Chan and Liew 1996). Reducing Leatherback bycatch has become a primary focus for many conservation projects around the world, and some mitigation efforts are showing promise (Watson et al. 2005; Gilman et al. 2006, 2011). However, threats to Leatherbacks—bycatch mortality and egg consumption, in particular—persist, and in some places, continue to hinder population recovery (Alfaro-Shigueto et al. 2011, 2012; Tapilatu et al. 2013; Wallace et al. 2013). For depleted Leatherback populations to recover, the most prevalent and impactful threats must be reduced wherever they occur, whether on nesting beaches or in feeding, migratory, or other habitats (Bellagio Report 2007; Wallace et al. 2011, 2013); a holistic approach that addresses threats at all life history stages needs to be implemented (Dutton and Squires 2011). Therefore, current conservation efforts, legal protections, and resources supporting those mechanisms must be maintained—and augmented, wherever possible—to reverse population declines and sustain stable and increasing population trends among Leatherback subpopulations. Regional and local efforts to protect Leatherbacks, their offspring, and their habitats should be designed to address threats at appropriate scales, and implemented with participation of appropriate stakeholders.
|Citation:||Wallace, B.P., Tiwari, M. & Girondot, M. 2013. Dermochelys coriacea. The IUCN Red List of Threatened Species. Version 2014.2. <www.iucnredlist.org>. Downloaded on 23 September 2014.|
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