|Scientific Name:||Dermochelys coriacea (Northwest Atlantic Ocean subpopulation)|
|Species Authority:||(Vandelli, 1761)|
|Red List Category & Criteria:||Least Concern ver 3.1|
|Assessor/s:||Tiwari, M., Wallace, B.P. & Girondot, M.|
|Reviewer/s:||Bolten, A.B., Casale, P., Chaloupka, M.Y., Dobbs, K., Dutton, P.H., Eckert, K.L., Mortimer, J.A., Musick, J.A., Nel, R., Pritchard, P.C.H., van Dijk, P.P. & Miller, J. & Limpus, C.|
|Contributor/s:||Meylan, A., Patino-Martinez, J., Turny, A., Tomas, J., Stewart, K., Chacon, D., DeFreitas, R., Diez, C., Eckert, S., Entraygues, M., Felix, M., Garner, J., Guada, H., Harrison, E., Kalamandeen, M., Livingstone, S. & Lloyd, C.|
The Northwest Atlantic Leatherback subpopulation nests in the southeastern U.S.A., throughout the mainland and insular Caribbean, and the Guiana Shield, and marine habitats extend throughout the North Atlantic, including the Gulf of Mexico, north beyond 50N, into the Mediterranean, and across the equator to northwestern Africa (Figure 1 in attached PDF). Several genetic nesting stocks have been identified within this subpopulation, but metapopulation dynamics support its designation as a single subpopulation, or regional management unit (Wallace et al. 2013, Dutton et al. 2013). Based on long-term time series datasets of abundance—i.e. annual counts of nesting females and nests—this Northwest Atlantic subpopulation is large (>50,000 nests yr-1, ~10,000 females yr-1; Table 1 in attached PDF) and has increased by 20.6% over the past three generations, and is projected to increase to >180,000 nests yr-1 in the next generation (by 2040) (Table 2 in attached PDF). Therefore, the Northwest Atlantic subpopulation is considered Least Concern under current IUCN Red List Criteria. However, future population increases depend on the success of current conservation efforts to protect Leatherbacks, their offspring, and their habitats being maintained—or augmented, wherever possible—throughout their enormous geographic distribution, but particularly in breeding and foraging areas, to ensure that current and future threats do not reach levels capable of causing population decreases. Thus, this “Least Concern” status should be considered as entirely conservation-dependent.
Comprehensive analyses of 17 existing datasets—including 14 time series datasets with ≥10 years of data—of abundance of nesting females or their nesting activities on beaches revealed different rookery trends within the Northwest Atlantic Leatherback subpopulation, but an overall subpopulation increase based on rookery trends weighted by rookery size relative to subpopulation size three generations ago (see Table 1 in attached PDF for all datasets used). Presently, the Northwest Atlantic subpopulation—i.e. from Florida, USA, throughout the Wider Caribbean—is large and increasing (Table 1 in attached PDF). To explore future projections of Northwest Atlantic Leatherback subpopulation abundance and trends, we also applied Criterion A4, which analyses the subpopulation trend within time intervals from the past, present, and future (Table 2 in attached PDF). According to our assessment of the data under Criterion A4, the Northwest Atlantic subpopulation will continue to increase over the next generation (Table 3 in attached PDF).
However, the precedent of the collapse of the historically large Pacific subpopulations (Santidrián Tomillo et al. 2007, Sarti Martínez et al. 2007, Tapilatu et al. 2013), is cause for concern for the Northwest Atlantic subpopulation. Significant threats persist in the Northwest Atlantic in nesting and foraging areas (see Wallace et al. 2011 and Eckert et al. 2012 for review), and this subpopulation’s enormous geographic distribution spans dozens of national and international jurisdictions, making effective conservation extremely challenging. Current efforts to protect Leatherback and their offspring, particularly in critical habitats already identified (e.g. Nova Scotia, Canada, nesting beaches in and adjacent waters off Trinidad, the Guiana Shield, etc.) must be maintained and even increased to sustain population growth in the Northwest Atlantic and prevent similar population declines observed for other Leatherback subpopulations.
Given the widespread, long-lived nature of Leatherbacks, and the Northwest Atlantic subpopulation in particular, Criterion A (i.e., decline in population of mature individuals over time) was the only appropriate criterion that could be used for this assessment; the restricted geographic range and small population size criteria (Criteria B, C and D) did not apply, and no population viability analysis was available (Criterion E).
Criterion A: 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. Briane et al. 2007, TEWG 2007), which are not available for all rookeries in this subpopulation. Therefore, we presented and analysed all abundance data in numbers of confirmed nests yr-1, as this metric was the most commonly available (Table 1 in attached PDF). See Sources of Uncertainty for discussion of caveats associated with these conversions.
We calculated annual and overall abundance trends for each rookery within the subpopulation, and then calculated the overall subpopulation trend by weighting rookery population trends by rookery abundance three generations ago relative to subpopulation abundance three generations ago. We only included time series datasets of ≥10 yr in trend estimations, although we included all rookeries for which we were obtained abundance values in the overall summary tables (see Table 1 in attached PDF). However, rookeries included in the assessment of trends account for the vast majority of the total abundance of this subpopulation. The most recent year for available abundance data across all rookeries 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).
To apply Criterion A, three generations (or a minimum of ten years, whichever is longer) of abundance data are required (IUCN 2011). For A2, data from three generations ago (~100 yr) are necessary to estimate population declines beginning three generations ago through the present (i.e. assessment) year. The challenges of this requirement on long-lived species like marine 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. The assessors considered extrapolating backward using population trends based on current datasets 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 than to assume that the population has always been in a decline (or increase) of the same magnitude as in the current generation (Table 1 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 this subpopulation, we assumed that the abundance at the beginning of an available time series dataset had not changed significantly in three generations, and therefore used the same abundance value in trend calculations.
Across rookeries, trends were significantly or slightly positive or stable (Table 1 in attached PDF). According to Dutton et al. (2013), Leatherbacks nesting in French Guiana, Suriname, Guyana, and to a lesser extent, Trinidad, are genetically similar, which suggests exchange of females, males, and new adult recruits to beaches other than their natal beaches. The highest abundance rookery—French Guiana—has shown high inter-annual fluctuations in abundance, but has remained stable over the past 40 years (Girondot et al. 2007). Nesting in Guyana and Suriname has increased since the 1990s, and Grand Riviere, Trinidad, hosts possibly the highest reported Leatherback nesting density in the world (TEWG 2007). Leatherback rookeries in Florida (U.S.A.) and St. Croix, U.S. Virgin Islands have also increased significantly since the 1970s (Dutton et al. 2005, Stewart et al. 2011).
Not all rookery trends were positive; the mainland Costa Rica rookery trends showed decreasing abundance since the 1990s though 2010 (Table 1 in attached PDF), as reported previously (Troëng et al. 2004). However, recent reports of high numbers of nesting Leatherbacks at rookeries in Panama and Colombia (Table 1 in attached PDF), and documented beach exchange by turtles among these rookeries (Ordoñez et al. 2007, Patiño-Martínez et al. 2008), indicate that the reported declines at Costa Rican rookeries might simply demonstrate that turtles nested outside of these study areas, rather than actual declines in abundance. Dutton et al. (2013) suggested that Leatherbacks nesting from Costa Rica to the Guiana Shield might represent a genetic cline, with a fuzzy boundary separating stocks. For these reasons, population abundance and trends among these rookeries should be considered together, not separately, to accurately assess nesting abundance.
We also applied A4 to this subpopulation, 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 impacting rookeries or the subpopulation will occur during that time. Based on available information, this appears to be a reasonable assumption; although threats to Leatherbacks persist globally (see Wallace et al. 2011 and Eckert et al. 2012 for review), conservation efforts in the Northwest Atlantic appear to have contributed to stable or increasing population trends for most rookeries (Dutton et al. 2005, Girondot et al. 2007, Hilterman and Goverse 2007, TEWG 2007, Stewart et al. 2011). Furthermore, in cases where present rookery sizes are unlikely to increase further (e.g. Trinidad, where density-dependent effects will likely regulate the population on the nesting beach), we did not use projected future abundance based on the current rate of population increase (doing so would have produced an impossible estimate of >1 million nests yr-1 on <1 km of beach), but instead assumed that abundance would remain constant. Even with this limitation in future abundance, the Northwest Atlantic Leatherback subpopulation could exceed 176,000 nests yr-1 (or >35,000 females yr-1) by 2040.
Criterion B: Extent of occurrence and area of occupancy both exceeded threatened category thresholds for this subpopulation under Criterion B.
Criterion C: To apply Criterion C, we first calculated the number of mature individuals in the subpopulation, i.e., the total number of adult females and males. First, we divided the current average annual number of nests (n=50,842, Table 1) by the estimated clutch frequency across rookeries (i.e. average number of clutches per female; n=5, Briane et al. 2007, TEWG 2007) to obtain an average annual number of nesting females. Next, we multiplied this value by the average re-migration interval (i.e. years between consecutive nesting seasons; n=2.5 yr; TEWG 2007, Dutton et al. 2005, Rivalan et al. 2005) to obtain a total number of adult females that included nesting as well as non-nesting turtles. Finally, to account for adult males, we assumed that the sex ratio of hatchlings produced on nesting beaches in the North Atlantic (approximately 75% female, or 3:1 female:male ratio) reflected the natural adult sex ratio. This calculation provided an estimated mature adult population of 33,810 individuals, which exceeded all threatened category thresholds under Criterion C.
Criterion D: Given the large number of mature individuals for this subpopulation (see above), no threatened category was triggered under Criterion D.
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 UncertaintyAlthough 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).
|Range Description:||Leatherbacks are distributed circumglobally, with nesting sites on tropical sandy beaches and foraging ranges that extend into temperate and sub-polar latitudes (See Eckert et al. 2012 for review). The Northwest Atlantic Leatherback subpopulation range extends throughout the North Atlantic Ocean, from the equator to beyond 50°N, and from the Gulf of Mexico into the Mediterranean (Wallace et al. 2010) (Figure 1 in attached PDF).|
Native:Albania; Anguilla; Antigua and Barbuda; Aruba; Bahamas; Barbados; Belize; Benin; Bermuda; Bosnia and Herzegovina; Brazil; Canada; Colombia; Costa Rica; Côte d'Ivoire; Croatia; Cuba; Cyprus; Dominica; Dominican Republic; Egypt; France (France (mainland)); French Guiana; Gambia; Ghana; Greece; Grenada; Guadeloupe; Guatemala; Guinea; Guinea-Bissau; Haiti; Honduras; Ireland; Israel; Italy; Jamaica; Lebanon; Liberia; Libya; Martinique; Mauritania; Mexico; Montenegro; Montserrat; Morocco; Netherlands Antilles; Nicaragua; Nigeria; Panama; Portugal; Puerto Rico; Saint Kitts and Nevis; Saint Lucia; Saint Vincent and the Grenadines; Senegal; Sierra Leone; Slovenia; Spain; Suriname; Syrian Arab Republic; Togo; Trinidad and Tobago; Tunisia; Turkey; Turks and Caicos Islands; United Kingdom; United States; Venezuela, Bolivarian Republic of; Virgin Islands, British; Virgin Islands, U.S.
|FAO Marine Fishing Areas:||
Atlantic – eastern central; Atlantic – northeast; Atlantic – northwest; Atlantic – southwest; Atlantic – western central; Mediterranean and Black Sea
|Range Map:||Click here to open the map viewer and explore range.|
Leatherbacks are a single species globally comprising 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 (RMU shapefiles can be viewed and downloaded at: http://seamap.env.duke.edu/swot). RMUs are functionally equivalent to IUCN subpopulations, thus providing the appropriate demographic unit for Red List assessments. There are seven Leatherback subpopulations: Northwest Atlantic Ocean, Southeast Atlantic Ocean, Southwest Atlantic Ocean, Northeast Indian Ocean, Southwest Indian Ocean, East Pacific Ocean, and West Pacific Ocean. Multiple genetic stocks have been defined within the Northwest Atlantic subpopulation—Florida (U.S.A.), the northern Caribbean (St. Croix, US Virgin Islands; British Virgin Islands, Puerto Rico), Costa Rica (and likely including Panama and Colombia), the Guianas (Guyana, Suriname, French Guiana), and Trinidad (Dutton et al. 2013)—that overlap significantly in migratory and feeding areas throughout the North Atlantic (TEWG 2007, Wallace et al. 2010, Eckert et al. 2012).
|Habitat and Ecology:||See the species-level account for details. For a thorough review of Leatherback biology, please see Eckert et al. (2012).|
Threats to Leatherbacks—and other marine turtle species—vary in time and space, and in relative impact to populations. Threat categories 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: 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 and for the Northwest Atlantic subpopulation (Wallace et al. 2011, 2013), 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 in three subpopulations and climate change was only scored in 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.
Although threats to Leatherbacks persist globally (see Wallace et al. 2011 and Eckert et al. 2012 for review), conservation efforts in the Northwest Atlantic appear to have contributed to stable or increasing population trends for most rookeries (Dutton et al. 2005, Girondot et al. 2007, Hilterman and Goverse 2007, TEWG 2007, Stewart et al. 2011). However, continued threats from fisheries bycatch in small- and large-scale fishing operations (Wallace et al. 2011, 2013), particularly those near nesting beaches (e.g. Lee Lum 2006) and in distant foraging areas (e.g. James et al. 2006, Stewart et al. 2013), could jeopardize the future state of this subpopulation’s abundance and trend. Thus, continued, effective efforts to mitigate bycatch impacts are absolutely necessary to ensure future population stability or increases for Northwest Atlantic Leatherbacks.
Leatherbacks are protected under various 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); 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 (e.g. Dutton et al. 2005; Chacón-Chaverri and Eckert 2007, Santidrián Tomillo et al. 2007, Sarti Martínez et al. 2007). 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 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).
Fortunately, conservation efforts in the Northwest Atlantic appear to have contributed to stable or increasing population trends for most rookeries (Dutton et al. 2005, Girondot et al. 2007, Hilterman and Goverse 2007, TEWG 2007, Stewart et al. 2011). However, continued threats from fisheries bycatch in small- and large-scale fishing operations and egg harvest for human consumption (Revuelta et al. 2012) could jeopardize the future state of this subpopulation’s abundance and trend (James et al. 2005; Lee Lum 2006; Wallace et al. 2011, 2013). To ensure successful Leatherback conservation, 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 sustain current population trends for the Northwest Atlantic Leatherback subpopulation. 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:||Tiwari, M., Wallace, B.P. & Girondot, M. 2013. Dermochelys coriacea (Northwest Atlantic Ocean subpopulation). In: IUCN 2013. IUCN Red List of Threatened Species. Version 2013.2. <www.iucnredlist.org>. Downloaded on 12 March 2014.|
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