[Federal Register Volume 85, Number 154 (Monday, August 10, 2020)]
[Rules and Regulations]
[Pages 48332-48421]
From the Federal Register Online via the Government Publishing Office [www.gpo.gov]
[FR Doc No: 2020-16277]



[[Page 48331]]

Vol. 85

Monday,

No. 154

August 10, 2020

Part II





Department of the Interior





-----------------------------------------------------------------------





Fish and Wildlife Service





-----------------------------------------------------------------------





50 Part 17





Department of Commerce





-----------------------------------------------------------------------





National Oceanic and Atmospheric Administration





-----------------------------------------------------------------------

50 CFR Parts 223 and 224





Endangered and Threatened Wildlife; 12-Month Finding on a Petition To 
Identify the Northwest Atlantic Leatherback Turtle as a Distinct 
Population Segment and List It as Threatened Under the Endangered 
Species Act; Final Rule

  Federal Register / Vol. 85, No. 154 / Monday, August 10, 2020 / Rules 
and Regulations  

[[Page 48332]]


=======================================================================
-----------------------------------------------------------------------

DEPARTMENT OF INTERIOR

Fish and Wildlife Service

50 Part 17

-----------------------------------------------------------------------

DEPARTMENT OF COMMERCE

National Oceanic and Atmospheric Administration

50 CFR Parts 223 and 224

[Docket No. 200717-0190]
RIN 0648-XF748


Endangered and Threatened Wildlife; 12-Month Finding on a 
Petition To Identify the Northwest Atlantic Leatherback Turtle as a 
Distinct Population Segment and List It as Threatened Under the 
Endangered Species Act

AGENCY: National Marine Fisheries Service (NMFS), National Oceanic and 
Atmospheric Administration (NOAA), Commerce; U.S. Fish and Wildlife 
Service (USFWS), Interior.

ACTION: Notification of 12-month petition finding.

-----------------------------------------------------------------------

SUMMARY: We, NMFS and USFWS, announce a 12-month finding on a petition 
to identify the Northwest Atlantic population of the leatherback turtle 
(Dermochelys coriacea) as a distinct population segment (DPS) and list 
it as threatened under the Endangered Species Act (ESA). In response to 
the petition, we completed a comprehensive status review of the 
species, which also constitutes the 5-year review of the species, to 
determine potential DPSs following the Policy Regarding the Recognition 
of Distinct Vertebrate Population Segments Under the ESA and to perform 
extinction risk analyses. Based on the best scientific and commercial 
data available, including the Status Review Report, and after taking 
into account efforts made to protect the species, we conclude that 
seven populations would meet the discreteness and significance criteria 
for recognition as DPSs, including the Northwest Atlantic population. 
However, even if we were to list them separately, all seven DPSs would 
meet the definition for endangered species (i.e., they are in danger of 
extinction throughout all or a significant portion of their range). The 
species is already listed as endangered throughout its range. We have 
determined that the listing of DPSs is not warranted, and therefore we 
do not propose any changes to the existing global listing.

DATES: This finding was made on August 10, 2020.

ADDRESSES: The Status Review Report are available on NMFS' website at 
https://www.fisheries.noaa.gov/species/leatherback-turtle.

FOR FURTHER INFORMATION CONTACT: Jennifer Schultz, NMFS Office of 
Protected Resources, (301) 427-8443, [email protected]. Persons 
who use a Telecommunications Device for the Deaf (TDD) may call the 
Federal Information Relay Service (FIRS) at 1-800-877-8339, 24 hours a 
day and 7 days a week.

SUPPLEMENTARY INFORMATION:

Background

    The leatherback turtle species as a whole was listed as an 
endangered species (one determined to be threatened with worldwide 
extinction) (35 FR 8491; June 2, 1970), under the Endangered Species 
Conservation Act of 1969, the precursor statute to the ESA (16 U.S.C. 
1531 et seq.). When the ESA was enacted in 1973, it specifically 
provided for continuity with the lists previously in effect under the 
Endangered Species Conservation Act. Section 4(c)(3) of the ESA 
directed that species on the lists of endangered foreign or native 
wildlife at the time the ESA took effect would be deemed ``endangered 
species'' under the ESA without interruption. See 39 FR 1444 (January 
9, 1974) (explaining transition provisions); 39 FR 1158, 1172 (January 
4, 1974) (setting out the final list of ``endangered foreign 
wildlife,'' including ``Turtle, Leatherback'' at 50 CFR 17.11).
    On September 20, 2017, the Blue Water Fishermen's Association 
petitioned NMFS and USFWS (together, the Services) to identify the 
Northwest (NW) Atlantic leatherback turtle population as a DPS and to 
list it as threatened under the ESA. On December 6, 2017, NMFS 
published a ``positive'' 90-day finding in the Federal Register (82 FR 
57565) announcing the determination that the petition presented 
substantial information indicating that the petitioned action may be 
warranted. At that time, NMFS also solicited information on leatherback 
turtles and announced that it would commence, jointly with USFWS, a 
status review of the entire listed species, pursuant to ESA section 
4(b)(3)(A) and 50 CFR 424.14. The resulting Status Review Report 
includes all information used to evaluate the petitioned actions and 
explains the process followed by the Status Review Team (i.e., the 
Team). The following summarizes that information; for additional 
details, please see the Status Review Report (see ADDRESSES).

ESA Statutory, Regulatory, and Policy Provisions and Evaluation 
Framework

    Under the ESA, the term ``species'' includes any subspecies of fish 
or wildlife or plants, and any DPS of any vertebrate fish or wildlife 
which interbreeds when mature (16 U.S.C. 1532(16)). The Services 
adopted a joint policy clarifying their interpretation of the phrase 
``distinct population segment'' for the purposes of listing, delisting, 
and reclassifying a species under the ESA (``Policy Regarding the 
Recognition of Distinct Vertebrate Population Segments Under the 
Endangered Species Act,'' 61 FR 4722 (Feb. 7, 1996; ``DPS Policy''). 
The DPS Policy stipulates two elements that must be considered: (1) 
Discreteness of the population segment in relation to the remainder of 
the species to which it belongs; and (2) the significance of the 
population segment to the species to which it belongs.
    Section 3 of the ESA defines an endangered species as any species 
which is in danger of extinction throughout all or a significant 
portion of its range and a threatened species as one which is likely to 
become an endangered species within the foreseeable future throughout 
all or a significant portion of its range (16 U.S.C. 1532(6) and (20)). 
Thus, we interpret an ``endangered species'' to be one that is 
presently in danger of extinction. A ``threatened species,'' on the 
other hand, is not presently in danger of extinction, but is likely to 
become so within the foreseeable future (that is, within a specified 
later time). In other words, the primary statutory difference between a 
threatened and endangered species is the timing of when a species may 
be in danger of extinction, either presently (endangered) or within the 
foreseeable future (threatened). The ESA uses the term ``foreseeable 
future'' to refer to the time over which identified threats are likely 
to impact the biological status of the species. The duration of the 
``foreseeable future'' in any circumstance is inherently fact-specific 
and depends on the particular kinds of threats, the life-history 
characteristics, and the specific habitat requirements for the species 
under consideration. The existence of threats to a species and the 
species' response to such threats are not, in general, equally 
predictable or foreseeable. Hence, in some cases, the ability to 
foresee a threat to a species is greater than the ability to foresee 
the species' exact response, or the timeframe of such a response, to 
that

[[Page 48333]]

threat. For purposes of making this 12-month finding, the relevant 
consideration is whether the species' population response (i.e., 
abundance, productivity, spatial distribution, diversity) is 
foreseeable, not merely whether the emergence of a threat is 
foreseeable. The foreseeable future extends only as far as we are able 
to reliably predict the species' population response to threats.
    Pursuant to the ESA and our implementing regulations, we determine 
whether a species is threatened or endangered based on any one or a 
combination of the following ESA section 4(a)(1) factors or threats (16 
U.S.C. 1533(a)(1), 50 CFR 424.11(c)):
    1. The present or threatened destruction, modification, or 
curtailment of its habitat or range;
    2. Overutilization for commercial, recreational, scientific, or 
educational purposes;
    3. Disease or predation;
    4. Inadequacy of existing regulatory mechanisms; or
    5. Other natural or manmade factors affecting its continued 
existence, which could include but are not limited to: Fisheries 
bycatch; vessel strikes; pollution (including marine debris and 
plastics, contaminants, oil and gas activities, and derelict fishing 
gear); natural disasters; climate change; and oceanographic regime 
shifts.
    Section 4(b)(1)(A) of the ESA requires us to make listing 
determinations based solely on the best scientific and commercial data 
available after conducting a review of the status of the species and 
after taking into account efforts being made by any State or foreign 
nation or political subdivision thereof to protect the species' 
existence (16 U.S.C. 1533(b)(1)(A)).

Approach to the Status Review

    The Services convened a team of NMFS and USFWS biologists (i.e., 
the Team) to gather and review the best available scientific and 
commercial data on the leatherback turtle, assess the discreteness and 
significance of populations by applying the DPS Policy, evaluate the 
extinction risk of any population segments that meet the DPS criteria, 
and document all findings in a report (i.e., the Status Review Report). 
Although the petitioner requested evaluation only of the NW Atlantic 
leatherback population, we instructed the Team to perform a 
comprehensive status review to identify and evaluate the status of all 
potential DPSs.
    The Team compiled information on leatherback turtle life history, 
biology, ecology, demographic factors, and threats. This included the 
information received in the petition and in response to the Federal 
Register request associated with the 90-day finding (82 FR 57565; 
December 6, 2017). The Team also requested leatherback nesting data 
from beach monitoring programs. To evaluate recent abundance and 
trends, unpublished nesting beach monitoring datasets were often the 
best available data (i.e., most recent and relevant). The Team assessed 
these data in terms of standardization (i.e., the use of standardized 
methodology), consistency (i.e., consecutive seasonal data collection), 
and duration of data collection (i.e., the number of years that data 
were collected). When evaluating threats, peer-reviewed information, 
specifically primary research with large sample sizes and long-term 
sampling duration, was often the best available data. In some 
locations, reports from governments or non-governmental organizations 
and expert opinion constituted the best available information. The Team 
also addressed the source and magnitude of any uncertainty and the 
impact on its conclusions.
    The Team evaluated the discreteness and significance of each 
population and provided their evaluation of whether each population 
would meet the criteria of the DPS Policy. The DPS Policy states that a 
population of a vertebrate species may be considered discrete if it 
satisfies one of the following conditions: (1) It is markedly separated 
from other populations of the same taxon as a consequence of physical, 
physiological, ecological, or behavioral factors (quantitative measures 
of genetic or morphological discontinuity may provide evidence of this 
separation); or (2) it is delimited by international governmental 
boundaries within which differences in control of exploitation, 
management of habitat, conservation status, or regulatory mechanisms 
exist that are significant in light of section 4(a)(1)(D) of the ESA 
(61 FR 4722, February 7, 1996). While the Team used the term ``DPS'' in 
describing and discussing populations that they concluded meet the 
requirements of discreteness and significance, it is important to note 
that the DPS term is used throughout the Status Review Report for ease 
of reference only. A DPS is formally recognized under the ESA only upon 
a listing action by the Services, and the Services cannot delegate 
authority to take formal listing actions to status review teams. The 
information compiled by the Team must be reviewed by the Services, 
which retain responsibility for making the listing determination after 
complying with all the requirements of Section 4 of the ESA and 
considering agency policies. Because we ultimately conclude for the 
reasons discussed in this finding that it would not be appropriate to 
disaggregate the existing global listing into DPSs, references in the 
Status Review Report (and in this finding when we are reviewing the 
information presented by the Team) must be understood as references to 
potential or hypothetical DPSs only.
    The Team evaluated significance in terms of the importance of the 
population segment to the overall welfare of the species, such as: (1) 
Persistence of the population segment in an unusual or unique 
ecological setting; (2) evidence that loss of the population segment 
would result in a significant gap in the range of the taxon; (3) 
evidence that the DPS represents the only surviving natural occurrence 
of a taxon that may be more abundant elsewhere as an introduced 
population outside its historic range; or (4) evidence that the 
population segment differs markedly from other populations of the 
species in its genetic characteristics.
    For each population segment that the Team determined would meet the 
criteria of the DPS Policy (which the Team and we refer to as a ``DPS'' 
for ease of reference), the Team performed an extinction risk analysis, 
which involved the evaluation of demographic factors and threats. 
Demographic factors reflect the impact that operative threats have had 
on the species. In some cases those threats or the impacts from the 
threats are continuing in nature. The demographic factors included 
abundance, productivity, spatial distribution, and diversity. Because 
sea turtles spend the majority of their lives at sea, where they are 
spread across vast distances, it is difficult to estimate total 
abundance. However, the number of nesting females can be counted 
directly, or estimated indirectly by counting the number of nests on 
beaches, during a nesting season. Females nest more than once in a 
season (i.e., clutch frequency, which is the average number of nests 
per season) and do not nest every season (i.e., remigration interval, 
which is the average number of years between successive nesting 
seasons). To calculate the index of nesting female abundance at a 
nesting beach, the Team summed the total number of nests over the most 
recent remigration interval (i.e., a run-sum) and divided this number 
by the clutch frequency. The Team performed these calculations only if 
available data were recent (i.e., last year of the remigration interval 
occurred in 2014 or more recently), consistent

[[Page 48334]]

(i.e., seasonal data collected for each year of the remigration 
interval), and collected in a standardized manner (i.e., data 
collection methods remained the same over the remigration interval), as 
further detailed in the Status Review Report. To provide a total index 
of nesting female abundance for each DPS, we summed the indices of 
nesting female abundance for all monitored beaches used by that DPS. 
The total index of nesting female abundance for each DPS is an index 
(rather than a census) because not all nesting beaches met these 
criteria. However, the nesting beaches that were not included were 
generally unmonitored or not recently monitored because they host few 
nesting females. Even where data were not sufficient to allow for a 
calculation of the index of nesting female abundance, the Team provided 
all available data to ensure the analysis would be as robust as 
possible.
    The Team evaluated the productivity for each DPS by evaluating 
nesting trends (through trend analyses or bar graphs) and productivity 
metrics. Where available data allowed it, they estimated the long-term 
trend for individual beaches using a Bayesian state-space model of 
stochastic exponential population growth (Boyd et al. 2017), where the 
rate parameter describes the annual percent change in observed nest 
counts (or female counts where applicable) over the period of data 
collection. This is further explained in the Status Review Report. To 
reflect current trends over approximately three remigration intervals, 
the criteria for trend analyses were as follows: Nesting data (i.e., 
nest or nesting female counts) consistently collected over nine or more 
years in a standardized manner (for that site), with the most recent 
data collection in 2014 or later and with a minimum average number of 
nests of 50 annually. The Team reported the median trend, along with 
the standard deviation (sd), 95 percent credible interval (CI), and an 
``f statistic'' which is the proportion of the posterior distribution 
with the same sign as the median (i.e., the confidence that the trend 
is positive or negative). When the data did not meet the criteria for 
performing trend analyses, the Team provided bar graphs and/or 
historical data in the Status Review Report. Based on the trend 
analysis (where possible) and the best available historical data, the 
Team characterized the nesting trend for each DPS as decreasing, 
stable, or increasing. The Team also evaluated the following 
productivity metrics (if available): Average size of nesting female; 
nesting female survivorship; remigration interval; clutch size; clutch 
frequency; internesting interval; incubation period; hatching success 
(the proportion of eggs in a nest that produce live hatchlings); and 
sex ratio. Each of these metrics contributes to the growth rate, or 
reproductive potential, of the population.
    For each DPS, the Team evaluated spatial distribution, which 
included the number and location of nesting beaches and foraging areas, 
as well as spatial structure (i.e., whether the DPS exists as a single 
population or several subpopulations connected by metapopulation 
dynamics). The Team also evaluated diversity, which like spatial 
distribution, is a measure of resilience. In general, diverse 
populations with broad spatial distributions and metapopulation 
dynamics are more resilient to threats and environmental changes than 
less diverse populations with narrow distributions.
    For each DPS, the Team next evaluated each of the ESA Section 
4(a)(1) factors (or ``threats'') as listed above (16 U.S.C. 1533(a)(1), 
50 CFR 424.11(c)). For each threat, the Team used the best available 
information to describe the threat, identify which life stages are 
affected, and describe the impact to the DPS with as much specificity 
as the best available information allowed to link the threat to the 
demographic factor it affected. The best available data often allow 
only for qualitative assessment. For each DPS, the Team identified the 
primary threat(s) to its continued existence, as well as other threats. 
The Team considered the impact of each threat individually, with the 
primary threat(s) given the greatest weight, and all threats 
cumulatively, to determine the extinction risk. To assess confidence in 
the extinction risk determination, the Team identified any sources of 
uncertainty and the impact of uncertainty on the conclusions. They 
analyzed all threats assuming the DPS had lost ESA protections going 
forward because a DPS would not receive such protections if it was not 
listed under the ESA. For example, a DPS would not have benefits of 
section 9 take prohibitions or section 7 consultations on actions that 
may affect the DPS.
    The Team performed an extinction risk assessment for each of the 
seven DPSs by evaluating the demographic factors and threats, as 
described above. Then, the Team voted, based on the best available 
data, on whether the extinction risk of each DPS was high, moderate, or 
low, following the definitions included in NMFS' internal guidance 
document, ``Guidance on Responding to Petitions and Conducting Status 
Reviews under the Endangered Species Act, Section II'' (i.e., NMFS' 
Guidance; November 9, 2017) and in the Status Review Report.
    After the Team completed its draft Status Review Report, the 
Services met to review and discuss that document and conservation 
efforts. The Services based our status determinations of the DPSs on 
the best scientific and commercial data available (as compiled and 
reflected in the Status Review Report) and after taking into account 
efforts by States and foreign nation, or any political subdivision 
thereof, to protect the species as mandated by the statute.

DPS Analysis

    The following is a summary of the DPS analysis conducted by the 
Team. For a detailed description of the Team's analyses of discreteness 
and significance, please see the Status Review Report. As a starting 
point, the Team considered seven leatherback populations that were 
previously identified as regional management units (RMUs) by Wallace et 
al. (2010) and recognized as subpopulations under the International 
Union for Conservation of Nature (IUCN) Red List (https://www.iucnredlist.org/species/6494/43526147). The Team found that seven 
leatherback populations met the discreteness and significance criteria 
per the DPS Policy and identified the following potential DPSs: 
Northwest (NW) Atlantic; Southwest (SW) Atlantic; Southeast (SE) 
Atlantic; SW Indian; Northeast (NE) Indian; West Pacific; and East 
Pacific.

Discreteness

    The Team evaluated all populations for discreteness and determined 
that each showed marked separation from the others as a consequence of 
behavioral and physical factors. Behavioral factors, especially 
returning to waters off a turtle's natal beach to breed, have prevented 
interbreeding, resulting in reproductive isolation, as indicated by 
genetic discontinuity.
    Although some populations use the same foraging areas, tagging and 
telemetry studies also demonstrate the discreteness of the populations 
at nesting beaches. Physical factors, such as land masses, ocean 
currents, and other oceanographic features, have established and 
reinforced barriers to gene flow among the seven populations.
    Genetic data provide the most compelling evidence for discreteness 
among the seven populations. The most recent and comprehensive global 
analysis of published and unpublished

[[Page 48335]]

mitochondrial deoxynucleic acid (mtDNA) sequence data (i.e., 28 
haplotypes, which are unique sequences of mtDNA) evaluated samples 
collected from 21 nesting sites representing key regions from all ocean 
basins (Dutton et al. 2007; Dutton et al. 2013; Shanker et al. 2011; 
Dutton and Shanker 2015); analyzing the evolutionary relationship of 
these data revealed three distinct haplogroups (i.e., similar 
haplotypes that cluster together, relative to other haplotypes) that 
are geographically segregated across the Atlantic, Indian, and Pacific 
Oceans (Dutton, unpublished data; NMFS and USFWS 2020). Early mtDNA 
analyses indicated strong genetic discontinuity, globally 
(FST = 0.415, P <0.001) and within ocean basins 
(FST = 0.203 to 0.253, P <0.001; Dutton et al. 1999). 
Wallace et al. (2010) combined these and other genetic data with 
nesting, flipper tagging, and satellite telemetry data to identify 
seven leatherback RMUs, which provided the starting point for our 
identification of discrete populations.
    From this starting point, the Team then evaluated more recent 
genetic data. Subsequent genetic analyses confirmed genetic 
discontinuity among the NW, SW, and SE Atlantic populations (Wallace et 
al. 2010; Dutton et al. 2013; Carreras et al. 2013; Molfetti et al. 
2013; Vargas et al. 2017). Elevated genetic differentiation at nuclear 
DNA (FST = 0.211-0.86) indicates that males, like females, 
likely return to the waters off their natal beaches to mate and that 
male-mediated gene flow may not be as pronounced as previously thought 
(Dutton et al. 2013; see Jensen et al. 2013). Nuclear (FST 
>0.126, P <0.001; Dutton et al. 2013) and mtDNA (FST >0.061, 
P = 0.05-0.001; Dutton et al. 2013; FST >0.061, P <0.01; 
Vargas et al. 2017) analyses indicate genetic discontinuity between the 
Atlantic populations and the SW Indian population. Preliminary mtDNA 
results for leatherback turtles nesting at Little Andaman Island, India 
(Shanker et al. 2011; Dutton and Shanker 2015), indicate that this 
population is closely related to the extinct Malaysian population, with 
which it shares common haplotypes. It is markedly different from the 
South African nesting population, as well as those in the West Pacific 
population (Dutton et al. 2007, 2013 and unpublished). Samples from 
extant and extirpated nesting aggregations of the NE Indian population 
(Shanker et al. 2011; Dutton and Shanker 2015; Dutton et al. 
unpublished data) are genetically differentiated from the SW Indian 
population (FST = 0.415, P <0.003; Dutton et al. 1999) and 
the West Pacific population (X2 = 49.346, P = 0.002; Dutton 
et al. 2007). There is genetic discontinuity between the West and East 
Pacific populations, as demonstrated by significant genetic 
differentiation between the samples from Solomon Islands in the western 
Pacific and Mexico or Costa Rica in the eastern Pacific (FST 
= 0.270 and 0.331, P <0.001; Dutton et al. 1999). Genetic discontinuity 
among all seven populations provides evidence for marked separation 
from the others and thus discreteness of each population.
    Tagging and telemetry studies confirm marked separation of the 
seven populations because nesting sites remain distant and isolated. 
Nesting females of one population have not been tracked to, or observed 
on, beaches used by another population, even though telemetry data 
indicate shared use of foraging areas by different populations.
    Telemetry studies demonstrate that females nesting on NW Atlantic 
beaches move throughout most of the North Atlantic from the Equator to 
about 50[deg] N latitude (Ferraroli et al. 2004; Hays et al. 2004; 
James et al. 2005a; James et al. 2005b; 2005c; Eckert 2006a; Eckert et 
al. 2006b; Hays et al. 2006; Doyle et al. 2008; Evans 2008; Dodge et 
al. 2014; Fossette et al. 2014; Aleksa 2017; Aleksa et al. 2018). 
Turtles originating from beaches of the NW Atlantic appear to mix at 
foraging areas throughout the North Atlantic Ocean (Fossette et al. 
2014), but their movements rarely extend into waters south of the 
Equator. Tagging studies further support the connectivity within and 
among nesting beaches and foraging areas of the North Atlantic Ocean 
(Tro[euml]ng et al. 2004; Br[auml]utigam and Eckert 2006; 
Chac[oacute]n-Chaverri and Eckert 2007; Turtle Expert Working Group 
(TEWG) 2007; S[ouml]nmez et al. 2008; Dutton et al. 2013b; Horrocks et 
al. 2016), but turtles tagged in the North Atlantic Ocean have never 
been found on nesting beaches in Brazil (SW Atlantic population) or 
Africa (SE Atlantic population). In the South Atlantic Ocean, post-
nesting females tracked from nesting beaches in Gabon and Brazil use 
the same foraging areas, including waters off SW Africa, in the south 
equatorial Atlantic and off SE Brazil and Uruguay (Almeida et al. 2011; 
Witt et al. 2011). Turtles incidentally captured in fisheries off South 
America (Billes et al. 2006, L[oacute]pez-Mendilaharsu et al. 2009) 
also demonstrate that turtles originating from the SW and SE Atlantic 
Ocean beaches share foraging areas. Despite such mixing at foraging 
areas, there is no evidence for the shared use of nesting beaches. 
Genetic data indicate that turtles return to their natal beaches to 
nest on opposite sides of the Atlantic Ocean (Dutton et al. 2013; 
Vargas et al. 2017), and no tag recoveries contradict these data.
    In the Indian Ocean, telemetry studies have been conducted at South 
African nesting beaches in the SW Indian Ocean (Hughes et al. 1998; 
Luschi et al. 2006; Robinson et al. 2016) and at Andaman Islands 
nesting beaches in the NE Indian Ocean (Namboothri et al. 2012; 
Swaminathan et al. 2019). South African nesting females showed diverse 
movements that were highly influenced by complex oceanographic currents 
and features that lead them to foraging destinations in the South 
Atlantic Ocean, SW Indian Ocean, and Mozambique Channel (Hughes et al. 
1998, Luschi et al. 2006, Robinson et al. 2016). About half of the 10 
post-nesting females tagged at the Andaman Islands moved westward: Two 
individuals reached the Mozambique Channel; the other half moved 
southeastward, past the Indonesian islands of Sumatra and Java, with 
one leatherback reaching an apparent foraging ground off NW Australia 
before transmissions stopped (Namboothri et al. 2012; Swaminathan et 
al. 2019). Despite overlap in one foraging area (i.e., reaching the 
Mozambique Channel), tagging data do not indicate movement between the 
distant nesting beaches.
    Within the Pacific Ocean, nearly all turtles tracked from East 
Pacific nesting beaches moved southward across the Equator to forage in 
open-ocean waters of the SE Pacific Ocean or in the coastal waters of 
Central America, Peru, and Chile. The movements of post-nesting females 
from the West Pacific Ocean are dependent on the season in which they 
nest, with winter-nesting females predominantly tracked into the 
Southern Hemisphere and summer-nesting females foraging in diverse 
coastal and oceanic ecosystems throughout the northern Indo-Pacific 
region (Benson et al. 2011). Telemetry data indicate little or no 
overlap with foraging destinations utilized by nesting females of the 
East and West Pacific populations (Bailey et al. 2012; Benson et al. 
2011). However, a genetic study of bycaught turtles off the coast of 
Chile and Peru indicated that 15 percent of leatherback turtles 
originated from West Pacific nesting beaches (Donoso and Dutton 2010), 
suggesting that foraging overlap may be more prevalent than estimated 
by telemetry data. Still, there is no genetic evidence for contemporary 
interbreeding between the two populations (Dutton et al. 2007), and 
telemetry and tagging data do not indicate movement between the distant

[[Page 48336]]

nesting beaches. Thus, flipper tagging and satellite telemetry data 
support the marked separation, and thus discreteness, of the seven 
populations at their nesting beaches.
    Physical factors likely shape and reinforce the behavior patterns 
that result in reproductive isolation. Though the species has a global 
range, with foraging areas extending into high latitudes, nesting 
mainly occurs on tropical or subtropical beaches. Post-hatchling 
dispersal is determined by the ocean currents they encounter off 
nesting beaches. While adults move throughout tropical and temperate 
waters irrespective of ocean currents, both males and females return to 
the waters off their natal nesting beach to mate. This natal homing is 
somewhat flexible, (Dutton et al. 2013; Jensen et al. 2013), creating 
reproductive isolation only among distant nesting sites, which may also 
be physically separated from one another by land masses and 
oceanographic barriers to gene flow. For example, leatherback turtles 
in the Atlantic Ocean are physically separated from those in the 
Pacific Ocean by the Americas. Though leatherback turtles have greater 
cold tolerance than other sea turtles, they do not appear to venture 
into latitudes greater than 47[deg] S or 71[deg] N (Eggleston 1971; 
Eckert et al. 2012). Therefore, the low latitude and cold waters of the 
Cape Horn Current likely prevent movement between the Atlantic and 
Pacific Oceans. Within ocean basins, nesting beaches of the discrete 
populations are separated by long distances of uninterrupted deep water 
(e.g., the East Pacific Barrier and the mid-Atlantic Barrier). While 
leatherback turtles clearly cross these open-ocean barriers to reach 
distant foraging areas, they do not appear to do so for nesting and 
breeding, but rather return to their natal region to breed and nest 
(Barragan et al. 1998; Dutton et al. 1999; Barragan and Dutton 2000; 
Dutton et al. 2013). Within ocean basins, currents shape post-
hatchlings' movement patterns, which they may retain as adults (e.g., 
Fossette et al. 2010; Benson et al. 2011). The NW Atlantic leatherback 
population appears to be physically separated from the SE and SW 
Atlantic populations by the current systems of the South and North 
Atlantic Gyres, respectively. NW Atlantic leatherback nesting beaches 
are adjacent to northward moving currents (e.g., Gulf Stream). 
Leatherback hatchlings from these nesting beaches, therefore, are 
transported northward, remaining in the North Atlantic Ocean. Those 
that survive return to their nesting beaches as adults, completing 
their life stages within the North Atlantic (Fossette et al. 2010; 
Chambault et al. 2017). The SE and SW Atlantic populations are 
similarly retained in the South Atlantic Ocean by the South Atlantic 
Gyre and the Benguela Current, which flows northward along the SE coast 
of Africa, restricting movement into the Indian Ocean. Within the 
Indian Ocean, the Somali Current runs between the nesting beaches of 
the SW and NE Indian populations. The NE Indian and West Pacific 
populations likely became isolated as a result of exposed land barriers 
between Indonesia, New Guinea, and the Philippines as a result of low 
sea levels within the past 6,000 years (Barber et al. 2000). Seasonal 
monsoons may also play a contemporary role by altering current 
directions and hatchling dispersal patterns (Benson et al. 2011; Gaspar 
et al. 2012). Thus, physical factors have likely helped to shape, or at 
least reinforce, the reproductive isolation among distant nesting 
beaches.
    Based on these data, the Team concluded that the seven populations 
demonstrate discreteness, or marked separation from each other, due to 
behavioral and physical factors. These are the NW Atlantic, SW 
Atlantic, SE Atlantic, SW Indian, NE Indian, West Pacific, and East 
Pacific populations.

Significance

    Each of the discrete populations is significant to the species 
because the loss of any one would result in a significant gap (i.e., a 
half or quarter of an ocean basin) in the range of the species. Several 
populations also persist in unique ecological settings. Each population 
likely possesses unique genetic characteristics and local adaptations 
as a result of thousands of years of reproductive isolation, but none 
have yet been identified because all genetic studies have involved 
neutral markers. Therefore, the Team did not rely on evidence of unique 
genetic characteristics and local adaptations for its significance 
finding.
    A loss of the NW Atlantic population would result in a gap (i.e., 
the entire North Atlantic Ocean) of the nesting and foraging range of 
the species. If the NW Atlantic population were extirpated, it is 
unlikely that leatherback turtles from other populations would 
recolonize the North Atlantic Ocean in an ecological time frame (i.e., 
tens to hundreds of years), leaving a significant gap in the range of 
the species. Extirpation of this population would also significantly 
reduce the genetic diversity of the species, as reflected by the 
possession of several unique haplotypes. Leatherback turtles of the NW 
Atlantic Ocean also occur in a unique ecological setting; this is the 
only DPS that regularly forages at high latitudes. Sightings have been 
documented as far north as Norway and Iceland (Brongersma 1972; Goff 
and Lien 1988; Carriol and Vader 2002; McMahon and Hayes 2006; Eckert 
et al. 2012). Such high latitude foraging is likely facilitated by the 
warm Gulf Stream, which meets cold water currents to create highly 
productive foraging areas. The Team concluded that the NW Atlantic 
population is biologically significant to the species.
    In the SW Atlantic Ocean, leatherback turtles only nest in a small 
area of the coastline of Brazil. All other nesting in South America 
occurs above the Equator or on the Pacific Coast. Therefore, the loss 
of this population would result in a gap of the nesting range of the 
species (i.e., the SW Atlantic coast). Although SE Atlantic leatherback 
turtles forage off the coasts of Brazil, Argentina, and Uruguay, they 
do not breed there. Rather, they return to the waters off western 
Africa to mate (Vargas et al. 2017). Therefore, if the SW Atlantic 
population were extirpated, it is unlikely that leatherback turtles 
from other populations would recolonize this region, leaving a 
significant gap in the nesting range of the species. The extirpation of 
this population would also significantly reduce the genetic diversity 
of the species, as reflected by the possession of unique haplotypes and 
high genetic diversity, despite the small population size (Vargas et 
al. 2017). The SW Atlantic population is biologically significant to 
the species.
    Leatherback turtles of the SE Atlantic population nest in West 
Africa and forage in the South Atlantic Ocean. This population is much 
more abundant than the SW Atlantic population, which also forages in 
the South Atlantic Ocean. Therefore, the loss of this population would 
result in a gap of the nesting range of the species (i.e., western 
Africa) and a significant reduction in the abundance of leatherback 
turtles foraging throughout the South Atlantic Ocean. The extirpation 
of this population would also significantly reduce the genetic 
diversity of the species, as reflected by the possession of unique 
haplotypes. The Team concluded that the SE Atlantic population is 
biologically significant to the species.
    In the SW Indian Ocean, leatherback turtles only nest in a small 
area along the South African and Mozambican coastlines. No other 
leatherback turtles nest in eastern Africa or in other areas throughout 
the entire western Indian Ocean. Therefore, the loss of this population 
would result in a gap of the

[[Page 48337]]

nesting range of the species (i.e., the SW Indian Ocean). The SW Indian 
population also occurs in a unique ecological setting: It is the only 
population to nest on temperate beaches. The warm Agulhas Current, 
adjacent to the nesting beaches, likely facilitates their high-latitude 
nesting. The Team concluded that the SW Indian population is 
biologically significant to the species.
    Leatherback turtles nest in small numbers in the NE Indian Ocean. 
These nesting sites are separated from other Indian Ocean nesting sites 
by at least 5,000 km. Although western Pacific nesting sites are 
closer, males and females return to the waters off their natal beaches 
to breed, preventing interbreeding among NE Indian and West Pacific 
populations. Therefore, the loss of this population would result in a 
gap of the nesting range of the species (i.e., the NE Indian Ocean). 
The extirpation of this population would also significantly reduce the 
genetic diversity of the species, as reflected by the possession of 
unique haplotypes. The Team concluded that the NE Indian population is 
biologically significant to the species.
    West Pacific leatherback turtles nest in small numbers primarily in 
four nations of the West Pacific Ocean. These nesting sites are 
separated from East Pacific nesting sites by over 10,000 km. Though NE 
Indian nesting sites are closer in distance, male and female philopatry 
prevents interbreeding. Therefore, the loss of this population would 
result in a gap of the nesting range of the species (i.e., the West 
Pacific Ocean). The loss of this population would also result in a gap 
of the foraging range of the species (i.e., the North Pacific Ocean). 
The extirpation of this population would also significantly reduce the 
genetic diversity of the species, as reflected by the possession of 
unique haplotypes. The West Pacific population is ecologically unique 
in two ways: It is the only population to forage in both hemispheres; 
and it nests year-round, with nesting peaks in the summer and winter. 
The Team concluded that the West Pacific population is biologically 
significant to the species.
    Leatherback turtles nesting on eastern Pacific coastlines also 
forage in the East Pacific Ocean. A loss of this population would 
result in a gap of the nesting range of the species (i.e., the East 
Pacific Ocean). Though West Pacific leatherback turtles may forage off 
the coasts of Peru and Chile, they do not breed there (Donoso and 
Dutton 2010). Therefore, if the East Pacific population were 
extirpated, it is unlikely that leatherback turtles from other 
populations would recolonize this region, leaving a significant gap in 
the nesting range of the species. The extirpation of this population 
would also significantly reduce the genetic diversity of the species, 
as the population possess several unique haplotypes. The East Pacific 
population is unique in having the smallest nesting female size, clutch 
size, and egg size of all populations, possibly reflecting unique 
foraging conditions that are subject to oceanographic regime shifts 
(e.g., the El Ni[ntilde]o Southern Oscillation, or ENSO). The Team 
concluded that the East Pacific population is biologically significant 
to the species.

DPS Summary

    The Team found that seven populations met the definition for 
discreteness. These populations are markedly separated as a result of 
the behavioral factors of movement (as demonstrated by satellite 
telemetry and flipper tagging studies) and philopatry, which has led to 
reproductive isolation (as demonstrated by genetic discontinuity). They 
are also physically separated by land masses, oceanographic features, 
and currents. The Team found these seven populations to be significant 
to the species because the loss of any one of them would result in a 
significant gap in the range of the species as well as a significant 
loss of genetic diversity, reducing the evolutionary potential of the 
species. Some populations also occur in a unique ecological setting. 
Thus, after reviewing the best available information, the Team 
identified the following populations as potential DPSs: NW Atlantic, SW 
Atlantic, SE Atlantic, SW Indian, NE Indian, West Pacific, and East 
Pacific. The Team defined the potential DPSs as leatherback turtles 
originating from nesting beaches within the boundaries for each DPS. 
The range of each DPS, which also includes foraging areas, thus extends 
beyond the nesting boundaries for most DPSs, and may overlap 
extensively with the range of another DPS. The boundaries are based on 
the best available genetic, telemetry, and observational data. When 
such data were not available, the Team used information on possible 
barriers to gene flow, such as oceanographic features. For ease of use, 
the Team applied political boundaries when this did not conflict with 
biological or oceanographic data. Additional information on the 
boundaries is available in the following sections, which summarize the 
extinction risk analysis for each DPS, and in the Status Review Report.

NW Atlantic DPS

    The Team defined the NW Atlantic DPS as leatherback turtles 
originating from the NW Atlantic Ocean, south of 71[deg] N, east of the 
Americas, and west of Europe and northern Africa; the southern boundary 
is a diagonal line between 5.377[deg] S, 35.321[deg] W and 16.063[deg] 
N, 16.51[deg] W. The northern boundary reflects a straight latitudinal 
line based on the northernmost documented occurrence of leatherback 
turtles (Brongersma 1972; Goff and Lien 1988; Carriol and Vader 2002; 
McMahon and Hayes 2006; Eckert et al. 2012). The southern boundary is a 
diagonal line between the elbow of Brazil, where the Brazilian current 
begins and likely restricts the nesting range of this DPS, and the 
northern boundary of Senegal. The boundary between Senegal and 
Mauritania was chosen because the SE Atlantic DPS does not appear to 
nest above this boundary (Fretey et al. 2007).
    The range of this DPS (i.e., all areas of occurrence) extends 
throughout the North Atlantic Ocean, including the Caribbean Sea, Gulf 
of Mexico (GOM), and Mediterranean Sea. Available data indicate that 
the NW Atlantic DPS occurs (at varying levels of frequency) in the 
waters of the following nations or territories: Albania, Algeria, 
Anguilla, Antigua and Barbuda, Aruba, Azores, Bahamas, Barbados, 
Belize, Bermuda, Bonaire, Bosnia and Herzegovina, Brazil, British 
Virgin Islands, Canada, Cape Verde, Cayman Islands, Colombia, Costa 
Rica, Croatia, Cuba, Cura[ccedil]ao, Cyprus, Denmark, Dominica, 
Dominican Republic, Egypt, France, French Guiana, Greece, Greenland, 
Grenada, Guadeloupe, Guatemala, Guyana, Haiti, Honduras, Iceland, 
Ireland, Israel, Italy, Jamaica, Lebanon, Libya, Madeira, Malta, 
Martinique, Mauritania, Mexico, Montenegro, Montserrat, Morocco, 
Netherlands Antilles, Nicaragua, Norway, Panama, Portugal, Slovenia, 
Spain, St. Barthelemy, St. Eustatius, St. Kitts and Nevis, St. Lucia, 
St. Maarten, St. Pierre and Miquelon, St. Martin, St. Vincent and the 
Grenadines, Suriname, Sweden, Syria, Trinidad and Tobago, Tunisia, 
Turkey, Turks and Caicos Islands, United Kingdom, United States 
(including Puerto Rico and the U.S. Virgin Islands (USVI), Venezuela, 
and Western Sahara.
    All nesting in this DPS occurs in the NW Atlantic Ocean, 
concentrated from the southeast United States throughout the Wider 
Caribbean Region (Dow et al. 2007). Leatherback nesting in the NW 
Atlantic can be grouped into several broad geographical areas, 
including the

[[Page 48338]]

U.S. mainland (primarily Florida), North Caribbean (including USVI and 
Puerto Rico), West Caribbean (Honduras to Colombia), and Southern 
Caribbean/Guianas (Venezuela to French Guiana; TEWG 2007). The largest 
nesting aggregations occur in Trinidad, French Guiana, and Panama. The 
northern-most confirmed nesting occurs in North Carolina, but there has 
been a crawl recorded as far north as Assateague Island National 
Seashore, Maryland (Rabon et al. 2003). No nesting occurs in the 
Mediterranean Sea (Casale and Margaritoulis 2010).
    Nesting occurs on unobstructed, high-energy beaches with either a 
deep water oceanic approach or a shallow water approach with mud banks, 
but without coral or rock formations (TEWG 2007). The main 
characteristics of leatherback nesting beaches include coarse-grained 
sand; steep, sloping littoral zone; obstacle-free approach; proximity 
to deep water; and oceanic currents along the coast (Hendrickson and 
Balasingam 1966 in Eckert et al. 2015). During the nesting season, 
adult females and males inhabit the waters off nesting beaches. During 
a nesting season, females generally stay within about 100 km of their 
nesting beaches, remaining close to the coast on the continental shelf, 
and engaging in shallow dives (Eckert et al. 2012). Intra-seasonal 
movement of greater than 100 km also occurs, especially between French 
Guiana and Suriname (Fossette et al. 2007; Georges et al. 2007), Panama 
and Costa Rica (Chac[oacute]n-Chaverri and Eckert 2007), and among 
Caribbean nesting beaches, including those on Trinidad (Brautigam and 
Eckert 2006; Georges et al. 2007; Horrocks et al. 2016). Adult males 
migrate from temperate foraging areas in the North Atlantic Ocean to 
waters off nesting beaches, typically arriving before the nesting 
season and remaining for the majority of the season (James et al. 
2005b; Doyle et al. 2008; Dodge et al. 2014).
    Foraging areas of the NW Atlantic DPS include coastal and pelagic 
waters of the North Atlantic Ocean (Eckert et al. 2012; Saba 2013; 
Shillinger and Bailey 2015). These waters include the GOM, North 
Central Atlantic Ocean, northwestern Atlantic shelf waters of the 
United States and Canada, waters along the southeastern U.S. coast, the 
Mediterranean Sea, and the northeastern Atlantic shelf waters of Europe 
and northwestern Africa (TEWG 2007). Some post-nesting females also 
remain in tropical waters to forage (Fossette et al. 2010). This DPS is 
mostly commonly associated with open-ocean and coastal shelf foraging 
areas off Nova Scotia (Canada), northeastern United States, GOM, 
northwestern Europe, and northwestern Africa (James et al. 2005a, 
2006b, 2007; Eckert 2006; Eckert et al. 2006; Fossette et al. 2010a; 
Fossette et al. 2010b; Dodge et al. 2014; Stewart et al. 2016; Aleksa 
et al. 2018). Fossette et al. (2014) analyzed available satellite 
telemetry data from 1995 to 2010 on post-nesting females (n = 93) as 
well as males (n = 4), females (n = 8), and a juvenile (n = 1) from 
foraging grounds throughout the Atlantic Ocean. They found widespread 
use of the North Atlantic Ocean (Fossette et al. 2014). High-use areas 
mainly occurred in the central (25 to 50[deg] N, 50 to 30[deg] W) and 
eastern Atlantic Ocean, in particular in the waters offshore Western 
Europe, around Cape Verde (year-round) and the Azores (October to 
March; Fossette et al. 2014). Fossette et al. (2014) found that 
seasonal high-use areas also occurred along the eastern U.S. coast 
(April to June and October to December) and off Canada (July to 
December). The GOM is also a high-use foraging area, with a peak in the 
northeast GOM during August and September (Aleksa et al. 2018). 
Overall, leatherback turtles of the North Atlantic population appear to 
have a diverse array of foraging habitat available.

Abundance

    The total index of nesting female abundance for the NW Atlantic DPS 
is 20,659 females. The nesting beaches with the greatest abundance have 
been included in this index, and most beaches with an unquantified 
number of nests likely host few nesting females. We based this index on 
24 nesting aggregations in 10 nations: Trinidad and Tobago (n = 
11,324), French Guiana (n = 2,519), Panama (n = 2,251), United States 
(n = 1,694), Costa Rica (n = 1,306), Suriname (n = 698), Grenada (n = 
499), Venezuela (n = 215), Guyana (n = 76), and Nicaragua (n = 10). 
With the possible exception of Colombia, our total index does not 
include 31 unquantified but likely small nesting aggregations for which 
data are not available. It also does not include outdated data 
published by Dow et al. (2007), which includes binned crawls, 
categorized as less than 25, 25 to 100, 100 to 500, 500 to 1000, or 
unknown abundance. Crawls or emergences (measured as females or tracks 
on beaches) include both successful egg-laying and unsuccessful 
nesting, so the number of crawls represents approximately two to 10 
times the number of nests (Dow et al. 2007). Because the Dow et al. 
data, which are more than 10 years old and do not provide the number of 
actual nests, may not be representative of recent nesting trends, we 
did not include them in our total index. To calculate the indices of 
nesting female abundance, we added the number of nests over the last 3 
years (representing the most recent remigration interval; Eckert et al. 
2012) and divided by the clutch frequency (site-specific values or, 
when such values were not available, the average of the site-specific 
values, i.e., 5.5 clutches per season).
    Our total index of nesting female abundance is based on the best 
available data for this DPS. It is the most robust estimate of nesting 
females at this time because it only includes available nesting data 
from recently and consistently monitored nesting beaches. Our total 
index does not include data from beaches where we were unable to 
quantify the number of nesting females, either due to the lack of 
recent or available nesting data or because only crawl data were 
reported (often on smaller nesting beaches). Scattered nesting may 
occur on beaches throughout the region, but because these beaches are 
not monitored, or have not been recently monitored, recent data are not 
available.
    Nesting in the NW Atlantic DPS is characterized by many small 
nesting beaches. Large nesting aggregations are rare; only about 10 
leatherback nesting beaches in the Wider Caribbean Region (about two 
percent of the DPS's total nesting sites) host more than 1,000 crawls 
annually (Dow Piniak and Eckert 2011). Only one site, Grande Riviere in 
Trinidad, hosts more than 5,000 nesting females, representing 29 
percent of the total index of nesting female abundance. Relatively 
large nesting aggregations are also found in Matura (Trinidad), 
Chiriqui Beach (Panama), and Cayenne and Remire Montjoly (French 
Guiana). In contrast, most known nesting beaches support a small 
nesting female abundance; 71 percent of the total nesting sites record 
annual crawls of less than 100 (Dow Piniak and Eckert 2011). The number 
of nesting females is unquantified at 31 beaches (i.e., the majority of 
nesting sites for the DPS). However, for the reasons identified above, 
most of those sites have small abundance levels as inferred from the 
numbers of crawls estimated by Dow et al. (2007). Therefore, our total 
index of nesting female abundance represents the most robust estimate 
allowed by the best available data and includes the majority of nesting 
females because the largest nesting aggregations were included. The 
data regarding additional nesting aggregations are not sufficiently 
recent, specific, or reliable for inclusion, and the contribution of 
these nesting

[[Page 48339]]

aggregations to the total index is expected to be small.
    Our total index of nesting female abundance is similar in 
comparison to other published estimates. TEWG (2007) estimated the 
abundance of NW Atlantic leatherback turtles using nesting data from 
2004 and 2005. At that time, the number of adult females (equating to 
total index of nesting female abundance in our analysis) was estimated 
to be approximately 18,700 (range 10,000 to 31,000). While a wide range 
was provided, the point estimate in TEWG (2007) is similar to, albeit 
slightly lower than, our total index of 20,659 nesting females. The 
most recent, published IUCN Red List assessment for the NW Atlantic 
Ocean subpopulation estimated a total of 20,000 mature individuals (The 
NW Atlantic Working Group 2019). Our total index, which only includes 
nesting females, exceeds their estimate, likely due to our use of a 3-
year remigration interval, which has increased at some locations in 
recent years (e.g., 4.5 years at St. Croix; K.R. Stewart, The Ocean 
Foundation and C. Lombard, USFWS, pers. comm., 2019).
    We conclude that the total index of nesting females for the NW 
Atlantic DPS is 20,659 females. The nesting beaches with the greatest 
abundance have been included in our total index, and most beaches with 
an unquantified number of nests likely host few nesting females. 
Current nesting female abundance is not at a level where stochastic or 
environmental changes would have catastrophic impacts, but the 
abundance at several nesting sites with previously high density has 
declined drastically. However, as we discuss below, a declining nest 
trend and several existing threats will likely continue to reduce this 
abundance.

Productivity

    The NW Atlantic DPS exhibits decreasing nest trends at nesting 
aggregations with the greatest indices of nesting female abundance. 
Though some nesting aggregations indicate increasing trends, most of 
the largest ones demonstrate declining nest trends. We evaluated nest 
trends by using nest count data consistently collected using a 
standardized approach for at least 9 years, with the last year of data 
in 2014 or more recently and with an average of more than 50 nests 
annually. When data did not meet these criteria, we evaluated bar 
graphs provided in the Status Review Report to consider all available 
data. Thus, these data are representative of the DPS because they 
include the largest nesting aggregations. With the possible exception 
of Colombia, nesting aggregations for which data are not available are 
likely small. Significant declines have been observed at nesting 
beaches with the greatest historical or current nesting female 
abundance, most notably in Trinidad and Tobago (Grande Riviere, Fishing 
Pond, and Tobago), Suriname, French Guiana (Awala-Yalimapo), Florida, 
and Costa Rica (Tortuguero). Therefore, these nest trends represent the 
best available data for this DPS.
    In Trinidad and Tobago, trends in annual nest counts were largely 
negative between 2009 and 2017, the years for which data were 
available. For Trinidad, we analyzed trends for three separately 
monitored beaches, including Grande Riviere, Matura, and Fishing Pond. 
The long-term trend was negative for Grande Riviere (median = -6.9 
percent; sd = 17.4 percent; 95 percent CI = -43.8 to 26.9 percent; f = 
0.682; mean annual nests = 13,272), positive for Matura (median = 1.8 
percent; sd = 15.1 percent; 95 percent CI = -29.2 to 33.0 percent; f = 
0.561; mean annual nests = 7,359), and negative for Fishing Pond 
(median = -19.3 percent; sd = 15.1 percent; 95 percent CI = -49.8 to 
12.0 percent; f = 0.916; mean annual nests = 3,892). For Tobago, the 
median trend was -0.9 percent annually (sd = 11.3 percent; 95 percent 
CI = -25.0 to 21.5 percent; f = 0.540; mean annual nests = 452).
    For French Guiana, we analyzed nest count data from 2002 to 2017 
for Awala-Yalimapo beach in the west and data from 1999 to 2017 for 
Cayenne and Remire Montjoly beaches in the east. There was a steep 
decline at Awala-Yalimapo, with a median trend of -19.4 percent 
annually (sd = 12.2 percent; 95 percent CI = -43.2 to 6.0 percent; f = 
0.942; mean annual nests = 3,200). In contrast to Awala-Yalimapo, nest 
counts at Cayenne and Remire Montjoly increased by 2.8 percent annually 
(sd = 12.9 percent; 95 percent CI = -24.9 to 27.9 percent; f = 0.596; 
mean annual nests = 3,498). In addition, leatherback nesting occurred 
on remote beaches in western French Guiana until 2013 (e.g., a high of 
4670 nests was found in 2003, with 1,270 mean annual nests from 2002 to 
2013), but we were unable to analyze trends because monitoring on these 
remote beaches has been reduced since approximately 2010 due to 
significant beach erosion and the disappearance of some previously 
monitored beaches.
    Suriname, Grenada, and Panama each had a single time series 
sufficient for trend analysis. For Suriname, we combined datasets from 
two beaches, Galibi and Braamspunt, which were monitored between 2001 
and 2017. Total nests in Suriname declined by -14.6 percent annually 
(sd = 9.6 percent; 95 percent CI = -36.4 to 4.5 percent; f = 0.953; 
mean annual nests = 4,586). In Grenada, data on the number of nesting 
tracks were collected on Levera beach between 2002 and 2018. There was 
a 7.1 percent annual increase in tracks at Levera during that period 
(sd = 8.7 percent; 95 percent CI = -10.5 to 25.3 percent; f = 0.827; 
mean annual tracks = 895). In Panama, the nest counts at Chiriqui beach 
increased by 0.8 percent annually (sd = 7.0 percent; 95 percent CI = -
14.1 to 14.6 percent; f = 0.557; mean annual nests = 4,463) between 
2004 and 2017.
    In Costa Rica, the four beaches for which we had sufficient data to 
analyze annual nest count trends mostly exhibited declining trends. 
Tortuguero experienced the steepest decrease, with a median trend of -
10.9 percent annually (sd = 4.2 percent; 95 percent CI = -19.5 to 2.2 
percent) for data collected between 1995 and 2017. Nest counts 
decreased by -3.8 percent annually at Pacuare beach (sd = 9.3 percent; 
95 percent CI = -22.6 to 16.9 percent) between 2004 and 2017, but 
increased by 1.8 percent annually (sd = 6.0 percent; 95 percent CI = -
10.8 to 14.2 percent) at the nearby Pacuare Nature Reserve between 1991 
and 2017. Nest counts at Estacion la Tortuga deceased slightly, with a 
median trend of -0.5 percent annually (sd = 7.0 percent; 95 percent CI 
= -15.7 to 13.1 percent) between 2002 and 2017.
    For the United States, we analyzed annual nest count trends for 
Florida (statewide data collected between 2008 and 2017), three beaches 
in Puerto Rico, including Culebra (1984 to 2017), Luquillo-Fajardo 
(1996 to 2017), and Maunabo (1999 to 2017), and Sandy Point National 
Wildlife Refuge in St. Croix, USVI (1982 to 2017). The median trend for 
Florida was a decline of -2.1 percent annually (sd = 13.0 percent; 95 
percent CI = -28.3 to 25.5 percent; f = 0.582; mean annual nests = 
1,288). Culebra nests decreased by -3.7 percent annually (sd = 5.3 
percent; 95 percent CI = -14.9 to 6.8 percent; f = 0.791; mean annual 
nests = 153), while nests increased by 15.9 percent annually at 
Luquillo-Fajardo (sd = 5.5 percent; 95 percent CI = -7.1 to 15.3 
percent; f = 0.805; mean annual nests = 283) and by 7.7 percent 
annually at Maunabo (sd = 4.9 percent; 95 percent CI = -2.7 to 17.4 
percent; f = 0.945; mean annual nests = 161). In St. Croix, nests 
increased by 1.7 percent annually (sd = 4.6 percent; 95 percent CI = -
7.8 to 10.7 percent; f = 0.660; mean annual nests = 399).
    These trend data are similar to other recent findings, adding 
further

[[Page 48340]]

confidence in declining trends at multiple large nesting aggregations. 
Because of concerns about declining nest counts throughout the region, 
the National Fish and Wildlife Foundation (NFWF) convened a NW Atlantic 
Leatherback Working Group (i.e., the Working Group) to assess recent 
nesting data and complete a region-wide trend analysis (NW Atlantic 
Leatherback Working Group 2018). The trend analyses conducted by the 
Working Group used leatherback nesting data from 23 sites from 14 
different nations with at least 10 years of data with consistent 
within-site methodology, analyzing data for three time periods: 1990 to 
2017, 1998 to 2017, and 2008 to 2017. Our approach to trend analyses 
was similar to that used by the Working Group in that both approaches 
involved Bayesian analyses of data meeting set criteria. However, the 
Team decided against aggregating the data over the DPS due to 
incongruity of data collection methods, collection dates and duration, 
and reporting. Despite these differences, the overall conclusion was 
the same--an overall declining nest trend.
    The Working Group found that regional, abundance-weighted trends 
were negative for all three time periods and became more negative in 
the more recent time series (NW Atlantic Leatherback Working Group 
2018). Specifically, overall nesting trends decreased at -4.21 percent 
annually from 1990 to 2017 and at -5.37 percent annually from 1998 to 
2017, with the most notable decrease (-9.32 percent annually) occurring 
during the most recent time frame of 2008 to 2017. While site-level 
trends showed variation within and among sites and across the time 
periods, overall the sites also reflected the same regional pattern: 
More negative trends were apparent during the most recent time frame. 
Seven sites had significant positive nesting trends from 1990 to 2017, 
but no sites exhibited significant positive trends from 2008 to 2017. 
The significant decline observed at Awala-Yalimapo, French Guiana (-
12.95 percent annually from 1990 to 2017, -19.05 percent annually from 
1998 to 2017, and -31.26 percent annually from 2008 to 2017), drove the 
regional results, but similar significant declines were found at other 
nesting beaches for the longer time period, including: St. Kitts and 
Nevis (-12.43 percent annually), Tortuguero, Costa Rica (-10.42 percent 
annually), Suriname (-5.14 percent annually), and Culebra, Puerto Rico 
(-4.61 percent annually). It should be noted that the other nesting 
beach in French Guiana (Cayenne) demonstrated an increasing trend (7.44 
percent annually from 1990 to 2017 and 8.19 percent annually from 1998 
to 2017). However, it exhibited a decreasing trend (-14.21 percent 
annually) from 2008 to 2017. While nesting increased over time at 
Cayenne, this increase has apparently not resulted from females 
shifting from Awala-Yalimapo, as turtles that nest at Cayenne are 
genetically distinct (Molfetti et al. 2013) and females tagged in 
Awala-Yalimapo are not seen in Cayenne or vice versa (NW Atlantic 
Leatherback Working Group 2018).
    These modeling results demonstrate that there has been a decline in 
NW Atlantic nesting from 1990 to 2017, with the most significant 
decreases occurring from 2008 to 2017. Some nesting beaches 
demonstrated positive trends for the longer time period. However, none 
showed significant increases over the most recent time period. The 
cause for the decline is uncertain, but the Working Group identified 
anthropogenic sources (e.g., fisheries bycatch), habitat losses, and 
changes in life history parameters (such as remigration interval) as 
potential drivers of the regional decline. While these results were 
taken into consideration by the Team when evaluating the extinction 
risk of the NW Atlantic DPS, the Team also performed its own trend 
analysis of the data provided to the Team so that the trends were 
calculated in a manner consistent with other DPSs. Regardless, both 
trend analyses conclude that the NW Atlantic DPS is experiencing a 
significant decline in nesting.
    In-water abundance studies of leatherback turtles are rare. 
Archibald and James (2016) assessed the relative abundance of turtles 
at a foraging area off Nova Scotia, Canada, from 2002 to 2015. This 
study evaluated opportunistic sightings per unit effort and found a 
mean density of 9.8 turtles per 100 km\2\, representing the highest in-
water density of leatherback turtles reported to date. Archibald and 
James (2016) concluded that the relative abundance of foraging 
leatherback turtles off Canada exhibited high inter-annual variability 
but, overall, showed a stable trend from 2002 to 2015. The authors 
reported that (at that time) these results were consistent with the 
stable or, in some cases, increasing trends reported for contributing 
NW Atlantic nesting beaches over the last decade (Dutton et al. 2005; 
Girondot et al. 2007; Fossette et al. 2008; McGowan et al. 2008; 
Stewart et al. 2011; Rivas et al. 2015). While there were no 
indications of a decreasing trend, the results should be interpreted 
with caution because of the small study area, opportunistic data 
collection, availability bias variance, and lack of understanding of 
the relative density outside the study area (Archibald and James 2016).
    Despite the declining trend in nesting, productivity parameters for 
the DPS are similar to the species' averages (though some may be 
declining, as we discuss below). While there is some variation, most 
productivity parameters are relatively consistent throughout the DPS. 
The overall survival rate for nesting females is relatively high at 85 
percent (Pfaller et al. 2018), with mean estimates of 0.70 to 0.99 in 
French Guiana (Rivalan et al. 2005, 2008), 0.89 in St. Croix (Dutton et 
al. 2005), and 0.89 to 0.96 on the Atlantic coast of Florida (Stewart 
et al. 2007, 2014). Remigration intervals range from 1 to 11 years 
(Schulz 1975; Boulon et al. 1996; Chevalier and Girondot 1998; 
Hilterman and Goverse 2007; Eckert et al. 2012; Stewart et al. 2014; 
Rivas et al. 2016; Garner et al. 2017). In St. Croix and St. Kitts, the 
median remigration interval appears to be increasing (4.5 years; K.R. 
Stewart, The Ocean Foundation and C. Lombard, USFWS, pers. 2019; K.M. 
Stewart, Ross University School of Veterinary Medicine and St. Kitts 
Sea Turtle Monitoring Network, pers. comm., 2019). Averaging all 
available data, the mean remigration interval for the DPS is 2.7 years, 
rounded to 3 years for use in our calculation of the index of nesting 
female abundance. Average clutch frequency per nesting season ranges 
from 3.6 to 8.3 throughout the region, with an overall mean of 5.5 
nests per season, interspersed with 9 to 10 day internesting intervals 
(Eckert et al. 2015; Garner et al. 2017). Recent records indicate that 
nesting females deposit 80 to 88 eggs per clutch. However, an early 
study by Carr and Ogren (1959) reported only 67 eggs per clutch. 
Hatching success is highly variable for nests that remain in situ, even 
for those that are viable and do not experience significant inundation 
or predation, with estimates as low as 8.9 percent in Costa Rica 
(Tro[euml]ng et al. 2007) and 10.6 percent in Suriname (Hilterman and 
Goverse 2007) and as high as 93.4 percent in Florida (Perrault et al. 
2012). Overall, hatching success is estimated at approximately 50 
percent (Eckert et al. 2012). Hatchling sex ratios often exhibit a 
female bias, but less so than for other sea turtle species, with 
estimated production of anywhere from 30 to 100 percent females in 
Suriname, Tobago, Colombia, and Costa Rica (Mrosovsky et al. 1984; 
Dutton et al. 1985; Godfrey et al. 1996; Leslie et al. 1996; Mickelson 
and

[[Page 48341]]

Downie 2010; Pati[ntilde]o-Mart[iacute]nez et al. 2012). However, the 
proportion of females documented in foraging individuals and strandings 
ranges from 57 to 70 percent (Murphy et al. 2006; James et al. 2007; 
TEWG 2007), and the ratio of females to males during an individual 
breeding season is thought to be closer to 1:1 (Stewart and Dutton 
2014).
    We conclude that the DPS exhibits a declining nest trend. In 
addition, there are indications of decreased productivity within the 
DPS. In St. Croix, one of the most thoroughly monitored nesting beaches 
in this DPS, the data from 1981 to 2010 indicate that hatching success 
and clutch frequency are declining and remigration intervals are 
increasing (Garner et al. 2017). Overall, we have a high degree of 
confidence in the decreasing nest trend and productivity metrics for 
this DPS, due to the large amount of data available from the largest 
nesting aggregations. We acknowledge that data are not available from 
all nesting beaches, but the data that we have relied upon is the best 
available and meets established standards. The declining trends reflect 
reduced nesting female abundance. In addition, longer remigration 
intervals and/or reduced clutch frequencies may play a role in this 
decline. The decline reflects a reduction in productivity that places 
the DPS at risk given the magnitude and duration of the decreasing 
trend.

Spatial Distribution

    The DPS has a broad spatial distribution for both foraging and 
nesting. There is significant genetic population structure, with 
subpopulations connected via various levels of gene flow and 
metapopulation dynamics. Tagging and telemetry studies indicate 
considerable mixing of leatherback turtles among nesting beaches and at 
multiple foraging areas throughout the North Atlantic Ocean.
    Nesting is widespread throughout the NW Atlantic beaches, occurring 
primarily as scattered, small aggregations throughout the Wider 
Caribbean, but with larger concentrations of nesting activity at 
certain sites in Trinidad, French Guiana, Suriname, Trinidad, Colombia, 
Panama, Costa Rica, Puerto Rico, St. Croix, and Florida (Horrocks et 
al. 2016).
    Genetic sampling in the NW Atlantic DPS has been generally 
extensive with good coverage of large populations in this region. 
However, sampling from some smaller Caribbean nesting aggregations is 
absent, and there are gaps in sampling or analysis for nesting sites 
along the coasts of South and Central America (e.g., Guyana, Venezuela, 
Colombia, and Panama). A comprehensive survey of genetic population 
structure in the Atlantic Ocean included large sample sizes from five 
nesting populations representative of the DPS and analysis of longer 
mtDNA sequences in combination with an array of 17 nuclear 
microsatellite DNA loci (Roden and Dutton 2011; Dutton et al. 2013). 
The microsatellite data revealed fine-scale genetic differentiation 
among neighboring subpopulations (Dutton et al. 2013): Trinidad, French 
Guiana/Suriname, Florida, Costa Rica, and St. Croix. The mtDNA data 
failed to find significant differentiation between Florida and Costa 
Rica or between Trinidad and French Guiana/Suriname. However, Dutton et 
al. (2013) show that the mtDNA sequence variation had relatively low 
statistical power to detect fine scale structure compared to the 
microsatellite DNA loci. The mtDNA homogeneity between Costa Rica and 
Florida, with differentiation demonstrated at nuclear DNA loci, 
suggests that Costa Rica may be the source of founders for the Florida 
population via one or multiple recent colonization events, likely 
indicating historic connectivity rather than ongoing demographic 
connectivity (Dutton et al. 2013). Likewise the French Guiana/Suriname 
and Trinidad populations were undifferentiated with mtDNA likely 
indicating historic connectivity. However, microsatellite DNA reveal 
fine-scale genetic structure that is consistent with tagging studies 
demonstrating a lack of nesting female movement between the two nesting 
aggregations (TEWG 2007). Significant genetic differentiation has also 
been reported for Martinique and Guadeloupe and the mainland French 
Guiana rookery (Molfetti et al. 2013). St. Croix likely represents a 
broader Northern Caribbean subpopulation of the NW Atlantic population 
that includes multiple neighboring island nesting aggregations in the 
USVI and Puerto Rico. However, sampling and analysis would be required 
to determine extent of fine scale structuring (NMFS unpublished data; 
Dutton et al. 2013). The Costa Rica (Tortuguero and Gandoca) and Guiana 
(French Guiana and Suriname) nesting aggregations are distinct 
subpopulations based on microsatellite and mtDNA results (Dutton et al. 
2013), but information on tag returns indicates movement of nesting 
females between adjacent beaches of Panama, Colombia, Venezuela and 
Guyana. Therefore, these nesting aggregations have ``fuzzy'' 
boundaries, likely a result of flexible natal homing. Nesting females 
use beaches up to 400 km apart between nesting seasons (Tro[euml]ng et 
al. 2004; Chac[oacute]n-Chaverri and Eckert 2007) and up to 463 km 
apart within the same nesting season (Stewart et al. 2014). Additional 
sampling of the remaining nesting sites will be required to determine 
the extent of fine-scale structuring within the NW Atlantic DPS. 
However, the available science indicates significant substructure 
within the DPS.
    Tagging studies indicate individual movement and gene flow among 
nesting aggregations. This is facilitated by the species' flexible 
natal homing, i.e., philopatry to a region, rather than a specific 
beach. In adjacent nesting sites in French Guiana and Suriname, five to 
six percent of nesting females were observed to shift from one site to 
the other within a season (TEWG 2007), while Schulz (1971) reported 
this proportion to be slightly higher at 8.5 percent. In contrast, 35 
percent of nesting females in Gandoca, Costa Rica, were estimated to 
nest at sites other than the study site during an individual season 
(Chac[oacute]n-Chaverri and Eckert 2007). The predisposition of nesting 
females to stray within a nesting season may be influenced by the 
proximity of alternative nesting sites within a range of approximately 
200 km (Horrocks et al. 2016). However, even within a given nesting 
season, females have been observed to move as far as 369 km (Grenada), 
463.5 km (Florida), and 532 km (Dominica) from their original location 
(Horrocks et al. 2016). Among nesting seasons, interchange between 
nesting locations also appears to be frequent and wide-ranging, with 
maximum distance separating two nesting sites for an individual female 
recorded as 1,849 km over an 8-year span (Horrocks et al. 2016).
    Genetic studies have revealed that turtles from different nesting 
aggregations use the same foraging areas. Analyzing 684 longline 
bycatch samples from across the NW Atlantic in a mixed stock analysis 
and microsatellite assignment, Stewart et al. (2016) found that 
leatherback turtles from Costa Rica were caught in a higher proportion 
in the GOM (43 percent) compared to the Northeast Distant fishing zone, 
an area in the northwestern Atlantic Ocean (6 percent), while turtles 
from Trinidad and French Guiana comprised 54 percent of bycatch in the 
GOM and 93 percent in the Northeast Distant fishing zone. A study of 
turtles foraging off Nova Scotia, Canada, similarly assigned most (82 
percent) of the 288 sampled turtles to Trinidad (n = 164) and French

[[Page 48342]]

Guiana (n = 72), with 15 percent (n = 44) from Costa Rica, and the 
remainder from St. Croix (n = 7) and Florida (n = 1; Stewart et al. 
2013). These proportions generally represent the relative population 
sizes for these breeding populations. Microsatellite DNA assignment of 
wild captured or stranded males (n = 122) throughout the NW Atlantic 
and Mediterranean found that all males originated from NW Atlantic 
nesting aggregations (Trinidad: 55 percent, French Guiana: 31 percent, 
and Costa Rica: 14 percent; Roden et al. 2017). No turtles were 
identified from St. Croix or Florida. One turtle that stranded in 
Turkey was assigned to French Guiana, while strandings in France were 
assigned to Trinidad or French Guiana (Roden et al. 2017).
    The mixing of nesting aggregations at foraging areas is also 
supported by several tagging and/or satellite telemetry projects, 
conducted in U.S. waters (Murphy et al. 2006; LPRC 2014; Dodge et al. 
2014, 2015; Aleksa et al. 2018), Canada (James et al. 2005a, 2005b, 
2005c, 2006b, 2007; Bond and James 2017), Atlantic Europe and 
Mediterranean (Doyle et al. 2008; Sonmez et al. 2008), and on nesting 
beaches of various nations (Hildebrand 1987; Hays et al. 2004; 
Ferraroli et al. 2004; Eckert 2006; Eckert et al. 2006; Hays et al. 
2006; TEWG 2007; Sonmez et al. 2008; Evans et al. 2008; Fossette et al. 
2010a, 2010b; Richardson et al. 2012; Bailey et al. 2012; Stewart et 
al. 2014; Fossette et al. 2014; Horrocks et al. 2016; Chambault et al. 
2017). For instance, turtles from Nova Scotian foraging grounds were 
tracked to nesting areas off Colombia, Trinidad, Guyana, and French 
Guiana (Bond and James 2017). The reverse has also been demonstrated: 
some leatherback turtles from the western Atlantic undertake annual 
migrations to Canadian waters to forage (James et al. 2005c), 
exemplified by post-nesting adults tracked to the waters off Nova 
Scotia from a variety of nesting locations, including French Guiana and 
Trinidad (Fossette et al. 2014), Costa Rica, Panama (Evans et al. 
2008), and Anguilla (Richardson et al. 2012). The eastern and western 
GOM also provide foraging areas for this DPS (Aleksa et al. 2018), as 
observed from tracks of post-nesting turtles from Florida (Hildebrand 
1987), Costa Rica (Tortuguero, Gandoca), and Panama (Chiriqu[iacute] 
Beach; Evans et al. 2008; Evans et al. 2012). Evans et al. (2008) 
suggested that the GOM may represent a significant foraging ground for 
leatherback turtles from the Caribbean coast of Central America.
    High use foraging areas may be identified through available 
telemetry data, but the migration routes to those areas may vary. 
Ferraroli et al. (2004) tracked leatherback turtles from French Guiana 
and found turtles dispersed widely throughout the North Atlantic but 
mostly followed two dispersion patterns: (1) Moving north to the Gulf 
Stream area, where they started following the general ocean 
circulation; and (2) traveling east, swimming mostly against the North 
Equatorial Current. Fossette et al. (2014) found a relatively broad 
migratory corridor when turtles departed their nesting sites in French 
Guiana/Suriname, and their movements overlapped with turtles from 
Grenada and Trinidad. Fossette et al. (2010a, 2010b) found that turtles 
tracked from nesting beaches in French Guiana, Suriname, and Grenada 
and turtles caught in waters off Nova Scotia and Ireland displayed 
three distinct migration strategies: (1) Heading northwest to fertile 
foraging areas off the Gulf of Maine, Canada, and GOM; (2) crossing the 
North Atlantic Ocean to areas off western Europe and Africa; and (3) 
residing between northern and equatorial waters. Essentially, tagging 
data coupled with satellite telemetry data indicate that leatherback 
turtles of the NW Atlantic DPS use the entire North Atlantic Ocean for 
foraging and migration (TEWG 2007).
    Although adults forage at multiple areas throughout the North 
Atlantic Ocean (Fossette et al. 2014), the range of juvenile 
leatherback turtles may be more restricted. Using an active movement 
model, Lalire and Gaspar (2019) found that most juveniles originating 
from nesting beaches in French Guiana and Suriname cross the Atlantic 
Ocean at mid-latitudes with north-south seasonal migrations; after 
several years, they reach the coasts of Europe and North Africa. Eckert 
(2002) reviewed the records of nearly 100 sightings of juvenile (less 
than 100 cm curved carapace length (CCL)) leatherback turtles and 
determined they are generally found in waters warmer than 26 [deg]C, 
suggesting that the first portion of their life is spent in tropical 
and subtropical waters. After exceeding 100 cm CCL, distribution 
extends into cooler waters (as low as 8 [deg]C), which is considered to 
be the primary habitat for the species (Eckert 2002).
    The wide distribution of nesting and foraging areas likely buffers 
the DPS against local catastrophes or environmental changes. The fine-
scale population structure, with movement of individuals and genes 
among nesting aggregations, indicates that the DPS has the capacity to 
withstand other catastrophic events.

Diversity

    The NW Atlantic DPS exhibits spatial diversity, as demonstrated by 
insular and continental nesting, multiple diverse foraging areas, and 
moderate genetic diversity. The DPS nests along both continental and 
insular coastlines. Nesting beach habitat also shows considerable 
diversity, ranging from coarse-grained, sandy beaches to silty, 
ephemeral shorelines whose dynamics are influenced by estuarine input. 
The breadth and, in some cases, transiency, of suitable nesting habitat 
in the western North Atlantic may contribute to consistent, low-level 
flexibility in natal homing, both within and among reproductive seasons 
(Br[auml]utigam and Eckert 2006), and this flexibility is thought to 
surpass that of other sea turtle species (TEWG 2007).
    This DPS exhibits some temporal variation in nesting. Nesting 
generally begins in March or April, peaks in May or June, and ends in 
July or August (Eckert et al. 2012). In French Guiana, a second small 
nesting peak was documented in Awala-Yalimapo during December and 
January. However, the number of nests deposited during that time frame 
decreased from 700 in 1986/1987 to 40 in 1992/1993, and now only a 
small number of individuals are observed to nest during that time 
(Girondot et al. 2007). Some evidence indicates that the timing of 
nesting may be modulated by environmental characteristics distant from 
the nesting beach, such as water temperatures at foraging grounds 
(Neeman et al. 2015).
    The foraging strategies are also diverse, with turtles using 
coastal and pelagic waters throughout the entire North Atlantic Ocean 
(Fossette et al. 2014). Foraging habitats include temperate waters of 
the GOM, North Central Atlantic Ocean, northwestern shelf (United 
States and Canada), southeastern U.S. coast, the Mediterranean Sea, and 
northeastern shelf (Europe; TEWG 2007). Some post-nesting females also 
remain in tropical waters (Fossette et al. 2010). Overall, leatherback 
turtles in the North Atlantic Ocean appear to have a diverse array of 
foraging habitat available.
    Genetic diversity of the DPS is moderate, with six mtDNA haplotypes 
(Dutton et al. 2013). In St. Croix, a unique haplotype occurs at high 
frequency. The Florida and Costa Rica nesting aggregations each possess 
one unique, low frequency haplotype.
    Based upon this information, we conclude that nesting location and 
habitat are diverse, providing some level of resilience against short-
term spatial and temporal changes in the

[[Page 48343]]

environment. However, high-abundance nesting occurs only at a few 
locations (e.g., Trinidad, French Guiana, and Panama). The foraging 
diversity likely provides resilience against local reductions in prey 
availability or catastrophic events, such as oil spills, by limiting 
exposure to a limited proportion of the total population. Moderate 
genetic diversity may provide the DPS with the raw material necessary 
for adapting to long-term environmental changes, such as cyclic or 
directional changes in ocean environments due to natural and human 
causes (McElhany et al. 2000; NMFS 2017). We conclude that such 
diversity provides some level of resilience to threats for this DPS.

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    Destruction and modification of leatherback turtle nesting habitat 
results from a variety of activities including coastal development and 
construction; beach erosion and inundation; placement of erosion 
control and nearshore shoreline stabilization structures and other 
barriers to nesting; beachfront lighting; vehicular and pedestrian 
traffic; beach sand placement; sand extraction; removal of native 
vegetation; and planting of non-native vegetation (Lutcavage et al. 
1997; Bouchard et al. 1998; USFWS 1999; Dow et al. 2007; Eckert et al. 
2012; NMFS and USFWS 2013). As a result, most nesting beaches are 
severely degraded by such activities that continue to cause adverse 
impacts throughout the range of the DPS.
Coastal Development and Construction
    In many areas, nesting habitat is under constant threat from 
coastal development and construction (Dow et al. 2007; Crespo and Diez 
2016; Flores and Diez 2016). Coastal development impacts include 
construction of buildings and pilings on the beach; increased erosion; 
artificial lighting; pollution; recreational beach equipment and other 
obstacles on the beach; beach driving; increased human disturbance; and 
mechanized beach cleaning (Lutcavage et al. 1997; USFWS 1999; Hernandez 
et al. 2007; Dow et al. 2007; Trinidad and Tobago Forestry Division et 
al. 2010; Flores and Diez 2016). Driftwood found on nesting beaches 
also has the potential to alter nesting beach habitat and obstruct 
nesting females and hatchlings, as seen in Gandoca, Costa Rica 
(Chac[oacute]n-Chaverri and Eckert 2007). These threats impact nesting 
habitat by reducing the amount and quality of suitable beaches, 
preventing or deterring nesting females from using optimal locations, 
destroying nests, eggs, and hatchlings, and preventing hatchlings from 
successfully reaching the ocean (USFWS 1999; Chac[oacute]n-Chaverri and 
Eckert 2007; Hernandez et al. 2007; Witherington et al. 2014). 
Development involving the construction of tall buildings and clearing 
of vegetation can also alter sand temperatures and skew sex ratios 
(Gledhill 2007).
    Development occurs to varying extents throughout the range of the 
DPS, but most leatherback nesting occurs in proximity to some coastal 
development. The Florida shoreline is extensively developed outside 
wildlife refuges (Witherington et al. 2011). In Grenada, nearly 20 
percent of all nests surveyed from 2001 to 2005 occurred in an area 
affected by development, resulting in ongoing run-off onto nesting 
beaches (Maison et al. 2010). In Trinidad, increasing rural and 
commercial beachfront development is a concern, especially on the east 
coast where the main nesting beaches are located (Trinidad and Tobago 
Forestry Division et al. 2010), including Grande Riviere, the largest 
nesting aggregation of this DPS. Likewise, several Tobago beaches are 
densely developed for commercial tourism, resulting in reduced turtle 
access to potential nesting sites due to buildings, umbrellas, and 
other recreational equipment (Trinidad and Tobago Forestry Division et 
al. 2010). Development in Puerto Rico, in particular Playa Grande-El 
Paraiso (i.e., Dorado Beach, which is considered to be the most 
important nesting beach in Puerto Rico), is also a notable concern 
(Crespo and Diez 2016; Flores and Diez 2016). There, ecosystems 
continue to be threatened by coastal development, even though the 
coastal zone is protected by the Maritime-Terrestrial Zone designation 
(i.e., Coastal Public Trust Lands; Flores and Diez 2016).
    Coastal development likely influences leatherback nest placement 
and subsequent nest success, which is the percentage of nesting 
attempts (i.e., emergences onto the beach) that result in eggs being 
deposited. On Margarita Island, Venezuela, Hernandez et al. (2007) 
found that leatherback nesting aggregated towards the portions of the 
beach with fewer risk factors, such as light pollution and 
concentrations of beach furniture. This change in nesting behavior 
resulted in females nesting in less optimum areas (e.g., areas with 
lower hatching success), thus affecting the reproductive potential of 
leatherback turtles in this region.
    The magnitude of development is also changing in some areas, where 
nest placement and success may be affected in the future. For instance, 
the area around Cayenne, French Guiana, is undergoing increased 
urbanization and recreational use (Fossette et al. 2008). In recent 
years, nesting has increased at Cayenne and eastern beaches compared to 
the western Awala-Yalimapo beaches (R[eacute]serve Naturelle de l'Amana 
data in Berzins 2018 and KWATA data in Berzins 2018). As such, more 
nesting in French Guiana is exposed to coastal development and the 
associated threats, and these threats are likely to continue and 
increase.
Beach Erosion and Inundation
    While erosion is often intensified due to anthropogenic influences, 
natural features in some areas result in high erosion rates and 
unstable beaches, thus affecting leatherback nesting. For instance, the 
Maroni River influence in the Guianas (French Guiana especially) has 
resulted in highly dynamic and unstable beaches, with shifting mudflats 
making nesting habitat unsuitable (Crossland 2003; Goverse and 
Hilterman 2003; Fossette et al. 2008). Beaches are often created and 
lost along the coast of French Guiana (Kelle et al. 2007). For example, 
remote beaches in western French Guiana experience significant beach 
erosion and several disappeared, reducing or preventing monitoring (and 
likely nesting). In Suriname, Braamspunt Beach at the mouth of the 
Suriname River is moving west, out of the established Wia Wia Nature 
Reserve and may disappear in the next several years (M. Hiwat, WWF, 
pers. comm., 2018). This is significant in that Braamspunt is currently 
the main nesting beach in Suriname. The second highest nesting area in 
Suriname, Galibi Beach, is also experiencing significant erosion and 
becoming narrower. Similar beach erosion is occurring in Guyana, as 
well as in Trinidad and Tobago (Reichart et al. 2003; Trinidad and 
Tobago Forestry Division et al. 2010). At some Trinidad and Tobago 
nesting sites (e.g., Fishing Pond, Matura, Grande Riviere, and Great 
Courland Bay), rivers emerge onto nesting beaches and create additional 
erosion during the nesting season (Godley et al. 1993; Lee Lum 2005), 
intensifying nest loss (up to 35 percent of nests; Trinidad and Tobago 
Forestry Division et al. 2010).
    Seasonal erosion also occurs at most Caribbean nesting beaches. A 
survey of Wider Caribbean Regions found that erosion/accretion was the 
highest threat to nesting habitat (Dow et al. 2007). For example, at 
Playa Gandoca, Costa Rica, erosion from strong coastal drift currents 
is thought to be one of the largest obstacles to hatching success, 
destroying greater than 10 percent of all

[[Page 48344]]

nests laid in some years (Chac[oacute]n-Chaverri and Eckert 2007). In 
2006 and 2007, coastal erosion and inundation accounted for 33 to 42 
percent of nest loss in southern Panama and 29 to 48 percent on 
Caribbean Colombia beaches (Pati[ntilde]o-Mart[iacute]nez et al. 2008).
    Inundation of nests is also a concern. Leatherback turtles 
generally nest closer to the water than other sea turtles (Caut et al. 
2010). If nests are laid too close to the high tide line, they are 
subjected to erosion and inundation, which can result in egg mortality 
from suffocation or curtailed embryonic development (Chac[oacute]n-
Chaverri and Eckert 2007; Caut et al. 2010). This inundation phenomenon 
occurs on multiple nesting beaches and is particularly of concern in 
areas with high tidal influence and dynamic coastlines. On Krofajapasi 
beach in Suriname, 31.6 percent of nests laid by females were below the 
spring high tide level and determined to be ``doomed'' clutches (Dutton 
and Whitmore 1983). Similarly, in Gandoca, Costa Rica, 37 percent of 
nests from 1990 to 2004 were laid in the low tide zone and would have 
been inundated if not relocated (Chac[oacute]n-Chaverri and Eckert 
2007). In St. Croix, 43 percent of the nests (with a range of 25 to 68 
percent) were considered to be ``doomed'' each season (McDonald-Dutton 
et al. 2001), but beginning in 1983, all doomed clutches were relocated 
to improve hatching success (Dutton et al. 2005). Without intervention, 
these nests would likely have been lost. On Awala-Yalimapo, French 
Guiana, 27 of 89 nests were overlapped by tide at least once during the 
incubation period, and the hatching success was on average 
significantly lower in overwashed nests (Caut et al. 2010). Observed 
mortality was 100 percent in the intertidal zone at sites along the 
coasts of Panama and Colombia, with an overall nest loss by erosion and 
inundation ranging from 16 to 48 percent among three major nesting 
sites (Pati[ntilde]o-Mart[iacute]nez et al. 2008). While levels of 
inundation and resulting declines in hatching success have been noted 
at multiple sites throughout the range of the NW Atlantic DPS, the 
specific impacts of inundation may be variable. Hilterman and Goverse 
(2007) noted that leatherback nests can tolerate relatively high levels 
of inundation, so hatching may still be successful despite proximity to 
the tide line. Because of this, and because it may affect natural sex 
ratios (Mrosovsky and Yntema 1980), the relocation of nests susceptible 
to inundation was abandoned in 2002 in Suriname (Hilterman and Goverse 
2007); only nests directly threatened by beach erosion are relocated, 
under certain circumstances. Other nations still relocate nests to 
reduce the impacts of erosion. However, as mentioned, such practices 
may result in cooler nests and affect sex ratios (Spanier 2008). While 
eggs relocated to hatcheries could have been lost under natural 
circumstances, due to coastal erosion and inundation in some areas 
(Dutton and Whitmore 1983, Chac[oacute]n-Chaverri and Eckert 2007), 
hatching success in relocated nests is often lower than in situ nests 
(Revuelta et al. 2014; Valentin-Gamazo et al. 2018; Florida Department 
of Environmental Protection unpublished data 2018).
    Such naturally dynamic areas make it difficult to protect nesting 
beach habitat and accurately assess leatherback nesting trends. This is 
particularly noteworthy given that nesting females use high energy, 
erosion-prone beaches, which often result in high nest loss 
(Chac[oacute]n-Chaverri and Eckert 2007; TEWG 2007; Spanier 2008; 
Trinidad and Tobago Forestry Division et al. 2010). However, 
leatherback turtles in the Guianas seem to have adapted to this 
constant geomorphological change of beaches. When new beaches develop, 
they may be colonized within months by nesting females, who take 
advantage of the fresh, clean sand (or seashells, in Guyana) and 
absence of entangling or deep-rooted beach vegetation (TEWG 2007).
    Nest site selection by leatherback turtles is still poorly 
understood (Maison et al. 2010), but nesting females may be changing 
their nesting patterns due to erosion. Spanier (2008) found that 
nesting females at Playa Gandoca, Costa Rica, appear to actively select 
nest sites that are not undergoing extensive erosion, with slope 
considered to be the cue for site selection. A similar result was found 
on Grande Riviere, Trinidad, with a nesting shift from east to west 
throughout the season as an apparent response to erosion on the eastern 
end of the nesting beach (Lee Lum 2005). Further, Maison et al. (2010) 
studied nest placement in Grenada and discovered that leatherback 
turtles seemed to respond to the accretion of the north facing beach 
and erosion of the east facing beach in 2005 by nesting more often on 
the north facing beach. If erosion is increasing in existing nesting 
locations, nesting may occur in areas with lower success rates, thus 
affecting productivity. In addition, leatherback nests are deeper than 
those of other sea turtles; water content and salinity typically 
increase with depth, leading to a decrease in sea turtle hatching 
success (Foley et al. 2006).
Erosion Control, Nearshore Shoreline Stabilization Structures, and 
Other Barriers
    A widespread strategy to reduce coastal erosion is to construct 
erosion control structures. However, these structures reduce the amount 
of available nesting habitat. Also, when beachfront development occurs, 
the site is often engineered to protect the property from erosion. This 
type of shoreline engineering, collectively referred to as beach 
armoring, includes sea walls, rock revetments, riprap, sandbag 
installations, groins and jetties. Beach armoring can result in 
permanent loss of a nesting beach through accelerated erosion and 
prevention of natural beach/dune accretion. These impacts can prevent 
or hamper nesting females from accessing suitable nesting sites (USFWS 
1999). Clutches deposited seaward of these structures may be inundated 
at high tide or washed out entirely by increased wave action near the 
base of the erosion control structures. As these structures fail and 
break apart, they spread debris on the beach, thus creating additional 
impacts to hatchlings and nesting females.
    In the southeastern United States, numerous erosion control 
structures that create barriers to nesting have been constructed. In 
Florida, the total amount of existing and potential future armoring 
along the coastline is approximately 24 percent (164 miles; FDEP, pers. 
comm., 2018). This assessment of armoring does not include other 
structures that are a barrier to sea turtle nesting, such as dune 
crossovers, cabanas, sand fences, and recreational equipment. 
Additionally, jetties have been placed at many ocean inlets in the 
United States to keep transported sand from closing the inlet channel. 
The installation of jetties resulted in lower loggerhead and green 
turtle nesting density updrift and downdrift of the inlets, leading 
researchers to propose that beach instability from both erosion and 
accretion may discourage turtle nesting (Witherington et al. 2005). 
Leatherback nesting near jetties and inlets is low, possibly reflecting 
their avoidance of such areas. There are some efforts, such as the 
Coastal Construction Control Line Program, that provide protection for 
Florida's beaches and dunes while allowing for continued use of private 
property. However, armoring structures on and adjacent to the nesting 
beach continue to be permitted and constructed on the nesting beaches 
of Florida, as in other nations where the DPS nests.
    Due to erosion, beach nourishment is a frequent activity in some 
developed

[[Page 48345]]

areas, and many beaches are on a periodic nourishment schedule. Beach 
nourishment may result in direct burial and disturbance to nesting 
females, if conducted during the nesting season. It may also result in 
changes in sand density, beach hardness, beach moisture content, beach 
slope, sand color, sand grain size, sand grain shape, and sand grain 
mineral content, if the placed sand is dissimilar from the original 
beach sand (Nelson and Dickerson 1988; USFWS 1999). These changes can 
affect nest site selection, digging behavior, incubation temperature 
(and hence sex ratios), gas exchange parameters within incubating 
nests, hydric environment of the nest, hatching success and hatchling 
emerging success (Lutcavage et al. 1997; Steinitz et al. 1998; Ernest 
and Martin 1999; USFWS 1999; Rumbold et al. 2001; Brock et al. 2009). 
On severely eroded sections of beach, where little or no suitable 
nesting habitat previously existed, beach nourishment has been found to 
result in increased nesting (Ernest and Martin 1999). However, on most 
beaches in the southeastern United States, nesting success typically 
declines for the first year or two following nourishment, even though 
more nesting habitat is available for turtles (Trindell et al. 1998; 
Ernest and Martin 1999; Herren 1999; Brock et al. 2009). Further, 
nourishment projects result in heavy machinery, pipelines, increased 
human activity and artificial lighting on the project beach, further 
affecting nesting females and beach habitat. Overall, the impacts of 
beach nourishment to this DPS are not as widespread as other threats to 
nesting habitat, as Dow et al. (2007) found that only four nations 
(Anguilla, Cuba, Mexico, and United States) reported frequent or 
occasional beach nourishment.
Artificial Lighting
    Coastal development also contributes to habitat degradation by 
increasing light pollution, which can result in hatchling and nesting 
female disorientation, altering behavior and leading to mortality. In 
Florida, from 2013 to 2017, a total of 341 leatherback nests 
(representing the whole or majority of hatchlings in the nest) and five 
nesting females were disoriented (FWC unpublished data 2018). 
Artificial lighting ranked as the third highest threat to nesting/
hatching turtles in the Wider Caribbean Region (Dow et al. 2007). For 
example, urban development is significant in Puerto Rico, with light 
pollution (as well as coastal erosion and deforestation) occurring near 
leatherback nesting beaches (Crespo and Diez 2016). Fortunately, some 
of the major nesting beaches in this DPS are located in comparatively 
remote areas, and large-scale development is currently less of an issue 
there (Trinidad and Tobago Forestry Division et al. 2010; NMFS and 
USFWS 2013). That said, even within the same country, light pollution 
is variable. Fossette et al. (2008) reported that in French Guiana, 
light pollution from residential areas is a problem at Cayenne Beach, 
but it is not an issue at Awala-Yalimapo. Similarly, lighting is not a 
significant problem on nesting beaches in Trinidad, but is a concern in 
Tobago (Trinidad and Tobago Forestry Division et al. 2010). With the 
risk of increased development in some of these relatively remote areas, 
additional light pollution is anticipated, and disorientation of 
hatchlings and adults from such lighting may become a bigger problem. 
In Costa Rica, beachfront lighting is increasing and may become 
problematic at Gandoca Beach (Chac[oacute]n-Chaverri and Eckert 2007) 
and Tortuguero (de Haro and Tro[euml]ng 2006).
    Light pollution has been managed to some extent (Witherington et 
al. 2014). Lighting in Florida is regulated by multiple rules and 
regulations including Florida statutes, the Florida Building Code, and 
local lighting ordinances (Witherington et al. 2014). In addition, the 
Florida Department of Transportation and local governments have adopted 
lighting-design standards. A total of 82 municipalities in Florida have 
adopted lighting ordinances to minimize the impact of lighting on 
adjacent sea turtle nesting beaches (Witherington et al. 2014). 
However, compliance and enforcement is lacking in some areas. Further, 
lighting away from areas covered by beachfront ordinances is 
unregulated, resulting in urban glow. Although outreach and 
conservation programs control the impacts of lighting in some other 
locations, such as Costa Rica, Mexico, and Puerto Rico (Lutcavage et 
al. 1997; Crespo and Diez 2016), a majority of nations do not have 
regulations in place.
Sand Extraction
    Extracting sand from nesting beaches for construction projects has 
a detrimental effect on the amount of available nesting beach habitat 
and also accelerates erosion (resulting in the aforementioned 
associated impacts). Sand mining occurs in most Wider Caribbean nations 
to varying extent and frequency (Dow et al. 2007). In particular, beach 
sand mining has been extensive at Matura Bay and Blanchisseuse in 
Trinidad (Trinidad and Tobago Forestry Division et al. 2010). Some 
nations regulate sand mining: In St. Lucia, the Conservation and 
Management Act of 2014 requires a certificate of environmental approval 
for projects removing sand from nesting beaches.
Removal of Native Vegetation
    In some nations, upland deforestation and the resultant deposition 
of debris and garbage can destroy or modify nesting beaches. The debris 
can block access of gravid (pregnant) females and fatally trap emergent 
hatchlings (Chac[oacute]n-Chaverri and Eckert 2007). The accumulation 
of logs reduces the amount of available nesting habitat, possibly 
forcing leatherback females to nest in suboptimal locations (TEWG 
2007). Deforestation due to coastal development is a notable concern in 
Puerto Rico (Crespo and Diez 2016).
Vehicular Traffic
    Beach driving also occurs in most nations throughout the range of 
this DPS (Chac[oacute]n-Chaverri and Eckert 2007; Dow et al. 2007; 
Trinidad and Tobago Forestry Division et al. 2010). In the United 
States, vehicular driving is allowed on certain beaches in Florida 
(e.g., Duval, St. Johns, and Volusia Counties). Beach driving reduces 
the quality of nesting habitat in several ways. Vehicle ruts on the 
beach can prevent or impede hatchlings from reaching the ocean 
following emergence from the nest (Mann 1977; Hosier et al. 1981; Cox 
et al. 1994; Hughes and Caine 1994). Sand compaction by vehicles 
hinders nest construction and hatchling emergence from nests (Mann 
1977; Gledhill 2007). Vehicle lights and vehicle movement on the beach 
after dark can deter females from nesting and disorient hatchlings. 
Additionally, vehicle traffic contributes to erosion, especially during 
high tides or on narrow beaches where driving is concentrated on the 
high beach and foredune.
Vegetation
    Beach vegetation (native and non-native) can affect turtle nesting 
productivity by obstructing nest construction and potentially drying 
the sand (resulting in egg chamber collapse). Vegetation can form 
impenetrable root mats that can invade and desiccate eggs and affect 
developing embryos, impede hatchling emergence, and trap hatchlings 
(Conrad et al. 2011). Non-native vegetation has invaded many coastal 
areas and often outcompetes native plant species (USFWS 1999). The 
occurrence of exotic vegetation (or loss of native vegetation) was 
recognized as a medium-ranked threat in many Wider Caribbean nations

[[Page 48346]]

(Dow et al. 2007). The Australian pine (Casuarina equisetifolia) is 
particularly harmful to sea turtles (USFWS 1999). Australian pines 
cause excessive shading of the beach that would not otherwise occur. 
Studies of loggerhead turtles in Florida suggest that nests laid in 
shaded areas are subjected to lower incubation temperatures, which may 
alter the natural hatchling sex ratio (Marcus and Maley 1987; Schmelz 
and Mezich 1988). Fallen Australian pines limit access to suitable nest 
sites and can entrap nesting females (Reardon and Mansfield 1997). The 
shallow root network of these pines can interfere with nest 
construction (Schmelz and Mezich 1988). Dense stands of Australian pine 
have overtaken many coastal areas throughout central and south Florida.
    While non-native vegetation can affect nesting habitat throughout 
the range of the DPS, native vegetation can also affect productivity. 
For instance, at Sandy Point, St. Croix, changing erosion-accretion 
cycles led to native Ipomoea pes-caprae, a creeping vine, extending 
into the nesting area in some years. Nesting females at Sandy Point 
typically avoided nesting in vegetation, resulting in more nests laid 
near the high-tide line (Conrad et al. 2011). As a result, Ipomoea pes-
caprae decreased nest productivity by reducing leatherback hatching and 
emergence (percentage of hatchlings that emerge from the nest) success 
rates (Conrad et al. 2011).
Mitigations to Habitat Modification
    Nesting habitat disruptions are minimized in some areas. Several 
areas in the NW Atlantic DPS range are under U.S. Federal ownership as 
National Wildlife Refuges in Florida (Archie Carr and Hobe Sound), 
Puerto Rico (Culebra and Vieques) and St. Croix (Sandy Point). Beaches 
in some Wider Caribbean countries are also protected. In Trinidad, 
Matura and Fishing Pond beaches were declared Prohibited Areas in 1990, 
and the nesting beach at Grande Riviere in 1997. In 1998, the Amana 
Nature Reserve, which includes Awala-Yalimapo beach and a 30 m wide 
marine fringe, was established in French Guiana. In Suriname, the Wia 
Wia Nature Reserve was implemented in 1961 (amended and enlarged in 
1966 to protect sea turtles), and in 1969, the Marowijne beaches were 
declared a sanctuary (the Galibi Nature Reserve; Schulz 1971). In 
addition, Tortuguero National Park, Costa Rica, was established in 1976 
to protect nesting habitat (Bjorndal et al., 1999). Terrestrial habitat 
in these areas is therefore protected from the above threats to some 
extent. USFWS and NMFS also designated as critical habitat for 
leatherback turtles the nesting beaches at Sandy Point, St. Croix (43 
FR 43688; September 26, 1978) and surrounding marine waters (44 FR 
17710; March 23, 1979), which benefits the turtles in this DPS. 
However, if ESA protections did not continue (i.e., if this species 
were no longer listed), these protections would be lost.
Marine Habitat Modifications
    In the marine environment, habitat threats include anthropogenic 
noise and offshore lighting. We discuss other threats to marine habitat 
and prey (e.g., marine pollution, oil exploration, and climate change) 
in later sections. Anthropogenic noise impacts the marine habitat of 
the DPS. Dow Piniak et al. (2012) measured hearing sensitivity of 
leatherback hatchlings. They found that hatchlings are able to detect 
sounds underwater and in air, responding to stimuli between 50 and 1200 
Hz in water and 50 and 1600 Hz in air, with maximum sensitivity between 
100 and 400 Hz in water and 50 and 400 Hz in air. This sensitivity 
range overlaps with the frequencies and levels produced by many 
anthropogenic sources used in the North Atlantic, including seismic 
airgun arrays, drilling, low frequency sonar, shipping, pile driving, 
and operating wind turbines. These noise sources may affect leatherback 
turtles' marine habitat and subsequently impact distribution and 
behavior. Offshore artificial lighting occurs in some marine waters of 
this DPS (Dow et al. 2007) but is less of a threat than beachfront 
lighting throughout the range of the DPS.
Summary
    We conclude that nesting females, hatchlings, and eggs are exposed 
to the loss and modification of nesting habitat, especially as a result 
of coastal development and armoring, erosion, and artificial lighting. 
These threats impact the DPS by reducing nesting and hatching success, 
thus, lowering the productivity of the DPS. Based on the information 
presented above, we conclude that habitat reduction and modification 
pose a threat to the NW Atlantic DPS.

Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    Overutilization is a threat to the NW Atlantic DPS, mostly due to 
poaching of turtles and eggs in certain nations. Legal harvest of 
turtles and eggs also occurs in some nations.
    While the vast majority of nations within the range of the NW 
Atlantic DPS protect leatherback turtles from harvest, it is legal in 
some Caribbean and Central American nations (Brautigam and Eckert 2006; 
Dow et al. 2007; Richardson et al. 2013; Horrocks et al. 2016). For 
example, the harvest of leatherback turtles over 20 pounds is allowed 
in Montserrat and Dominica from October 1 to May 31; Saint Lucia allows 
leatherback turtles over 65 pounds to be taken from October 2 to 
February 27; and St. Kitts and Nevis allows take of leatherback turtles 
over 350 pounds from October 2 to February 27 (Montserrat Turtles Act 
2002; Br[auml]utigam and Eckert 2006). In some nations, commercial use 
is prohibited, but traditional use is allowed, which can still diminish 
protection. In Colombia, subsistence fishing of sea turtles is 
permitted, and indigenous use is allowed in Honduras. Traditional or 
cultural use is permitted in Belize with prior approval (Br[auml]utigam 
and Eckert 2006). However, regular leatherback nesting does not occur 
in Belize, and its occurrence in surrounding waters is infrequent, 
reducing the impact of such mortality. Legal harvest throughout the 
range of this DPS is not monitored, and the precise magnitude of this 
threat is not clear. However, we conclude that legal harvest of turtles 
is significant because, when it occurs, nesting turtles are targeted, 
removing the most important individuals from the population. More 
often, leatherback eggs, rather than turtle meat, are harvested (TEWG 
2007; Pati[ntilde]o-Mart[iacute]nez et al. 2008), reducing productivity 
in the DPS.
    Poaching of turtles and eggs occurs throughout the NW Atlantic DPS, 
and Dow et al. (2007) ranked it as a threat for all turtle species on 
the beaches in the Wider Caribbean Region. In Panama, interviews with 
locals revealed that the development of a new way for cooking 
leatherback turtle meat has resulted in a recent increase of its 
consumption in Changuinola, Bocas del Toro Province (CITES Secretariat 
2019). Adult turtles are killed in Panama and on remote beaches in 
Trinidad and Tobago (Tro[euml]ng et al. 2002; Ordo[ntilde]ez et al. 
2007; Trinidad and Tobago Forestry Division et al. 2010). Most 
poaching, however, targets eggs, and the level often is determined by 
how much monitoring and activity to deter poachers occur on the nesting 
beaches. Some of the highest levels of egg poaching occur throughout 
Costa Rica (Tro[euml]ng et al. 2004). Tro[euml]ng et al. (2007) found 
that, at a minimum, between 13 to 21.5 percent of nests between 2000 
and 2005 were illegally

[[Page 48347]]

collected at Tortuguero. Poaching of leatherback nests was higher 
outside Tortuguero National Park (minimum 33 percent) than within the 
National Park (minimum 9 percent) in 2005 (de Haro and Tro[euml]ng 
2006). At Pacuare Playa, Costa Rica, 55 percent of nests were poached 
in 2012 (Fonseca and Chac[oacute]n 2012) and 42 percent were poached in 
2017, which was the lowest level since Latin American Sea Turtles 
(LAST) started to monitor in 2012 (LAST 2017). Poaching at Gandoca 
Beach has decreased over time (previously 100 percent of nests were 
poached), but rates still averaged 15.5 percent annually from 1990 to 
2004 (Chac[oacute]n-Chaverri and Eckert 2007). In the Dominican 
Republic, poaching is also high. Revuelta et al. (2012) determined the 
poaching of clutches in Jaragua National Park and Saona Island ranged 
from 0 to 100 percent from 2006 to 2010, with averages of 19 percent on 
western Jaragua National Park beaches, 71 percent on eastern Jaragua 
National Park beaches, and 74 percent on Saona. Poaching also occurs at 
relatively high levels in Colombia (e.g., 22 to 31 percent of clutches 
at Playona in 2006 and 2007; Pati[ntilde]o-Mart[iacute]nez et al. 2008) 
and, to some extent, in most other Caribbean nations (e.g., Guyana and 
Grenada). Poaching is likely more prevalent, and occurs at higher 
levels, on unmonitored or unprotected beaches (Dow et al. 2007; TEWG 
2007; Tro[euml]ng et al. 2007; Trinidad and Tobago Forestry Division et 
al. 2010; K. Charles, Oceans Spirits Inc., pers. comm., 2018).
    Poaching has been significantly reduced at some nesting beaches. In 
Suriname, high levels of egg poaching (at least 26 percent of nests) 
occurred in the late 1990s, but due to better monitoring and 
enforcement, that level has been significantly reduced (Hilterman and 
Goverse 2007; M. Hiwat, WWF, pers. comm., 2018). Poaching was also a 
major problem in Trinidad, but levels have been reduced with more 
people monitoring the beach (Trinidad and Tobago Forestry Division et 
al. 2010). The Marine Turtle Conservation Act of 2004 (MTCA) funds 
activities in Panama in an attempt to reduce poaching. At Chiriqui 
Beach, Panama, intense monitoring efforts have attempted to reduce 
poaching. However, of the monitored nests, 29 leatherback nests (0.7 
percent) were still poached in 2017 (Sea Turtle Conservancy 2017). 
Further, poaching in Panama outside the monitored areas still occurs, 
with the clandestine sale of eggs widespread (Brautigam and Eckert 
2006). In St. Croix, almost 100 percent of nests were lost to poaching 
prior to 1981 (Garner et al. 2017). However, the establishment of the 
USFWS Sandy Point National Wildlife Refuge has reduced egg poaching to 
0 to 1.8 percent annually as a result of nightly patrols (Garner et al. 
2017).
    Poaching of eggs is widespread throughout the Caribbean, especially 
on beaches of Costa Rica, Dominican Republic, and Colombia. The total 
number of individuals affected by poaching cannot be quantified at this 
time. However, we conclude that many eggs and some adults are affected 
by illegal poaching at nesting beaches. Adults and eggs are also 
exposed to legal harvest in some nations. The legal and illegal harvest 
of nesting females reduces both abundance (through loss of nesting 
females) and productivity (through loss of reproductive potential), 
resulting in a high impact to the DPS. Legal and illegal egg harvest 
reduces productivity only. Thus, we conclude that overutilization poses 
a threat to the DPS.

Disease or Predation

    For the NW Atlantic DPS, information on diseases is limited, but 
predation is a well-documented threat.
    Much of the available information on disease in leatherback turtles 
was obtained by necropsy of stranded large juvenile and adult turtles; 
the health implications of various conditions reported in this species 
are incompletely understood. Solitary large intestinal diverticulitis 
of unknown etiology was found in 31 subadult and adult leatherback 
turtles stranded in U.S. waters (Stacy et al. 2015). All lesions were 
chronic and unrelated to the cause of death in all cases, although risk 
of perforation and other complications are possible. Adrenal gland 
protozoal parasites were found in 17 leatherback turtles in North 
American waters examined from 2001 to 2014; it is not currently known 
whether parasitism affects adrenal function (Ferguson et al. 2016). In 
addition, leatherback turtles are hosts for several trematode parasites 
(flatworms), known species of which also occur in hard-shelled sea 
turtles (Manfredi et al. 1996, Greiner et al. 2013). In general, 
trematodes are frequently encountered without any apparent clinical 
effect on the turtle host but can affect some heavily parasitized 
individuals. With regard to other types of potential disease-causing 
organisms, there are a small number of reports of bacterial infections 
in stranded individuals (Poppi et al. 2012; Donnelly et al. 2016). A 
variety of other bacteria have been documented in nesting females on 
beaches in Costa Rica (Santoro et al. 2008) and St. Kitts (Dutton et 
al. 2013); the majority of identified bacterial species may be 
considered as potential or opportunistic pathogens for sea turtles. A 
putative case of fibropapilloma, a virus-associated tumor-causing 
disease in sea turtles, has been reported in a leatherback; this 
disease is considered very rare in the species (Huerta et al. 2002).
    An in-water health assessment was performed on 12 turtles directly 
caught at-sea and seven turtles bycaught in fishing gear in the NW 
Atlantic Ocean (Innis et al. 2010). Most were determined to be in good 
health, but several exhibited evidence of past injuries. The blood 
chemistry of entangled turtles indicated stress, seawater intake, and 
reduced food consumption associated with entanglement. In addition, 
Perrault et al. (2012) examined baseline blood chemistry metrics (i.e., 
plasma protein electrophoresis, hematology, and plasma biochemistry) as 
indicators of health for nesting females in Florida. They found that 
multiple measures of maternal health significantly correlated with 
leatherback hatching and emergence success (the percentage of 
hatchlings that emerge from the nest).
    From these data, we estimate that the exposure of eggs, juveniles, 
and adults to disease is low. The impact of disease cannot be 
quantified at this time as we have no documentation of any deaths or 
reductions in productivity directly related to disease. However, 
disease may compound the effects of or have synergistic effects with 
other threats to the species and related physiologic derangements. We 
conclude that disease, alone or in combination with other threats, is 
likely a threat to the DPS.
    Throughout the range of the DPS, predation is a threat to 
leatherback eggs, hatchlings, and adults. Eckert et al. (2012) provides 
an exhaustive list of the documented predators for each life stage and 
area. For eggs in the NW Atlantic DPS, predators include ants (Dorylus 
spininodis), fly larvae (Diptera spp.), locust larvae (Acrididae spp.), 
mole crickets (Scapteriscus didactylus), ghost crabs (Ocypode 
quadratus), vultures (Cathartidae), dogs (Canis familiaris), cattle 
(Bos taurus; due to trampling), armadillo (Dasypodidae), opossum 
(Didelphis marsupialis), coati (Nasua spp.), and raccoons (Procyon 
lotor); see Eckert et al. 2012).
    In particular, dog predation of eggs occurs in many areas (e.g., 
Colombia, French Guiana, Guyana, Panama, Puerto Rico, and Trinidad and 
Tobago). In Trinidad, where the largest nesting aggregation occurs, 
feral dogs are

[[Page 48348]]

considered to be the primary threat to eggs, even above poaching and 
coastal erosion (Trinidad and Tobago Forestry Division et al. 2010). On 
Chiriqui Beach, Panama, 54 percent of the monitored leatherback nests 
were depredated by dogs in 2003 and approximately eight percent in 2004 
(Ordo[ntilde]ez et al. 2007). Such predation may been reduced as a 
result of protection efforts funded by the MTCA. In Playa California, 
Maunabo, Puerto Rico, more than 30 percent of the leatherback nests 
were depredated by stray dogs in 2012 (Crespo and Diez 2016). A public 
outreach project in Puerto Rico was established in 2013 to reduce this 
impact. Puerto Rico is a U.S. territory; if ESA protections were 
removed, it is likely that predation rates would be higher.
    Egg predation by other species is also a notable concern in some 
areas. On Gandoca Beach, Costa Rica, dipteran larvae infestation 
exceeded 75 percent of nests in 2005 and 2006 (Gautreau et al. 2008). 
In French Guiana, on average, mole crickets preyed on 18 percent of all 
eggs (Maros et al. 2003). These threats are likely to continue, as no 
predator screening typically occurs in Wider Caribbean nations due to 
the potential for increased poaching as well as logistical difficulties 
in these areas of high density nesting. Nest loss to predators was 
found to be the seventh ranked threat to turtles (all species, not 
specific to leatherback turtles) on nesting beaches in the Wider 
Caribbean Region, and have been noted to frequently occur in Honduras, 
Mexico, Panama, Puerto Rico, and Venezuela (Dow et al. 2007).
    Hatchlings are preyed upon by a wide variety of species, including 
mole crickets, ghost crabs, horse-eye jack fish (Caranx latus), gray 
snapper (Lutjanus griseus), tarpon (Megalops atlanticus), vultures, 
hawks (Accipitridae), gulls (Larus spp.), night heron (Nyctanassa 
violacea), frigate birds (Fregatidae), dogs, mongoose (Atilax 
paludinosus), coati, and raccoons (Eckert et al. 2012). Again, dogs are 
a serious threat to leatherback hatchlings in some areas, and 
especially in Puerto Rico (Crespo and Diez 2016).
    There are few documented predators to subadults and adult 
leatherback turtles, presumably because of their large size and pelagic 
behavior. Predation by sharks (Elasmobranchii) and killer whales 
(Orcinus orca) has been reported in Barbados and St. Vincent, 
respectively (Caldwell and Caldwell 1969; Horrocks 1989). Sharks have 
also been reported to prey on nesting females off St. Croix, USVI 
(DeLand 2017; Scarfo et al. 2019). Over the past 6 years, researchers 
at Sandy Point have observed an apparent increase in injuries to 
leatherback turtles (K. Stewart, NMFS, pers. comm., 2019). These 
injuries, many of them consistent with shark predation, affect up to 70 
percent of all nesting females at the beach (Scarfo et al. 2019). While 
some turtles probably survive these encounters, it is unknown how many 
encounters result in mortality or reduced nesting effort. Jaguars 
(Panthera onca) prey on nesting females in some areas, including 
Suriname, French Guiana, Guyana, and Costa Rica (see Eckert et al. 
2012). While three nesting females were killed by jaguars at 
Tortuguero, Costa Rica, from 1998 to 2005, this mortality is only 
considered to be a minor threat and is therefore unlikely to cause a 
population decline on its own (Tro[euml]ng et al. 2007). Archibald and 
James (2018) examined 228 leatherback turtles for injuries off Atlantic 
Canada and on Matura, Trinidad, and found 15.7 percent of turtles 
exhibited injuries of suspected predatory origin.
    Predation on early life stages is natural; however, at high rates, 
it reduces the viability of the DPS (see the Status Review). Predation 
primarily reduces productivity via reduced egg and hatching success and 
the loss of hatchlings. Predation on nesting females reduces abundance 
and productivity. We conclude that predation is a threat to the NW 
Atlantic DPS.

Inadequacy of Existing Regulatory Mechanisms

    Many regulatory mechanisms (including state, Federal and 
international) have been promulgated to protect leatherback turtles, 
eggs, and nesting habitat throughout the range of the NW Atlantic DPS. 
We reviewed the objectives of each regulation and to what extent they 
adequately address the targeted threat (i.e., the threat that the 
regulation was intended to address). The effectiveness of many 
international regulations was evaluated by Hykle (2002), who found that 
international instruments often do not realize their full potential, 
either because they do not include all key countries, do not 
specifically address sea turtle conservation, are handicapped by the 
lack of a sovereign authority that promotes enforcement, or are not 
legally binding.
    National regulatory mechanisms are described in full in the Status 
Review Report. Although these regulatory mechanisms provide some 
protection to the species, most inadequately reduce the threat they 
were designed to address, generally as a result of poor implementation 
or incomplete enforcement. Specifically, existing regulatory mechanisms 
continue to be inadequate to control impacts to nesting beach habitat 
and overutilization (harvest of turtles and eggs) for this DPS. In 
addition, regulatory mechanisms are inadequate to reduce several other 
threats including bycatch in fishing gear, vessel strikes, and marine 
debris. Despite existing regulatory mechanisms, bycatch from fisheries 
(discussed in detail along with existing regulatory mechanisms in the 
Fisheries Bycatch section), incomplete nesting habitat protection, and 
poaching remain major threats to the DPS.

Fisheries Bycatch

    Fisheries bycatch is the primary threat to the NW Atlantic DPS. 
Bycatch occurs throughout the range of the DPS, affecting juveniles, 
subadults, and adults.
    Finkbeiner et al. (2011) analyzed sea turtle bycatch across all 
commercial U.S. fisheries from 1990 to 2007. They examined sea turtle 
bycatch reduction based on the year a particular fishery implemented 
bycatch reduction measures. Prior to implementing bycatch reduction 
measures, approximately 3,800 leatherback interactions, of which 2,300 
were lethal, occurred in U.S. Atlantic Ocean and GOM commercial 
fisheries annually. After bycatch reduction measures were implemented, 
1,400 leatherback turtles, 40 of those dead, were estimated to be taken 
annually in the Atlantic Ocean. The Atlantic/GOM pelagic longline 
fishery was responsible for the most annual interactions (n = 900) and 
mortality events (n = 17) in the Atlantic Ocean, followed by the 
southeast Atlantic/GOM shrimp trawl fishery (Finkbeiner et al. 2011). 
These estimates represent minimum numbers of actual bycatch and 
mortality. Because the observer coverage for these fisheries is low (so 
some bycatch may not be observed and observed effort may not be a true 
representation of actual fleet effort), not all fisheries are observed 
and thus some are not included in these estimates. Interactions are 
difficult to observe if gear modifications are in place, and so the 
methods used are conservative (Finkbeiner et al. 2011).
    In the Wider Caribbean Region, reports of leatherback bycatch in 
fisheries are common. In a survey of Caribbean nations, Dow et al. 
(2007) ranked fisheries bycatch among the highest in-water threat to 
sea turtles. Many fisheries in less industrialized nations are coastal 
and small-scale, but these fisheries are reported to have significant 
ecological impacts due to their high bycatch discards and impacts

[[Page 48349]]

to the marine environment (Shester and Micheli 2011). Of particular 
concern are leatherback bycatch in artisanal nearshore and offshore 
gillnet, longline and trawl fisheries (Barrios-Garrido and Montiel-
Villalobos 2016). Information on fisheries bycatch is collected mostly 
from stranding records but also from fisher surveys (Moncada et al. 
2003; Delamare 2005; Madarie 2006, 2010, 2012) and observations of 
nesting females. Hilterman and Goverse (2007) recorded fisheries 
related injuries on nesting females in Suriname. In 2002, 16.9 percent 
of the nesting females had fisheries- related injuries; in 2003, at 
least 18.3 percent had such injuries; and in 2005, 9 percent (Hilterman 
and Goverse 2007). From 2000 to 2003, an average of 28 leatherback 
turtles stranded on the Suriname survey beaches. Although no cause of 
death was immediately apparent, Hilterman and Goverse (2007) indicated 
that the mortalities were fisheries-related, based upon the fisheries 
that occur offshore with high bycatch and documented fisheries-related 
injuries on nesting leatherback turtles at the same time. On the 
western oceanic nesting beaches of French Guiana, injuries consistent 
with fisheries interactions (e.g., scars, wounds) were recorded on 8.4 
percent (n = 1,259) of nesting females in 2003 (Morisson et al. 2003). 
In Venezuela, 55 percent of strandings from 2001 to 2007 (n = 57) 
exhibited evidence of fisheries interactions (Barrios-Garrido and 
Montiel-Villalobos 2016). Most recently, an injury assessment of 228 
leatherback turtles from two foraging areas off the Atlantic coast of 
Canada and Trinidad nesting beaches found 19 percent of turtles 
exhibited injuries indicative of entanglement in lines or nets, and 17 
percent showed evidence of hooks; 62 percent of turtles assessed 
exhibited a minimum of one external injury (Archibald and James 2018).
    Fisheries bycatch also occur in the Mediterranean and eastern North 
Atlantic Ocean. Casale et al. (2003) analyzed 411 records of 
leatherback turtles in the Mediterranean, of which 152 were collected 
from Italy. Most of these records were from fishery captures (n = 170) 
or found in unknown circumstances (n = 127). Of those reported by 
fishermen, set or drift nets had the highest number of interactions 
(29.4 percent), followed by unknown fishing equipment (22.9 percent), 
longlines (20.6 percent), unspecified nets (12.9 percent), other 
fishing equipment (9.4 percent), and trawls (4.7 percent). The main 
fisheries affecting turtles in the Mediterranean (all turtle species, 
not just leatherback turtles) are Spanish and Italian surface 
longlines, North Adriatic Italian trawls, Tunisian trawls, Turkish 
trawls, Moroccan driftnets, and Italian driftnets (Cami[ntilde]as 
2004). The same types of fishing gear from other nations also affect 
turtles, but the bycatch numbers are lower (Cami[ntilde]as 2004). 
Stranding records from Portugal from 1978 to 2013 found that 49 of 275 
leatherback turtles exhibited evidence of fishery interactions (the 
cause of stranding could not be determined in most cases due to 
decomposition state; Nicolau et al. 2016). Multifilament nets accounted 
for approximately 41 percent of the strandings, followed by 
monofilament nets, traps/pots, and longlines. Coastal artisanal 
fisheries were recognized as a particular threat in Portugal.
    Based upon these summary reports and stranding assessments, it is 
clear that fisheries have a large impact on the NW Atlantic DPS. In the 
following paragraphs, we review information on specific gear 
interactions, including the following fisheries: Gillnet, longline, 
trawl, pot/trap, and other.
Gillnet Fisheries
    Gillnet fisheries are common throughout the range of this DPS. Due 
to the nature of the gear and fishing practices (e.g., relatively long 
soak times), bycatch in gillnets is among the highest source of direct 
sea turtle mortality (Upite et al. 2013; Wallace et al. 2013; Upite et 
al. 2018). Upite et al. (2018) evaluated observed fishery interactions 
and post-interaction mortality and determined a 79 percent sea turtle 
mortality rate for Northeast and Mid-Atlantic gillnet gear from 2011 to 
2015. Wallace et al. (2013) calculated leatherback bycatch in gillnets 
throughout the NW Atlantic Ocean of 0.015 turtles/set, with a 21 
percent median mortality rate (not considering post-interaction 
mortality). This gear was classified as having a relatively high 
bycatch impact on the NW Atlantic leatherback population. Small scale 
fisheries are of particular concern, given the magnitude of bycatch, 
nearshore distribution, and limited monitoring (Lewison et al. 2015). 
When nets are used in waters off nesting beaches, where leatherback 
turtles mate, nesting females and mature males are often captured and 
killed.
    The largest documented bycatch of leatherback turtles in gillnet 
gear occurs off the coast of Trinidad. Lee Lum (2006) estimated that 
more than 3,000 leatherback turtles were captured by coastal surface 
gillnets off Trinidad annually, with an approximate 30 percent 
mortality rate. These captures involved adult turtles, occurring off 
the north and east coasts of Trinidad during January to August, i.e., 
the breeding and nesting season, when nesting females and adult males 
occur in the waters off nesting beaches (Lee Lum 2006). Gilman et al. 
(2010) extrapolated leatherback bycatch estimates (Lee Lum 2006; 
Gearhart and Eckert 2007) to the entire Trinidad Spanish mackerel and 
king mackerel surface gillnet fishery, and estimated that almost 7,000 
turtles were captured in 2000. Additionally, Eckert et al. (2013) 
worked with drift gillnet fishermen to identify leatherback bycatch hot 
spots off the north and east coasts of Trinidad (where the nesting 
beaches are), with capture probability increasing from March to July 
and a secondary peak in October.
    Whereas most of the documented leatherback bycatch off Trinidad 
occurs in surface drift gillnet fisheries, bottom set gillnet fishing 
also captures leatherback turtles (Gass 2006; S. Eckert, WIDECAST, 
pers. comm., 2018). The magnitude of effort and turtle bycatch in this 
fishery are lower than for surface nets, but mortality rates are higher 
(approximately 70 percent; Gass 2006). As such, the bottom set gillnet 
fishery is thought to have a comparable level of mortality to the drift 
gillnet fishery (approximately 500 to 1,000 leatherback turtles 
annually; Gass 2006; S. Eckert, WIDECAST, pers. comm., 2018). The Sea 
Turtle Recovery Action Plan for the Republic of Trinidad and Tobago 
noted that drowning in gillnets is that nation's most significant cause 
of sea turtle mortality (Trinidad and Tobago Forestry Division et al. 
2010). Bond and James (2017) tracked a female from Canadian waters to a 
nesting beach off Trinidad, but the turtle was confirmed dead, 
entangled in coastal fishing gear, just prior to the date of her first 
predicted nesting event. Venezuelan fishers have also been seen hauling 
leatherback turtles from Trinidad waters into their boats (Brautigam 
and Eckert 2006). Together, drift and bottom-set gillnets off the 
Trinidad beaches, which host the largest nesting aggregation in the 
DPS, are estimated to kill well over 1,000 leatherback turtles 
annually, and they thus pose a large threat to the DPS.
    High levels of gillnet bycatch occur in other Caribbean and South 
American nations, also off major nesting beaches. In French Guiana, 
bycatch was confirmed to be high in the Maroni estuary (Chevalier 2001; 
Girondot 2015). In 2003, 26 leatherback turtles were caught in coastal 
gillnets and released off the Cayenne and Montjoly nesting sites 
(Gratiot et al. 2003 in TEWG 2007). Delamare (2005) conducted fishermen 
interviews and estimated an average of 1,149 leatherback captures in 
2004 and

[[Page 48350]]

2005 by bottom-set or drifting gillnets in French Guiana. No estimate 
of mortality was provided, but it is likely similar to Trinidad 
fisheries, i.e., 70 and 30 percent, respectively. In Suriname, a World 
Wildlife Fund survey of fishermen estimated leatherback bycatch in 
drifting gillnets at 584 in 2006, 174 in 2010, and 424 in 2012 (Madarie 
2006; Madarie 2010; Madarie 2012). Most of the turtles were captured 
alive. In Colombia, 10 to 40 leatherback turtles are killed annually by 
gillnets (Pati[ntilde]o-Mart[iacute]nez et al. 2008). Longline and 
driftnet gillnet fisheries in Moroccan waters off the northwestern 
Africa coast capture approximately 100 leatherback turtles annually 
(Benhardouze et al. 2012).
    Although not at as high a rate as in the Caribbean (based upon 
observed interactions), gillnet bycatch occurs in U.S. and Canadian 
waters. Although South Carolina, Georgia, Florida, Louisiana, and Texas 
have prohibited gillnets in their State waters, active gillnet 
fisheries remain in other states and U.S. Federal waters. No cumulative 
estimates of leatherback bycatch in gillnet fisheries in U.S. waters 
are available due to the limited observed interactions. However, from 
2003 to 2017, fishery observers recorded lethal and non-lethal bycatch 
in fixed sink, drift sink, and drift floating gillnets throughout the 
U.S. Atlantic Exclusive Economic Zone (EEZ) and GOM (NMFS unpublished 
data). From 2012 to 2016, 27 leatherback turtles (coefficient of 
variation = 0.71, 95 percent CI over all years: 0-68) were bycaught 
with 21 mortalities in sink gillnet gear in the Georges Bank and Mid-
Atlantic regions (Murray 2018). From 1989 to 1998, U.S. drift pelagic 
gillnets captured 54 leatherback turtles, but that gear is no longer 
used. Hamelin et al. (2017) reviewed leatherback entanglement records 
reported by Canada in Atlantic Canadian waters between 1998 and 2014. 
Gillnets, mainly targeting groundfish, were involved in 24 of 205 
entanglements (11.7 percent), particularly in Newfoundland and Labrador 
(n = 15). Often, gillnet entanglements involve the vertical lines 
associated with gear (M. James, DFO, pers. comm., 2019).
    Gillnet bycatch occurs in the eastern North Atlantic Ocean and in 
the Mediterranean Sea. As in other areas, sea turtles have the 
potential to interact with set gillnets and drift gillnets. The United 
Nations (UN) established a worldwide moratorium on drift gillnet 
fishing effective in 1992; the General Fisheries Commission for the 
Mediterranean prohibited driftnet fishing in 1997; a total ban on 
driftnet fishing by the European Union fleet in the Mediterranean went 
into effect in 2002; and the International Commission for the 
Conservation of Atlantic Tunas (ICCAT) banned driftnets in 2003. 
Nevertheless, unregulated driftnetting continued to occur in some areas 
(e.g., the Mediterranean Sea and off Europe; Pierpoint 2000; 
Cami[ntilde]as 2004). In the Atlantic Ocean, leatherback bycatch has 
been reported from NE Atlantic tuna driftnet fisheries by English, 
French and Irish vessels (Pierpoint 2000). Of 20 leatherback turtles 
found in nets in British and Irish waters (1980 to 2000), eight were 
caught in the NE Atlantic tuna driftnet fishery (with 25 percent 
mortality) and one was caught in a hake gillnet (Pierpoint 2000).
    Historically, driftnet fishing in the Mediterranean Sea caught 
large numbers of sea turtles. And today an estimated 600 illegal 
driftnet vessels operate in the Mediterranean, including fleets based 
in Algeria, France, Italy, Morocco, and Turkey (Environmental Justice 
Foundation 2007). Out of 411 records of leatherback turtles (stranded, 
captured, sighted, or found in unknown circumstances) in the 
Mediterranean Sea, 170 turtles were captured by fishermen, of which 
29.4 percent were caught by set or drift nets (Casale et al. 2003). 
Driftnets and gillnets in Greece, Israel, Italy, Tunisia and Turkey 
have reported documented leatherback interactions, and occasional 
leatherback bycatch occurs in Croatian artisanal gillnet fisheries 
(Cami[ntilde]as 2004; Ergene and Ukar 2017). In particular, Karaa et 
al. (2013) reviewed 36 leatherback bycatch records from Tunisia 
fisheries in the Gulf of Gabes, and found that gillnets are the 
dominant threat to leatherback turtles in the region. A similar result 
(e.g., gillnets being a high threat to leatherback turtles in the area) 
was found in the Adriatic Sea (Lazar et al. 2012). The first 
leatherback recorded on the Aegean coast of Turkey was caught in a 
gillnet (Taskavak et al. 1998). Further, a review by Casale (2008) 
found that leatherback turtles are taken in the drift gillnet fishery 
in Spain at a rate of 0.065 turtles/day-boat.
    Throughout the range of the NW Atlantic DPS, effective gillnet 
bycatch reduction measures have not been required, but measures to 
reduce leatherback bycatch have been discussed in some areas (e.g., 
Trinidad; Eckert 2013). If nations have a closed season for fishing, at 
least in the nesting season (e.g., Suriname; Madarie 2006), nesting 
females are afforded some level of protection from gillnet bycatch. 
Some nations have prohibited gillnet gear; St. Barthelemy does not 
allow trammel nets in its territorial waters and St. Lucia prohibits 
fishing within 100 meters of shore to protect nesting turtles. There 
are gillnet and trammel net restrictions in Curacao (Ministry of 
Health, Environment, and Nature 2014, UN Environment Programme 2017). 
In the United States, gillnets with stretched mesh seven inches and 
larger are prohibited at certain times off North Carolina and Virginia 
to protect sea turtles (50 CFR 223.206(d)(8); 71 FR 24776, April 26, 
2006). While no gear modifications are currently required for U.S. 
gillnet fisheries, Federal U.S. fisheries are subject to section 7 of 
the ESA, 16 U.S.C. 1536(a)(2), and through formal consultations on 
specific fisheries, measures may be required to minimize the impact of 
incidental take in gillnets (NMFS 2013). Regardless of some of these 
protective measures, gillnet bycatch (especially off nesting beaches) 
results in the loss of thousands of mature individuals annually.
Longline Fisheries
    Leatherback turtles are known to interact with longline fishing 
gear, most commonly pelagic longlines (Lewison et al. 2004; Zollett 
2009; Wallace et al. 2010; Wallace et al. 2013). There is significant 
concern over the effects of pelagic longline fishing, which extends 
globally throughout temperate and tropical waters, including several 
high pressure fishing areas in the North Atlantic Ocean (Fossette et 
al. 2014; Gray and Diaz 2017). In international waters, numerous flag 
states have high seas longline fisheries that frequently catch 
leatherback turtles (Lewison et al. 2004). Individuals are found 
entangled and hooked in this gear, mostly by the flippers (Witzell and 
Cramer 1995; Coelho et al. 2015; Huang 2015). Leatherback bycatch in 
longlines throughout the NW Atlantic Ocean was calculated at 0.062 
turtles per set, classifying the gear as a relatively low bycatch 
impact relative to other sea turtle populations (Wallace et al. 2013; 
Lewison et al. 2015). However, because longline fisheries are 
widespread across leatherbacks' distribution and use millions of hooks 
each year, they pose a large threat to the NW Atlantic DPS and are 
estimated to kill thousands of individuals (mature and immature) 
annually.
    Pelagic longline fishing is widespread throughout the range of the 
DPS and involves a number of nations, so an accurate estimate of total 
bycatch is difficult to obtain. In the Atlantic Ocean from 2002 to 
2013, the largest longline fishing fleets belonged to Taiwan, Japan, 
Spain, Belize, and China, with the Taiwanese fleet comprising the 
largest distant-water longline effort throughout

[[Page 48351]]

the region (Angel 2014; Huang 2015). In an assessment of the impact of 
ICCAT fisheries on sea turtles, Gray and Diaz (2017) estimated 
leatherback interactions with pelagic longlines in the ICCAT area from 
2012 to 2014 (15 to 16 fleets). Using a combination of published and 
assigned sea turtle bycatch rates as a function of estimated fishing 
effort submitted to ICCAT by its members, Gray and Diaz (2017) found a 
high degree of overlap in the central North Atlantic Ocean and 
equatorial waters (some of which are outside this DPS). Within the NW 
Atlantic region, an estimated 7,138 leatherback interactions occurred 
in 2012, 6,036 in 2013 and 4,991 in 2014 (Gray and Diaz 2017). Applying 
a reasonable estimated mortality rate of 21.4 percent, as seen in other 
high seas pelagic longline gear (Huang 2015), results in an average 
annual estimated mortality of 1,296 leatherback turtles from 2012 to 
2014. However, this is likely an underestimate of total mortality, as 
the high seas mortality rate in Huang (2015) was based upon the 
disposition of the turtle when boarded and therefore did not account 
for post-interaction mortality; 240 of 459 leatherback turtles caught 
from 2002 to 2013 were alive and 121 were of unknown status (Huang 
2015). Angel et al. (2014) conducted a risk assessment of turtles from 
the impacts of tuna fishing in the ICCAT region and found the NW 
Atlantic RMU (which is comparable to the NW Atlantic DPS; Wallace et 
al. 2010) has high-moderate vulnerability to longline gear, with as 
many as 270 million longline hooks annually from 2000 to 2009. In 
particular, Fossette et al. (2014) analyzed leatherback satellite 
tracks (converted to densities) overlaid with longline fishing effort 
from 1995 to 2009 in the Atlantic Ocean. In the North Atlantic Ocean, a 
total of four seasonal high-susceptibility areas were identified: one 
in the central northern Atlantic in international waters, one along the 
east coast of the United States, and one each in the Canary and Cape 
Verdean basins (Fossette et al. 2014). These areas partly occurred in 
the EEZs of eight nations (Cape Verde, Gambia, Guinea Bissau, 
Mauritania, Senegal, Spain/Canaries, United States, and Western 
Sahara). Given the species' flexible diving behavior, it is reasonable 
to expect that turtles are likely to encounter pelagic longlines 
throughout the Atlantic Ocean, regardless of whether they are engaged 
in foraging or migratory behavior (Fossette et al. 2014).
    Bycatch in U.S. Atlantic and GOM pelagic longlines has been 
extensively studied in the last decade. Current estimates of 
leatherback interactions with the U.S. Atlantic pelagic longline 
fishery are lower than previous years. In the late 1990s and early 
2000s, estimates of Atlantic U.S. pelagic longline bycatch were around 
1000 leatherback turtles annually (NMFS 2001; Yeung 2001; NMFS 2018), 
with bycatch rates of about 0.15 to 0.5 turtles per 1000 hooks (Watson 
et al. 2005). In 2005, after the United States required pelagic 
longline gear modifications (50 CFR 635.21), the fleet was estimated to 
have interacted with 351 leatherback turtles outside experimental 
fishing operations (Walsh and Garrison 2006). NMFS (2018) estimated 239 
leatherback interactions in the U.S. Atlantic pelagic longline fishery 
in 2011, 596 in 2012, 363 in 2013, 268 in 2014, 299 in 2015, and 339 in 
2016. The majority of interactions occurred in the GOM, Mid-Atlantic 
Bight, Northeast Coastal, and Northeast Distant areas (NMFS 2018). The 
post-interaction mortality estimate for the most recently available 3-
year period (2013 to 2015) for leatherback turtles is 30.13 percent (L. 
Desfosse, NMFS, pers. comm., 2018). Based on the average leatherback 
interaction estimate for the entire U.S. pelagic longline fleet from 
2011 to 2016 (351), the estimated average annual mortality for the U.S. 
pelagic longline fishery is 106 leatherback turtles.
    Leatherback interactions also occur in Canadian pelagic longline 
fisheries. From summer to fall, primarily on the Scotian Shelf, 
encounters with leatherback turtles have been documented in the large 
pelagic longline fishery since 2001 (DFO 2012). With observer coverage 
ranging from 5 to 30 percent since 2001, there were 102 reported 
interactions with pelagic longlines from 2001 to 2005, and 36 from 2006 
to 2010 (DFO 2012). Mortality rates are estimated to be in the range of 
21 to 49 percent, resulting in an estimated mortality of 13 to 44 
leatherback turtles annually. Based on an analysis of Canadian observer 
data from 2002 to 2010, the bycatch rate in this fishery is estimated 
to have declined from 120-190 leatherback turtles annually from 2002 to 
2006 to 60-90 leatherback turtles annually from 2006 to 2010, largely 
as a result of gear modifications (Hanke et al. 2012).
    In the Mediterranean Sea, longlining is prevalent. Drifting 
longlines targeting swordfish (Xiphias gladius), albacore (Thunnus 
alalunga), and bluefin tuna (T. thynnus) are considered to be the most 
dangerous fishing gear for turtles in the Mediterranean Sea (Lucchetti 
and Sala 2010). Drifting longlines (mainly for albacore tuna) in Spain, 
Italy, Greece, and Albania have documented leatherback interactions 
(Cami[ntilde]as 2004). In the western Mediterranean, swordfish 
longlines appeared to be responsible for most of the leatherback 
bycatch and entanglements (Cami[ntilde]as 1998; Cami[ntilde]as 2004). 
Casale et al. (2003) reviewed bycatch rates for longline fisheries 
targeting swordfish and estimated the average Mediterranean longline 
bycatch rates at 0.0025 leatherback turtles/1000 hooks, with a maximum 
rate of 0.0510 leatherback turtles/1000 hooks in the Tyrrhenian Sea of 
Italy (Casale et al. 2003; Casale and Margaritoulis 2010). Of 170 
leatherback fishery captures in fisheries from the Mediterranean Sea, 
approximately 35 involved longlines (Casale et al. 2003). While 
leatherback turtles are encountered in Mediterranean longlines, 
loggerheads are the most common species caught; only 0.1 percent of 
turtles captured during an observer program in Spain, Italy and Greece 
were leatherback turtles (3 out of 2,370 observed turtles; Laurent et 
al. 2001). However, given the extensive longline effort in the 
Mediterranean Sea (Casale 2008), leatherback bycatch in the 
Mediterranean is still a concern. Lewison et al. (2004) estimated a 
range of 250 to 10,000 leatherback turtles bycaught in the 
Mediterranean in 2000, with 6 percent observer coverage.
    Longline bycatch of leatherback turtles in the range of the NW 
Atlantic DPS also occurs in waters off Cape Verde (Melo and Melo 2013; 
Coelho et al. 2015), Morocco (Benhardouze et al. 2012), and Brazil 
(Pacheco et al. 2011). Given the wide distribution of both pelagic 
longline gear and leatherback turtles, bycatch of individuals in 
longline gear can occur wherever and whenever the gear and sea turtle 
distribution overlap.
    Large circle hooks (non-offset) have been found to reduce 
leatherback bycatch by as much as 55 percent compared to traditional J-
style hooks (Andraka et al. 2013; Coelho et al. 2015). While the 
vessels of certain nations may employ large circle hooks, there are no 
obligations for international longline fleets to adopt such bycatch 
mitigation measures (Richardson et al. 2013). In 2005, an ICCAT 
resolution encouraged circle hook research (ICCAT 2005), but no legally 
binding measure to require circle hooks exists (Gilman 2011). Without 
the widespread use of non-offset circle hooks, it is likely that the 
high bycatch rates of leatherback turtles in pelagic longline gear will 
continue throughout the North Atlantic high seas fisheries.
    Since 2004, the United States has issued regulations that require

[[Page 48352]]

modifications to pelagic longline gear in the U.S. Atlantic and GOM to 
reduce the bycatch and post-interaction mortality of sea turtles; these 
regulations (50 CFR 635.21(c)(2)) specify hook type and size (18/0 or 
16/0 circle hooks depending on the area), bait type, use of turtle 
disentangling equipment and handling guidelines. Swimmer et al. (2017) 
recently analyzed pelagic longline interactions before (1992 to 2001) 
and after (mid-2004 to 2015) these regulations were promulgated. 
Throughout the study period, 844 leatherback turtles were captured. 
Overall, turtle bycatch was highest in the Northeast Distant 
statistical reporting area (0.3 turtles/1000 hooks), followed by the 
Northeast Coastal, GOM, and Caribbean areas. Bycatch rates were higher 
for years prior to 2004; after the regulations, Atlantic leatherback 
bycatch rates declined by 40 percent (0.13 to 0.078 turtles/1000 
hooks). Within the Northeast Distant area alone, where additional 
restrictions include a large circle hook (18/0) and limited use of 
squid bait, rates declined by 64 percent (0.44 to 0.16 turtles/1000 
hooks; Swimmer et al. 2017). Gilman and Huang (2017) found similar 
results: Fish versus squid bait lowered catch rates of leatherback 
turtles, and wider circle hooks reduced leatherback catch rates 
relative to narrower J and tuna hooks. Capture probabilities are lowest 
when using a combination of circle hook and fish bait.
    Efforts have been made to reduce interactions in Canadian waters as 
well. Circle hook use has been recommended in the swordfish-directed 
Canadian longline fleet since 2003, whereas corrodible circle hooks 
have been required in the pelagic longline fishery since 2012 (DFO 
2013; C. MacDonald, DFO, pers. comm., 2019). There is no mandatory hook 
size restriction for the Canadian longline fleet, but license holders 
almost exclusively use 16/0 circle hooks (C. MacDonald, DFO, pers. 
comm., 2019). De-hooking and line-cutting kits are required on 
swordfish longline fishery vessels (C. MacDonald, DFO, pers. comm. 
2019).
    Some fishing fleets in the Atlantic Ocean (e.g., U.S., Canadian, 
ICCAT vessels) use large circle hooks and modified bait, but these 
measures are not required in all areas (Watson et al. 2005; Gilman et 
al. 2007; Gilman 2011). Some nations in the Wider Caribbean Region have 
implemented circle hook provisions; in Belize, the high seas fishing 
fleet adopted the use of circle hooks on 10 percent of the fleet and 
are required to report capture of sea turtles by longlines (Belize 
Fisheries Department 2017). Because the measures are not widely 
required, the number of vessels that do not employ bycatch reduction 
measures is likely higher than the number of vessels that do, and so we 
conclude on the basis of the best available information that 
leatherback bycatch in pelagic longline fisheries is still a 
significant threat (Lewison et al. 2015).
    Leatherback interactions with bottom longlines also occur. Directed 
shark fisheries using bottom longlines in the Atlantic Ocean and GOM 
may capture or entangle leatherback turtles (NMFS 2012), and the GOM 
reef fishery is also anticipated to take leatherback turtles (NMFS 
2011). On February 7, 2007, NMFS published a rule that required 
commercial shark bottom longline vessels to carry the same dehooking 
equipment as the pelagic longline vessels; this rule was promulgated to 
reduce post-interaction mortality (72 FR 5633).
    The Canadian east coast groundfish longline fishery targets a wide 
variety of groundfish species, including cod, haddock, pollock and 
white hake. Observer coverage has ranged from 2 to 30 percent depending 
on area, and there have been no reported interactions of leatherback 
turtles in the observer database since 2001 (DFO 2012). However, there 
have been three reports from Quebec logbooks and 10 reports of 
interactions with groundfish longline gear to non-governmental groups 
(DFO 2012). This indicates that the risk of interactions in this gear 
may be higher than documented through the observer program.
    Bottom longlines are also used in the Mediterranean Sea (Casale 
2008). While there have not been any documented leatherback captures 
from this gear type, loggerheads have been caught at high rates in 
Tunisia, Libya, Greece, Turkey, Egypt, Morocco, and Italy (Casale 
2008), and interactions with leatherback turtles are possible.
    Commercial pelagic longline fisheries do not operate in some 
Caribbean nations, such as in Panama where effort is limited to vessels 
under six tons (Executive Decree 486, December 28, 2010). However, 
other Caribbean nations allow commercial pelagic longline fishing, and 
many find leatherback turtles with longline hooks (R[eacute]serve 
Naturelle de l'Amana data in Berzins, Office National de la Chasse et 
de la Faune Sauvage, pers. comm., 2018 and KWATA data in Berzins 2018). 
While no longlines exist in the Caribbean Dutch nations of Bonaire, St. 
Eustatius and Saba, there are efforts to introduce circle hooks into 
the trolling fishery (Ministry of Economic Affairs 2014). We consider 
longline bycatch to be a widespread threat to this DPS, likely 
resulting in the loss of thousands of individuals annually.
Trawl Fisheries
    Leatherback turtles may interact with bottom and midwater trawl 
gear throughout the North Atlantic Ocean. The highest reported trawl 
bycatch of leatherback turtles of the NW Atlantic DPS is likely from 
the southeastern U.S. shrimp trawl fishery. Epperly et al. (2003) 
anticipated an average of 80 leatherback mortalities a year in shrimp 
trawl interactions, dropping to an estimate of 26 leatherback 
mortalities in 2009 due to reduction in fishing effort (Memo from Dr. 
B. Ponwith, SEFSC, to Dr. R. Crabtree, SERO, January 5, 2011). The 2014 
NMFS Southeast U.S. Shrimp Fishery Biological Opinion estimated 167 
annual leatherback captures (144 mortalities) in the Atlantic Ocean and 
GOM shrimp otter trawl fishery, with an additional 34 captures in try 
nets (single nets testing for shrimp concentrations; NMFS 2014). The 
majority of these interactions were in the GOM. However, a more recent 
study of the GOM and southeastern U.S. Atlantic coast shrimp otter 
trawl fishery found fewer leatherback captures: From 2007 to 2017, only 
3 leatherback turtles were reported in the observer data (with coverage 
levels around 2 percent of nominal days at sea; Babcock et al. 2018).
    In the mid-Atlantic and northeastern U.S. waters, observers 
reported 9 leatherback captures in bottom otter trawl gear and 5 
captures in midwater trawls from 1993 to 2017 (NMFS unpublished data 
2018). In the Wider Caribbean Region, leatherback turtles are reported 
captured in trawls in French Guiana (Ferraroli et al. 2004; TEWG 2007), 
Guyana (Reichart et al. 2003), Suriname (Madarie 2010), Trinidad 
(Forestry Division et al. 2010), and Venezuela (Alio et al. 2010).
    Since 1980, there were eight reports of leatherback turtles 
incidentally captured by trawl gear in British and Irish waters 
(Pierpoint 2000). In the Mediterranean Sea, leatherback bycatch in 
bottom trawls off Tunisia (Caminas 2004) and Egypt (Casale 2008) has 
also been reported.
    Trawl bycatch reduction measures (e.g., turtle excluder devices 
(TEDs) are in place in some nations. The southeastern U.S. shrimp 
fishery has required TEDs since the early 1990s. However, TEDs that 
were initially required for use in the U.S. Atlantic Ocean and GOM 
shrimp fisheries were less effective for leatherback turtles as 
compared to smaller, hard-shelled turtle species, because the TED 
openings were

[[Page 48353]]

too small to allow leatherback turtles to escape. To address this 
problem, NMFS issued a final rule on February 21, 2003, to amend the 
TED regulations (68 FR 8456) to require modified TEDs in the 
southeastern United States (Atlantic Area and GOM Area) that exclude 
leatherback turtles, as well as large benthic immature and sexually 
mature loggerhead and green sea turtles. TEDs are also required in 
summer flounder trawls operating off Virginia (south of Cape Charles) 
and North Carolina (64 FR 55860, October 15, 1999; 67 FR 19933, April 
17, 2002).
    TEDs are also used outside the United States. Shrimp harvested with 
commercial fishing technology that may adversely affect sea turtles 
generally cannot be imported into the United States per Public Law 101-
162, Section 609(b), enacted on November 21, 1989 (16 U.S.C. 1537 
note). The import ban does not apply to nations that have adopted sea 
turtle protection programs comparable to that of the United States 
(i.e., require and enforce TED use) or whose fishing activity does not 
present a threat to sea turtles (e.g., nations fishing in areas where 
sea turtles do not occur). Although most certifications are done on a 
national basis, the U.S. State Department guidelines allow some 
individual shipments of TED-harvested shrimp from uncertified countries 
with appropriate documentation. Approximately 40 nations are currently 
certified to import shrimp into the United States, and five fisheries 
have been determined as having their products eligible for importation 
with proper documentation (83 FR 22739, May 16, 2018). Specifically, on 
May 8, 2018, the U.S. State Department certified 13 nations on the 
basis that their sea turtle protection programs (e.g., use of TEDs) are 
comparable to that of the United States: Colombia, Costa Rica, Ecuador, 
El Salvador, Gabon, Guatemala, Guyana, Honduras, Mexico, Nicaragua, 
Nigeria, Panama, and Suriname. It also certified 26 shrimp-harvesting 
nations and one economy as having fishing environments that do not pose 
a danger to sea turtles. In addition, one fishery from a non-certified 
nation within the range of the NW Atlantic DPS (the French Guiana 
domestic trawl fishery) has been authorized to import shrimp products, 
provided certain documentation accompanies the imports. Sixteen nations 
have shrimping grounds only in cold waters where the risk of taking sea 
turtles is negligible: Argentina, Belgium, Canada, Chile, Denmark, 
Finland, Germany, Iceland, Ireland, the Netherlands, New Zealand, 
Norway, Russia, Sweden, the United Kingdom, and Uruguay. Ten nations 
(Bahamas, Belize, China, the Dominican Republic, Fiji, Jamaica, Oman, 
Peru, Sri Lanka, and Venezuela) and Hong Kong only harvest shrimp using 
small boats with crews of less than five that use manual rather than 
mechanical means to retrieve nets or catch shrimp using other methods 
that do not threaten sea turtles. Use of such small scale technology is 
not believed to adversely affect sea turtles. For those nations within 
the geographical range of the NW Atlantic DPS, the threat of shrimp 
trawling is minimized with TED use.
    TEDs are also required in trawl fleets in Trinidad, Belize, Brazil, 
and Venezuela, but those gear modifications do not currently meet the 
U.S. certification protocol. On June 20, 2019, the European Union 
passed a regulation (PE-CONS 59/1/19 Rev 1) that requires technical 
measures concerning: The taking and landing of marine biological 
resources; the operation of fishing gear; and the interaction of 
fishing activities with marine ecosystems. Specific to sea turtles, the 
regulation requires shrimp trawl fisheries to use a TED in European 
Union waters of the Indian and West Atlantic Oceans, consisting of 
waters around Guadeloupe, French Guiana, Martinique, Mayotte, 
R[eacute]union and Saint Martin.
    TEDs are not required in Mediterranean trawls. Some nations, like 
Belize, St. Barthelemy, Venezuela (industrial fishing only), and the 
Caribbean Netherlands (Bonaire, St. Eustatius, Saba), have banned 
trawling (Bolivarian Republic of Venezuela Official Gazette N[deg] 
5.877, March 14, 2008; Ministry of Economic Affairs 2016; Belize 
Fisheries Department 2017), and Costa Rica does not allow the issuance 
of any new permits for shrimp trawling (Costa Rica Ministry of 
Environment and Energy 2017). Curacao prohibits fishing in its 
territorial waters and inland bays with dragnets (and certain fish 
traps). These initiatives reduce the impact of trawling on leatherback 
turtles.
Pot/Trap Fisheries
    Leatherback turtles are commonly entangled in the vertical lines of 
pot and trap gear. Entanglements have been mostly reported from U.S. 
and Canadian waters, but line entanglements have occurred in other 
areas where similar gear is used (e.g., Britain; Godley et al. 1998).
    Due to high numbers of entanglement reports, a Sea Turtle 
Disentanglement Network (STDN) was established by NMFS in the 
northeastern United States (Maine to Virginia) in 2002. This program 
relies primarily on reports from the public and subsequent 
documentation and disentanglement by trained responders. From 2008 to 
2017, 267 leatherback entanglements were reported in vertical fishing 
line (STDN unpublished data). Of those fisheries that could be 
identified, 79 were lobster, 21 were fish traps or fish lines, 18 were 
conch (or a combination of conch and lobster), and 5 were crab gear; 
144 entanglements were from unidentifiable fishing gear. While most 
unknown vertical line entanglements likely involve pot/trap gear, this 
cannot always be conclusively determined. The majority of the 
leatherback turtle reports (67 percent) were from Massachusetts waters. 
Of the 267 leatherback entanglements, 221 were released alive and 46 
were found dead.
    Given the nature of their injuries, it is probable that not all 
animals released alive from entanglements survived. Currently there are 
limited empirical data on leatherback survival from pot/trap 
entanglements. Innis et al. (2010) found that at least some of the 
disentangled individuals were able to resume normal behavior and 
migratory patterns, but two leatherback turtles were entangled at least 
twice, and a third disentangled turtle had significant forelimb skin 
and muscle injuries. The effects of entanglement may be sub-lethal 
initially, but could result in subsequent mortality. By assessing the 
injuries experienced by each turtle that was documented to have been 
entangled and using NMFS' post-interaction mortality guidance (NMFS 
2017), the resulting mortality rate for northeastern U.S. vertical 
fishing line interactions for all sea turtle species combined was 
calculated at 55 percent from 2013 to 2017 (NMFS unpublished data). 
When the mortality estimate includes those turtles that were not 
disentangled and assumed to have died, the rate increases to 61 
percent. As a result (and applying the latest 5 year mortality rate to 
the last 10 years of entanglement data), 147 to 163 leatherback turtles 
died from vertical fishing line gear (most of which were likely pot/
trap gear) in the northeastern U.S. waters from 2008 to 2017, based on 
opportunistically reported data. An additional 36 leatherback turtles 
were reported entangled in trap/buoy lines from North Carolina to Texas 
from 2008 to 2017 (STSSN unpublished data). Of those 36 entanglements, 
32 turtles were found alive and 4 dead, but these southeastern U.S. 
numbers do not incorporate potential post-interaction mortality so the 
total lethal interactions were likely higher. Further, this information 
is likely an underestimate of actual

[[Page 48354]]

entanglements and mortality given the opportunistic reporting nature of 
the program; therefore, it is clear that leatherback interactions with 
vertical fishing lines are a threat to this DPS.
    Entanglements in Canadian waters are also frequently reported under 
circumstances similar to the U.S. STDN program, i.e., opportunistically 
by fishermen or the public. Between 1998 and 2014, 205 leatherback 
entanglements were reported in Canada along the Atlantic coast, with 
most from Nova Scotia (136) and Newfoundland (40; Hamelin et al. 2017). 
Entanglements mostly involved pot fisheries (44 percent; n = 91), 
including snow crab (n = 37), inshore lobster (n = 31), rock crab (n = 
10), whelk (n = 8), and hagfish (n = 3) fisheries. Trap net fisheries 
were involved in 26 percent of the entanglements (n = 53). Of the 
overall 205 reports, the majority of turtles were reported alive and 
successfully released (n = 174), and the other 15 percent (n = 31) were 
reported dead in gear. However, the number of dead turtles is likely an 
underestimate of actual entanglement-associated mortality (Hamelin et 
al. 2017).
    Leatherback turtles are also found entangled in vertical fishing 
lines in European waters. Since 1980, 83 leatherback turtles were 
bycaught in British and Irish waters, with the method of capture 
identified in 58 cases (Piedpoint 2000). The majority of captures (n = 
36) were rope entanglements, usually buoy lines used in pot fisheries 
for crustaceans or whelk, with a 61 percent recorded mortality 
(Pierpoint 2000).
    Some types of aquaculture use vertical lines similar to pot/traps 
and may pose an entanglement risk (Price et al. 2017). Four leatherback 
turtles (two alive, two dead) in Canadian and U.S. waters have been 
opportunistically reported in aquaculture gear to date (Price et al. 
2017). However, as this industry is anticipated to grow in the near 
future, leatherback interactions with aquaculture lines, and subsequent 
injury or mortality, may increase.
    These data comprise the best available information on pot/trap 
fishery interactions with the NW Atlantic DPS. However, due to the high 
probability of underreporting leatherback turtle entanglements by 
fishers, the ad hoc nature of public reporting, and the uncertainty 
about post-release survivorship, the leatherback mortality rate due to 
entanglements in vertical lines is likely underestimated (Hamelin et 
al. 2017). Estimates indicate that approximately 622,000 vertical lines 
are deployed from fishing gear in U.S. waters from Georgia to the Gulf 
of Maine (Hayes et al. 2018). There are currently no existing 
mitigation measures to reduce leatherback bycatch in vertical fishing 
lines, but efforts to reduce the amount of vertical lines in the water 
to assist with large whale conservation in the United States may help 
reduce the impact to the DPS (https://www.greateratlantic.fisheries.noaa.gov/protected/whaletrp/).
Other Gear Types
    Leatherback turtles are also susceptible to bycatch in pound nets, 
weirs, and purse seine fisheries. In the United States, pound nets set 
in Virginia waters have entangled leatherback turtles. On June 23, 
2006, NMFS issued a regulation (71 FR 36024) requiring offshore pound 
nets set in a portion of the lower Chesapeake Bay from May 6 through 
July 15 of each year to use modified pound net leaders, a gear 
modification consisting of vertical hard lay lines spaced at least two 
feet apart on the top portion of the leader, and eight inch or smaller 
stretched mesh on the bottom portion of the leader. From 2013 to 2017, 
16 leatherback turtles have been found entangled in the hard lay lines 
of the leaders, of which two were dead (NMFS 2018). While individuals 
may continue to be entangled in modified pound net leaders, the impact 
of the pound net fishery on the NW Atlantic DPS is likely minor given 
the few nets set in the lower Chesapeake Bay using this gear 
(approximately four to six) and the frequency of live interactions. 
From 2008 to 2017, the STDN also documented leatherback captures in 
weirs set off Massachusetts; these turtles were found alive, either 
entangled in the netting (n = 2) or free swimming in the weir (n = 4).
    Purse seines are used to catch a variety of fish species and are 
commonly used in the ICCAT area to catch tuna (Angel et al. 2014). 
Leatherback captures have occurred in Atlantic purse seine fisheries, 
and this bycatch may have a minor impact on the DPS. In British and 
Irish waters, two leatherback turtles were reported to be captured in 
purse seine gear between 1980 and 2000 (Pierpont 2000). Clermont et al. 
(2012) reported a total capture of 67 leatherback turtles in more than 
9000 observed Atlantic purse seine sets between 1995 and 2011, with 
only four found dead (representing 10 percent observer coverage). Most 
of the interactions were adults (75 percent). However, not all of the 
purse seine effort reported by Clermont occurs in the NW Atlantic DPS 
range. Thus, purse seine interactions with this DPS may be a fraction 
of the total captures reported. For those purse seines in the ICCAT 
region using fish aggregating devices and for those setting over free-
swimming tuna schools, the effort (through 2011) was concentrated in 
the tropics, off West Africa between Namibia and Mauritania and off 
Venezuela (Clermont et al. 2012; Angel et al. 2014). While leatherback 
and purse seine interactions may occur where distribution and effort 
overlap, the magnitude of the purse seine impacts on the NW Atlantic 
DPS is lower than the bycatch values presented in Clermont et al. 
(2012). Further, Angel (2014) found that the direct impacts on turtles 
from purse seine fishing operations appears to be minor in comparison 
to the impacts from longline fishing, especially as most purse seine 
captures are released alive.
Summary of Fisheries Bycatch
    We conclude that most immature and adult leatherback turtles of 
this DPS are exposed to bycatch in multiple fisheries throughout their 
range. Bycatch in gillnet fisheries, in particular, is a major threat 
with high mortality rates (Lee Lum 2006; Gilman et al. 2010; Girondot 
2015), annually killing thousands of NW Atlantic leatherback turtles. 
When set off nesting beaches, gillnets result in high mortality of 
nesting females and mature males (Lee Lum 2006; Eckert 2013). Longline 
bycatch is considered to be a widespread threat throughout the DPS and 
a primary source of leatherback mortality (Lewison et al. 2004), 
resulting in the death of thousands of leatherback turtles annually. In 
general, bycatch mortality reduces abundance by removing individuals 
from the population. When nesting females are killed, it also reduces 
productivity. We conclude that fisheries bycatch is the primary threat 
to the NW Atlantic DPS.

Vessel Strikes

    Vessel strikes are a threat to the NW Atlantic DPS. Injuries from 
vessel strikes may include blunt force trauma and propeller parallel 
slicing wounds affecting the carapace, flippers, head, and/or 
underlying organs (Work et al. 2010). Most of what is known about 
vessel strikes comes from stranding records; the most extensive 
stranding network is found in the United States: The Sea Turtle 
Stranding and Salvage Network (STSSN). In the United States (Maine 
through Texas), 957 leatherback turtles were reported stranded, 
captured, or entangled from 2008 to 2017, and of those, 204 had 
probable vessel-related injuries (STSSN unpublished data). For example, 
at least 72 leatherback turtles stranded in

[[Page 48355]]

Massachusetts with vessel strike wounds between 2006 and 2018, 
including at least three adult females that had previously been 
documented nesting in the Caribbean (Dourdeville et al. 2018; Mass 
Audubon Wellfleet Bay Wildlife Sanctuary, unpublished data, 2019). It 
is sometimes difficult to determine whether the vessel related wounds 
occurred before or after the turtle died (Stacy et al. 2015). However, 
a recent study estimated that approximately 93 percent of Florida 
stranded turtles with vessel strike wounds were killed by those 
injuries (Foley et al. 2019). Based on the best available information, 
it is reasonable to conclude that approximately 190 leatherback turtles 
were killed as a result of vessel strikes in U.S. Atlantic and GOM 
waters from 2008 to 2017. This number is likely an underestimate as 
strandings represent a small percentage of turtles that are injured or 
die at sea, and many vessel strikes are not reported, detected, or 
recovered.
    Vessel strikes have been documented in other nations as well, 
including in Portugal (Nicolau et al. 2016), Britain (Godley et al. 
1998), and off the coast of Tunisia in the Strait of Sicily (Karaa et 
al. 2013; Caracappa et al. 2017). While there is very limited 
observational information on vessel collisions in Atlantic waters of 
Canada, there has been at least one recorded vessel strike (DFO 2012). 
More recently, an injury assessment of leatherback turtles (n = 228) on 
Atlantic Canada foraging grounds and on a Trinidad nesting beach found 
only 1.3 percent of turtles exhibited injuries consistent with vessel 
strikes (Archibald and James 2018). However, this low injury rate may 
indicate that there is low survivorship of vessel strikes. Females with 
carapace damage from propellers have been also observed on Costa Rican 
nesting beaches (de Haro et al. 2006).
    Leatherback behavior data can help predict the potential for vessel 
strikes. Based on telemetry data for leatherback turtles (n = 15) on 
the northeastern U.S. shelf, leatherback turtles spent over 60 percent 
of their time in the top 10 m of the water column and over 70 percent 
of their time in the top 15 m (Dodge et al. 2014). Additional turtle-
borne camera and autonomous underwater vehicle research in the waters 
off Massachusetts suggests that turtles surface frequently and engage 
in subsurface swimming (within the top 2 m) when occupying shallow, 
well-mixed, coastal environments, increasing the probability of a 
vessel strike (Dodge et al. 2018). Based on 24 free swimming 
leatherback turtles tagged in Canadian waters from 2008 to 2013, 
Wallace et al. (2015) found these leatherback turtles primarily 
occupied the upper 30 m of the water column and had shallow 4 to 6 
minute dives. Given most leatherback activity occurs in the top 15 to 
30 meters of the water column in temperate shelf waters of the NW 
Atlantic Ocean and vessel traffic is high along the U.S. East coast, 
the risk of vessel strikes is likely higher than the documented 
interactions would suggest (DFO 2012; Hamelin et al. 2014).
    While observational data are limited, it is reasonable to conclude 
that, based upon the best available information, mortality due to 
vessel strikes may occur wherever vessel traffic and leatherback 
distribution (juvenile and adult) overlap. The impact is likely 
minimized in areas with less frequent vessel traffic (e.g., less 
developed areas) and decreased leatherback turtle presence. Nesting 
females and mature males may be especially vulnerable to vessel strikes 
because they occur in the waters off nesting beaches, which are coastal 
areas where vessel traffic is more prevalent. Vessel strikes affect the 
NW Atlantic DPS by lowering abundance (if the interaction results in 
mortality) and affecting future reproductive potential (productivity) 
when nesting females are killed. We conclude that vessel strikes pose a 
threat to the NW Atlantic DPS.

Pollution

    Pollution includes contaminants, marine debris, and ghost fishing 
gear. The detection of pollution impacts on leatherback turtles is 
opportunistic and thus likely underestimated. While plastic ingestion 
is not always fatal, it can reduce ability to feed, affect swimming 
behavior and buoyancy control, potentially lead to chemical 
contamination and chronic effects, and weaken physical condition, which 
could impair the ability to avoid predators and survive threats (Nelms 
et al. 2016). Entanglement in marine debris results in injuries that 
can reduce fitness, cause eventual death, reduce ability to avoid 
predators, reduce ability to forage and/or swim efficiently due to 
drag, and lead to starvation or drowning (Nelms et al. 2016). Pollution 
on the beach and in the water occurs throughout the range of the NW 
Atlantic DPS.
    Dow et al. (2007) defined marine pollution as agriculture, 
petroleum, sewage, industrial runoff, vessel discharges, declining 
water quality, and marine debris. They found pollution in the marine 
environment to be among the greatest threats to all sea turtle species 
in the Wider Caribbean Region. Dow et al. (2007) defined beach 
pollution as agriculture, petroleum/tar, sewage, industrial runoff, and 
beach litter/debris; they found pollution on the beach to be a threat. 
Pollution on the beach and in the water occurs throughout the range of 
the NW Atlantic DPS.
    Leatherback turtles are susceptible to adverse effects from 
pollution. Marine pollution, including direct contamination and 
structural habitat degradation, can also affect leatherback habitat. In 
particular, the Mediterranean is an enclosed sea, so organic and 
inorganic wastes, toxic effluents, and other pollutants rapidly affect 
the ecosystem (Cami[ntilde]as 2004).
    Of particular concern, due to their immune, reproductive, and 
endocrine disrupting nature, are persistent organic pollutants (POPs), 
such as polychlorinated biphenyls (PCBs), polybrominated diphenyl 
ethers (PBDEs), and pesticides (Bergeron et al. 1994; Bishop et al. 
1991, 1998; Keller et al. 2004). These chemicals have been identified 
in both adults and eggs in several areas occupied by this DPS. Guirlet 
et al. (2010) measured maternal transfer of organochlorine contaminants 
(OCs) from 38 nesting females in French Guiana. PCBs were found to be 
the dominant OC, followed by pesticides, but OC concentrations were 
lower than concentrations measured in other marine turtles (potentially 
due to the lower trophic level diet and offshore foraging areas). All 
OCs detected in nesting adults were detected in eggs, suggesting a 
maternal transfer of OCs. In French Guiana, hatching success has been 
shown to be low when OCs are present in the sand (most likely 
originating from pesticide use in plantations and malaria prophylaxis 
(Guirlet 2005). However, a link between OCs and embryonic mortality 
could not be determined (Guirlet et al. 2010). Stewart et al. (2011) 
also recorded PCB, OC, and PBDE concentrations for nesting and stranded 
leatherback turtles in the southeastern United States. Their results 
also suggested maternal transfer of POPs in leatherback turtles, but 
Stewart et al. (2011) found higher levels of PCBs and pesticides than 
those found in French Guiana (Guirlet et al. 2010). While finding that 
leatherback contaminant concentrations were substantially lower than 
concentrations in other reptile studies that demonstrated toxic 
effects, Stewart et al. (2011) suggested that sub-lethal effects 
(especially on hatchling body condition and health) may nevertheless be 
occurring in this species. De Andres et al. (2016) similarly monitored 
PCB and PBDE concentrations in eggs laid in Costa Rica (18 nests). POP 
levels were similar to those reported in French

[[Page 48356]]

Guiana nesting females (Guirlet et al. 2010) and slightly lower than 
those in Florida (Stewart et al. 2011). Further, De Andres et al. 
(2016) found a significant negative relationship between PBDE levels 
and hatching success, suggesting potential harmful effects of these 
contaminants on leatherback reproduction. OCs (and mercury) have also 
been documented in turtles that stranded in the United Kingdom (Godley 
et al. 1998). A leatherback that stranded off the coast of Wales, U.K. 
was found with PCB levels one-to-three orders of magnitude higher than 
the lowest levels reported for fish taken in the North Atlantic, but 
similar to the lowest concentrations reported from oceanic cetaceans 
(Davenport et al. 1990). Even with the recent restriction of the use of 
POPs, due to the widespread persistent nature of these chemicals and 
continuing atmospheric deposition (Ross et al. 2009) it is probable 
that similar chemical concentrations occur in other areas of this DPS.
    Other contaminants have also been documented in leatherback turtles 
and their eggs. Heavy metals (e.g., arsenic, cadmium, chromium, 
mercury, lead, etc.) enter the environment from a variety of sources 
(Guirlet et al. 2008; Perrault 2012). In particular, mercury can affect 
a variety of functional processes in wildlife, including the nervous, 
excretory and reproductive systems (Wolfe et al. 1998). Mercury, 
cadmium, and lead were recorded in nesting females (n = 46) and eggs in 
French Guiana (Guirlet et al. 2008). Maternal transfer of all three 
elements was documented, and female lead levels increased throughout 
the nesting season (Guirlet et al. 2008). This could be explained, in 
part, by external contamination via ingestion of contaminated prey or 
polluted water during nesting, as the French Guiana coast environment 
is exposed to significant environmental pollution via anthropogenic and 
natural sources. While mercury concentrations were lower than values 
reported for other sea turtle species, cadmium levels documented in 
French Guiana were at the same level shown to impact gonadal 
development in other turtle species and may impact reproductive 
processes and lower fertility (Guirlet et al. 2008). In Massachusetts, 
entangled turtles had significantly higher blood lead concentrations 
than directly captured turtles (Innis et al. 2010). While similar to 
those reported in French Guiana (Guirlet et al. 2008), blood 
concentrations of mercury and cadmium were at levels high enough to 
induce carcinogenic, teratogenic, and toxic effects in a variety of 
species (Innis et al. 2010).
    Mercury and selenium have also been recorded in nesting females and 
eggs in Florida and St. Croix. Animals persistently exposed to mercury 
can experience selenium deficiency, which is of concern because 
selenium is important to hatching and emergence success (Perrault et 
al. 2011). However, high levels of selenium can be toxic and negatively 
impact hatching success (Perrault et al. 2013). Mercury concentrations 
in nesting females from Florida were found to be higher than in St. 
Croix, which could be a result of different migratory and foraging 
areas, whereas hatchling blood mercury values were higher in St. Croix 
(Perrault et al. 2011; Perrault et al. 2013). It is interesting to note 
that in St. Croix, no correlations were found between mercury or 
selenium concentrations and hatching or emergence success, which is 
different from results in Florida (Perrault et al. 2011; Perrault et 
al. 2013). Hazard quotient results by Perrault et al. (2013, 2014) 
imply that mercury and selenium levels could pose a threat to 
leatherback turtle reproductive success and/or hatchling health and 
survival. Leatherback hatching and emergence success rates are already 
low compared to other species of sea turtles (Bell et al. 2004; 
Perrault et al. 2011), so the impacts of pollution and contamination on 
hatching success is a notable concern. In addition, mercury was found 
to be higher in adults than juveniles/sub-adults stranded along the 
U.S. Atlantic coast, suggesting potential physiological concerns due to 
accumulation and ongoing inputs into the environment (Perrault et al. 
2012). It is clear that additional long-term research is needed to 
better understand the relationship of non-essential elements in turtle 
development and reproduction.
    Marine debris (most notably plastic pollution) is a threat 
throughout the range of the NW Atlantic DPS (Girondot 2015). Several 
global reviews have outlined the persistent and widespread nature of 
the issue, both as an ingestion and an entanglement threat (Mrosovsky 
et al. 2009; Schuyler et al. 2014; Nelms et al. 2016; Lynch 2018). Law 
et al. (2010) assessed plastic content at the surface of the western 
North Atlantic Ocean and Caribbean Sea from 1986 to 2008, and found the 
highest concentration of plastic debris was observed in subtropical 
latitudes and associated with large-scale convergence zones, which 
include foraging areas targeted by leatherback turtles.
    Ingestion of marine debris is a concern for leatherback turtles, 
especially given the similarity of their preferred prey (e.g., 
gelatinous zooplankton) to some plastics. In particular, plastic bags 
appear similar to jellyfish in the marine environment, leading to 
mistaken and inappropriate triggering of the sensory cue to feed 
(Schuyler et al. 2014; Nelms et al. 2016). While plastic ingestion is 
not always fatal, it can reduce ability to feed, affect swimming 
behavior and buoyancy control, potentially lead to chemical 
contamination and chronic effects, and weaken physical condition, which 
could impair the ability to avoid predators and survive threats (Nelms 
et al. 2016).
    Marine debris ingestion can occur in any location, but given the 
enclosed nature of the sea and intense human pressure, the 
Mediterranean Sea in particular is a hot spot for plastic marine debris 
and other pollutants (Cami[ntilde]as 2004; Cozar et al. 2015). Marine 
debris ingestion has been documented from leatherback turtles stranded 
in Tunisia (Karaa et al. 2013), Israel (Levy et al. 2005), the northern 
Adriatic Sea (Poppi et al. 2012), and the Strait of Sicily (Caracappa 
et al. 2017). Of particular note, 30 to 73 percent of turtles stranded 
in the Bay of Biscay (France) were found to have ingested plastic 
annually from 1979 to 1999 (out of 87 leatherback turtles necropsied; 
Duguy et al. 2000). The seasonal rate of ingestion was inversely 
related to the abundance of jellyfish, leading the authors to propose 
that the depletion of jellyfish led to debris ingestion as potential 
prey. Cozar et al. (2015) conclude that the effects of plastic 
pollution on marine life are anticipated to be frequent in the high 
plastic-accumulation region of the Mediterranean Sea.
    In U.S. waters, marine debris ingestion has also been documented in 
stranded leatherback turtles. However, ingestion does not always cause 
mortality and is typically an incidental finding. Of 41 leatherback 
turtles necropsied from North Carolina to Texas between 2008 and 2017, 
17 had ingested plastics or marine debris (STSSN unpublished data 
2018). From Maine to Virginia during that same time period, 10 
necropsies detected ingestion, but the total number of necropsied 
turtles, out of the 677 strandings in the region, is currently unknown. 
It is likely that many more stranded turtles ingested some level of 
marine debris (STSSN unpublished data 2018). Out of 33 leatherback 
turtles examined in New York Bight (an area with dense population), 30 
percent had

[[Page 48357]]

synthetic material ingestion, mostly consisting of thin, clear plastic 
(Sadove et al. 1989). Of two leatherback turtles stranded in North 
Carolina during 2017 whose gastrointestinal tracts were analyzed, 
microplastics were present in both (Duncan et al. 2018).
    Marine debris ingestion is not limited to microplastics or plastic 
bags. Off the northeastern U.S. coast, necropsies of disentangled 
leatherback turtles that have died post-release have documented 
considerably large pieces of plastic (e.g., 83 by 35 cm) in their 
stomachs (Innis et al. 2010). These numbers likely underestimate the 
true marine debris ingestion rate because many turtles likely ingest 
marine debris and do not strand.
    Leatherback turtles can also become entangled in marine debris. 
From 2008 to 2017, the Northeast U.S. STDN documented 24 entanglements 
from miscellaneous sources not attributed to obvious fisheries 
entanglements, as described above (STDN unpublished data). These 
unknown entanglements could involve a myriad of sources but are 
considered as entangling marine debris. The Sea Turtle Recovery Action 
Plan for the Republic of Trinidad and Tobago noted that entanglement in 
lost or abandoned fishing gear (primarily nets) poses a threat to 
leatherback turtles in the marine and terrestrial environment (Forestry 
Division et al. 2010).
    Marine debris is also a problem on nesting beaches and can reduce 
nesting success. Pollution and debris often are deposited on high 
energy beaches, which are also the preferred nesting habitat of 
leatherback turtles (TEWG 2007). Coastal and inland littering (which 
can ultimately reach the sea) is a problem throughout Trinidad and 
Tobago, and ocean borne debris is particularly prevalent on the east 
and north coasts, which host the main leatherback nesting beaches 
(Trinidad and Tobago Forestry Division et al. 2010). Extensive debris 
on nesting beaches is not uncommon throughout the Caribbean, often 
carried by rivers to the sea and later washed ashore (e.g., in Costa 
Rica; Chac[oacute]n-Chaverri and Eckert 2007). Debris on nesting 
beaches may impede females during the nest-site selection stage, limit 
and degrade the amount of habitat available, and/or result in aborted 
nesting attempts (Chac[oacute]n-Chaverri and Eckert 2007). If line or 
netting is encountered on nesting beaches, entanglement of nesting 
females and hatchlings is also a risk.
    The majority of the NW Atlantic DPS is exposed to pollution 
throughout all life stages. These threats are a result of the developed 
nature of many of the nations within the range of the DPS. The degree 
of impact is difficult to quantify, especially given the widespread 
nature of pollution and the diverse types of impacts. Contaminants may 
affect this DPS by reducing productivity, if hatching success is 
lowered, and by lowering abundance, if contamination results in 
mortality. Marine debris affects the DPS by lowering abundance, when it 
causes death through ingestion or entanglement, and reducing 
productivity, when hatchlings and nesting females are affected. While, 
we do not have quantitative estimates of the number of individuals that 
are killed or injured as a result of pollution, we conclude that it is 
prevalent throughout the range of the DPS and constitutes a threat to 
the NW Atlantic DPS.

Oil and Gas Exploration

    Oil and gas activities have the potential to impact the NW Atlantic 
DPS directly (e.g., exposure to oil following oil spills) and 
indirectly (e.g., increased probability of vessel strikes and habitat 
degradation/destruction). In addition to lethal effects, sublethal 
effects may occur and include displacement from primary foraging areas 
with accompanying energy costs (TEWG 2007).
    Several areas within the range of the NW Atlantic DPS have intense 
oil and gas development and exploration close to major nesting beaches. 
The potential for oil spills is of particular concern in the Wider 
Caribbean Region due to its effect on all life stages in the marine 
environment. The biggest oil producing nations in South America are 
Brazil, Mexico, Venezuela, and Colombia. Although only three Caribbean 
nations currently have exportable oil and natural gas reserves 
(Barbados, Cuba, and Trinidad and Tobago, with Trinidad and Tobago the 
only significant exporter), in 2017, a major oil field was discovered 
off Guyana, which will likely lead to extensive new development and 
extraction. As a result, marine traffic is likely to increase in the 
area as well as the possibility for oil spills. In Panama, 
contamination from oil spills, primarily in area of the Trans-Isthmus 
oil pipeline and the Panama Canal, is of particular concern 
(Br[auml]utigam and Eckert 2006; Ruiz et al. 2006). Some Caribbean 
nations (e.g., Belize, French Guiana) have permanent moratoria on oil 
and gas exploration in offshore waters.
    In the United States, oil and gas extraction primarily occurs in 
the GOM (BOEM 2016; BOEM 2017), an area with leatherback foraging and 
migratory habitat (Aleksa et al. 2018). Increased shipping traffic and 
marine noise due to oil and gas explorations in the GOM pose a direct 
threat for leatherback turtles in foraging grounds and migratory 
routes, due to the potential for vessel strikes and harassment (Wallace 
et al. 2017; Ward 2017). Oil spills regularly occur in the GOM, from 
small amounts of varying types of oil product to large catastrophic 
spills. In 2010, a major oil spill occurred in the north-central GOM, 
affecting important foraging habitat used by leatherback turtles 
(Deepwater Horizon NRDA Trustees 2016). Evans et al. (2012) tracked a 
post-nesting leatherback from Chiriqui Beach, Panama, into the GOM 
during the Deepwater Horizon oil spill. The track followed similar 
tracks from turtles in previous years and did not seem to change once 
entering areas with visible oil slicks (on two occasions). Injuries to 
leatherback turtles caused by the GOM Deepwater Horizon oil spill could 
not be quantified (Deepwater Horizon NRDA Trustees 2016). However, 
given that the GOM is important habitat for leatherback turtles (Aleksa 
et al. 2018) and leatherback turtles were documented in the Deepwater 
Horizon oil spill zone during the oil spill period, the Deepwater 
Horizon NRDA Trustees (2016) concluded that leatherback turtles were 
exposed to Deepwater Horizon oil, and some portion of those exposed 
likely died.
    In Atlantic Canada, impacts from oil and gas may also occur. 
Several petroleum production projects occur offshore of Nova Scotia 
(https://www.cnsopb.ns.ca/offshore-activity/offshore-projects). Howard 
(2012) determined that oil pollution from coastal refineries, ships, 
small engine vessels, and oil and gas exploration and production is a 
risk to leatherback survival in Canada. There are also offshore oil and 
gas platforms in the North (United Kingdom, Denmark) and Mediterranean 
Seas, where similar impacts to leatherback turtles may also occur (EU 
Offshore Authorities Group 2018; https://euoag.jrc.ec.europa.eu/node/63). In particular, the Mediterranean Sea has been declared a ``special 
area'' by the International Convention for the Prevention of Pollution 
from Ships (MARPOL), in which deliberate petroleum discharges from 
vessels are banned, but numerous repeated offenses are still thought to 
occur (Pavlakis et al. 1996). Some estimates of the amount of oil 
released into the region is as high as 1,200,000 metric tons (Alpers 
1993). Direct oil spill events also occur, as in Lebanon in 2006 when 
10,000 to 15,000 tons of heavy fuel oil spilled into the eastern

[[Page 48358]]

Mediterranean (UN Environment Programme 2007).
    In summary, oil and gas activities are prevalent in foraging, 
migratory, and offshore nesting habitats of the NW Atlantic DPS, 
potentially exposing all life stages to oil associated threats, such as 
direct miring in oil, oil ingestion, vessel strikes, and nesting beach 
contamination. Oil and gas activities have the potential to affect this 
DPS by reducing productivity (e.g., if hatching success is reduced by 
oil spills) and potentially lowering abundance (e.g., if oil exposure 
results in mortality). As such, oil and gas activities are a threat to 
the NW Atlantic DPS.

Natural Disasters

    Natural disasters, such as hurricanes and other storms, and natural 
phenomena, such as Sargassum events on or near nesting beaches, pose a 
threat to the NW Atlantic DPS.
    Hurricanes are common in the Caribbean and southeastern United 
States. Hurricanes and tropical storms impact nesting beaches by 
increasing erosion and sand loss and depositing large amounts of 
debris. In 2017, Hurricane Maria devastated the islands of Dominica, 
St. Croix, and Puerto Rico, and even though the nesting season was 
nearly over, many beaches were impacted, including Maunabo, Puerto Rico 
(one of the most abundant nesting beaches on the island; R. Espinoza, 
Conservaci[oacute]n ConCiencia, pers. comm., 2017). Dewald and Pike 
(2014) found that a lower level of leatherback nesting attempts 
occurred on sites that were more likely to be impacted by hurricanes. 
These types of storm events may ultimately affect the amount of 
suitable nesting beach habitat, potentially resulting in reduced 
productivity, especially as leatherback turtles typically nest on high 
energy beaches (TEWG 2007).
    Hurricanes may also result in egg loss by destroying and inundating 
nests. However, hurricanes are usually aperiodic so the impacts are 
expected to be infrequent. Hurricanes also typically occur after the 
peak of the leatherback hatching season and would not be expected to 
affect the majority of incubating nests (USFWS 1999). That said, 
according to the Intergovernmental Panel on Climate Change (IPCC), 
climate change may be increasing the frequency and patterns of 
hurricanes (IPCC 2014) potentially causing such impacts to nests to 
become more common in the future.
    Sargassum is a genus of macroalgae found in temperate and tropical 
waters. When large amounts of Sargassum wash ashore, they form thick 
mats that have the potential to disrupt females' nesting activities and 
impede hatchlings' access to the ocean (Maurer et al. 2015). In 2011 
and 2015, large amounts of Sargassum were present in the Caribbean 
(mainly Trinidad and Tobago and Grenada) and frequently washed ashore, 
covering large expanses of sandy shoreline on nesting beaches. While 
females still nested in these areas, hatchlings needed intervention to 
reach the ocean (Wang and Hu 2016; Audroing, TVT, pers. comm., 2018; K. 
Charles, Ocean Spirits Inc., pers. comm., 2018). Most recently, large 
amounts of Sargassum were found in 2018 on Caribbean beaches, causing 
Barbados to declare a national emergency in June 2018. Such widespread 
blanketing of Sargassum on leatherback nesting beaches throughout the 
Caribbean has the potential to impact future hatching success and 
survival.
    In summary, natural disasters and phenomena have the potential to 
impact the NW Atlantic DPS. However, given the infrequent and temporary 
nature of the occurrences, only a small proportion of eggs, hatchlings, 
and nesting females are exposed to these threats. Impacts include egg 
and hatchling mortality that affect productivity of the DPS. Seasonal 
losses at individual beaches may be large, but we do not expect such 
impacts to be spatially or temporally widespread. However, we conclude 
that natural disasters pose a threat to the DPS.

Climate Change

    Climate change is a threat to the NW Atlantic DPS. The impacts of 
climate change include increases in temperatures (air, sand, and sea 
surface); sea level rise; increased coastal erosion; more frequent and 
intense storm events; and changes in ocean currents. These impacts may 
affect leatherbacks through alterations of the incubation environment, 
reduction of nesting habitat, and changes in prey as described in the 
following subsections.
    Modeling results show that global warming (rise in average surface 
temperature) poses a ``slight risk'' to females nesting in French 
Guiana and Suriname relative to those nesting in Gabon, Congo, and West 
Papua (Dudley et al. 2016). As global temperatures continue to 
increase, some beaches will experience changes in sand temperatures, 
which in turn will alter the thermal regime of incubating nests. 
Changing sand temperatures at nesting beaches may result in changing 
sex ratios of hatchling cohorts and reduced hatching output (Hawkes et 
al. 2009). Leatherback turtles exhibit temperature-dependent sex 
determination (Binckley and Spotila 2015) and warmer temperatures 
produce more female embryos (Mrosovsky et al. 1984; Hawkes et al. 
2007). In the NW Atlantic DPS, the pivotal temperature (the temperature 
at which a sex ratio of 1:1 is produced) is estimated to be between 
29.25 [deg]C and 30.5 [deg]C (Eckert et al. 2012) but there are 
variations in measurements (Girondot et al. 2018), over time, and among 
locations. An increase over that temperature would result in more 
female hatchlings. Such increases in female hatchling output have 
already been documented (Pati[ntilde]o-Mart[iacute]nez et al. 2012), 
and with an increase in temperatures from climate change, these trends 
are likely to continue if other nesting factors remain constant. For 
example, Pati[ntilde]o-Mart[iacute]nez et al. (2012) developed a model 
to relate measured incubation temperature to sex ratio and estimated 
that females nesting at Caribbean Colombian beaches currently produce 
approximately 92 percent female hatchlings. Under all future climate 
change scenarios, complete feminization could occur as soon as 2021 
(Pati[ntilde]o-Mart[iacute]nez et al. 2012). In St. Eustatius, 
leatherback hatchling production was female biased from 2002 to 2012, 
with less than approximately 24 percent of males produced every year 
(Lalo[euml] et al. 2016). Future warming air temperatures will 
exacerbate this female bias, and female leatherback sex ratios are 
projected to consistently reach 95 percent after 2028 on that island, 
which has dark and light sand beaches (Lalo[euml] et al. 2016). Warming 
trends in Costa Rica are expected to be higher than the global average 
and resulting female-biased sex ratios are also expected (Gledhill 
2007). While the assumption is that most nesting beaches will become 
female-biased due to increased sand temperatures, this may not be the 
case in all areas. In Grenada, increased rainfall (another effect of 
climate change) was found to have a cooling influence on nests, so that 
more male producing temperatures (less than 29.75 [deg]C) were found 
within the clutches (Houghton et al. 2007). Further, due to the 
tendency of nesting females to deposit some clutches in the cooler 
intertidal zone of beaches, the effects of long-term climate on sex 
ratios may be mitigated (Kamel and Mrosovsky 2004; Pati[ntilde]o-
Mart[iacute]nez et al. 2012).
    Hatching success is affected by warming temperatures. Extremely 
high sand/nest temperatures are anticipated to result in embryonic 
mortality (Gledhill 2007, Santidri[aacute]n Tomillo et al. 2012, 
Valentin-Gamazo et al. 2018). In Costa Rica, warmer conditions can

[[Page 48359]]

exacerbate the effects of biotic contamination and mold infestations of 
developing embryos (Gledhill 2007), resulting in reduced hatching 
success.
    Temperature increases are likely to be associated with more extreme 
precipitation and faster evaporation of water, leading to greater 
frequency of both very wet and very dry conditions that reduce 
productivity (Pati[ntilde]o-Mart[iacute]nez et al. 2014; 
Santidri[aacute]n Tomillo et al. 2015). These impacts may affect nests 
in different ways, but the result (e.g., reduced hatching output) is 
similar. Very wet conditions may inundate nests or increase fungal and 
mold growth, reducing hatching success (Pati[ntilde]o-Mart[iacute]nez 
et al. 2014). Very dry conditions may affect embryonic development and 
decrease hatchling output. Under climate change scenarios, very dry 
conditions are expected for St. Croix, an area already showing 
decreased productivity and reduced first time nesting female abundance 
(Santidri[aacute]n Tomillo et al. 2015; Garner et al. 2017). 
Santidri[aacute]n Tomillo et al. (2015) assessed climatic conditions on 
hatchling output at four nesting sites (Sandy Point, St. Croix; 
Pacuare, Caribbean Costa Rica; Playa Grande, Pacific Costa Rica; 
Maputaland, South Africa), and found that St. Croix had the highest 
projected warming rate (+ 5.4 [deg]C), highest absolute temperature and 
lowest precipitation levels. With these further increases in dryness 
and air temperatures, hatchling productivity is expected to be 
compromised by the end of the 21st century in this area 
(Santidri[aacute]n Tomillo et al. 2015). Santidri[aacute]n Tomillo et 
al. (2015) suggested that the lack of rain is what reduces 
developmental success and hatchling emergence. However, Rafferty et al. 
(2017) evaluated long-term climate data for St. Croix, using climate 
data collected from a nearby weather station, and found no significant 
trend in incubation temperatures or precipitation that could be 
associated with observed decreases in productivity at this location.
    Finally, incubation temperatures can also influence hatchling 
morphology and locomotion (Mickelson and Downie 2010). Leatherback 
hatchlings originating from nests incubated at lower temperatures 
exhibited carapace and front flipper length-width ratios that 
significantly improved their crawling speeds relative to those 
hatchlings incubated at high temperatures (Mickelson and Downie 2010).
    Sea level rise is another threat to leatherback turtles. Thornalley 
et al. (2018) found that the Labrador Sea deep convection and the 
Atlantic Meridional Overturning Circulation, a system of ocean currents 
in the North Atlantic, have been unusually weak over the past 150 years 
or so, and this weakened state may have modified northward ocean heat 
transport, as well as atmospheric warming by altering ocean-atmosphere 
heat transfer. Further, the documented weakening of this system is 
related to above-average sea level rise along the U.S. East Coast 
(Caesar et al. 2018). Sea level rise may result in intensified erosion 
and loss of nesting beach habitat (Fish et al. 2005; Fuentes et al. 
2010; Fonseca et al. 2013). In Bonaire, up to 32 percent of the current 
beach area could be lost with a 0.5 m rise in sea level, with lower, 
narrower beaches being the most vulnerable (Fish et al. 2005). Ussa 
(2013) predicted a 20 to 25 percent loss in beach areas due to sea 
level rise by the year 2100 within the Archie Carr National Wildlife 
Refuge, Florida, as well as areas adjacent to the Refuge. With the 
threat of increasing sea level rise, protection of developed coastlines 
often involves shoreline armoring that reduces the amount of beach 
available, thus creating a smaller amount of space for turtles to nest 
(Hawkes et al. 2009). Along such developed coastlines, rising sea 
levels may cause severe effects on eggs, because nesting females are 
forced to deposit eggs seaward of shoreline armoring, potentially 
subjecting them to repeated tidal inundation and/or egg exposure from 
exacerbated wave action near the base of these structures.
    Sea level rise is expected to result in more nests being inundated, 
reducing hatching success. On Playona Beach, Colombia, Pati[ntilde]o-
Mart[iacute]nez et al. (2014) found that nests in wet sand suffered 
higher mortality (emergence success of zero percent for wettest nests 
to 64 percent for the driest nests), suggesting that nesting success 
should be expected to decrease under future climate change sea level 
rise scenarios. Inundation is likely to reduce hatching success 
(Pati[ntilde]o-Mart[iacute]nez et al. 2008; Caut et al. 2010) and will 
continue to occur (or worsen) with sea level rise.
    However, leatherback turtles may be less susceptible than other 
species of sea turtles to loss of nesting habitat, because they exhibit 
lower nest-site fidelity (Dutton et al. 1999). Nesting beaches in the 
Guianas are already highly dynamic and interseasonally variable, and 
leatherback nesting females have been successful in those areas despite 
the fact that some beaches disappear between nesting years (Plaziat and 
Augustinus 2004; Kelle et al. 2007; Caut et al. 2010). If global 
temperatures increase and there is a range shift northwards, beaches 
not currently used for nesting could in the future become used by 
leatherback turtles, potentially offsetting some loss of accessibility 
to beaches in southern portions of the range. Leatherbacks' behavioral 
flexibility may allow for opportunities to colonize new beaches, but 
whether turtles can colonize nesting areas that become available, 
either thermally or geographically, by climate change, and whether 
these colonized areas provide incubation regimes that will lead to 
successful nesting, emergence success, and hatchling fitness cannot be 
known at this time (Hawkes et al. 2009).
    Observed changes in marine systems are associated with other 
aspects of climate change, including rising water temperatures, as well 
as related changes in ice cover, salinity, oxygen levels, and 
circulation. Ocean temperatures of the U.S. northeastern continental 
shelf and surrounding NW Atlantic waters have warmed faster than the 
global average over the last decade (Pershing et al. 2015). New 
projections for the U.S. northeastern shelf and NW Atlantic Ocean 
suggest that this region will warm two to three times faster than the 
global average and existing projections from the IPCC may be too 
conservative (Saba et al. 2015). This increase in northeastern shelf 
waters is relevant for NW Atlantic leatherback turtles, as they rely on 
U.S. and Canadian waters to forage during the warmer months (James 
2005a, 2006b, 2007; Dodge 2014, 2015).
    Global warming is expected to expand leatherback foraging habitats 
into, and increase residency time in, higher latitude waters (James et 
al. 2006a; McMahon and Hays 2006; Robinson et al. 2009). For example, 
leatherback turtles have extended their range in the Atlantic north by 
around 200 km per decade over the last two decades as warming has 
caused the northerly migration of the 15 [deg]C sea surface temperature 
(SST) isotherm, the lower limit of thermal tolerance for leatherback 
turtles (McMahon and Hays 2006). Documented weakening of the Meridional 
Overturning Circulation is related to above-average warming in the Gulf 
Stream region and an associated northward shift of the Gulf Stream 
(Caesar et al. 2018). This weakening of the deep, cold-water 
circulation in the North Atlantic is likely to continue to occur with 
global warming. Migratory routes may be altered by climate change as 
increasing ocean temperatures shift range-limiting isotherms north 
(Robinson et al. 2009). Post-nesting females from French Guiana were 
found to migrate northward toward the Gulf Stream north wall, targeting 
similar habitats in terms of physical characteristics, i.e., strong 
gradients of

[[Page 48360]]

SST, sea surface height, and a deep mixed layer (Chambault et al. 
2017). Hatchling dispersal may also be affected by changes in surface 
current and thermohaline circulation patterns (Hawkes et al. 2009; Pike 
2013).
    The effects of global warming are difficult to predict, but changes 
in reproductive behavior (e.g., remigration intervals, timing and 
length of nesting season) could occur (Hawkes et al. 2009; Hamann et 
al. 2013). Robinson et al. (2014) found that the median nesting date at 
Sandy Point (St. Croix) occurred on average 0.17 days earlier per year, 
between 1982 and 2010. However, Neeman et al. (2015) found that 
increased temperatures at the foraging grounds tend to delay 
leatherback nesting. Temperatures at the nesting beaches (Playa Grande, 
Costa Rica; Tortuguero, Costa Rica; and St. Croix) did not affect the 
timing of leatherback nesting (Neeman et al. 2015). Because the 
relation between temperatures (local sea surface and the foraging 
grounds) and timing of nesting is complex, Neeman et al. (2015) 
indicated that further study is needed at the nesting beaches to 
determine how environmental conditions change within the season and how 
these changes affect nesting success. Robinson et al. (2014) suggests 
that shifts in the nesting phenology may make the Atlantic populations 
more resilient to climate change.
    Extreme precipitation events over most of the mid-latitude and 
tropical regions will very likely become more intense and more frequent 
(IPCC 2014). Changes in the frequency and timing of storms or changes 
in prevailing currents could lead to increased beach loss via erosion 
(Van Houtan and Bass 2007; Fuentes and Abbs 2010). More frequent and 
intense storm events will have the same effect on leatherback nesting 
success as previously described for natural disasters.
    In summary, climate change is likely to affect multiple life stages 
of turtles in the NW Atlantic DPS. Likely impacts include altering sex 
ratios and reducing nest success, reducing nesting beach habitat and 
nests due to sea level rise and storms, and potentially changing 
distribution. Climate change therefore has the potential to alter 
productivity and diversity. These impacts could be more severe in 
certain areas with more dynamic beach environments, or could be 
widespread throughout the DPS. Impacts are likely to range from small, 
temporal changes in nesting season to large losses of productivity. 
That said, leatherback turtles are considered to be the best able to 
cope with climate change of all sea turtle species due to their wide 
geographic distribution and relatively weak nesting site fidelity. 
Overall, we conclude that climate change is a threat to the NW Atlantic 
DPS.

Conservation Efforts

    Next we consider ``conservation efforts'' under Section 4(b)(1)(A) 
(16 U.S.C. 1533(b)(1)(A)).\1\ There are numerous efforts to conserve 
the leatherback turtle. The following conservation efforts apply to the 
NW Atlantic DPS (for a description of each effort, please see the 
section on conservation efforts for the taxonomic species): African 
Convention on the Conservation of Nature and Natural Resources (Algiers 
Convention); Central American Regional Network; Convention on the 
Conservation of Migratory Species of Wild Animals; Convention on 
Biological Diversity; Convention on International Trade in Endangered 
Species of Wild Fauna and Flora; Convention Concerning the Protection 
of the World Cultural and Natural Heritage (World Heritage Convention); 
Convention for the Protection and Development of the Marine Environment 
of the Wider Caribbean Region, Specially Protected Areas and Wildlife 
(SPAW); Convention on the Conservation of European Wildlife and Natural 
Habitats; Convention for the Co-operation in the Protection and 
Development of the Marine and Coastal Environment of the West and 
Central African Region (Abidjan Convention); Memorandum of 
Understanding Concerning Conservation Measures for Marine Turtles of 
the Atlantic Coast of Africa (Abidjan Memorandum); Convention for the 
Protection and Development of the Marine Environment of the North East 
Atlantic; Convention on Nature Protection and Wildlife Preservation in 
the Western Hemisphere (Washington or Western Hemisphere Convention); 
Convention for the Protection and Development of the Marine Environment 
of the Wider Caribbean Region (Cartagena Convention); Cooperative 
Agreement for the Conservation of Sea Turtles of the Caribbean Coast of 
Costa Rica, Nicaragua, and Panama (Tri-Partite Agreement); Council 
Regulation (EC) No. 1239/98 of 8 June 1998 Amending Regulation (EC) No. 
894/97 Laying Down Certain Technical Measures for the Conservation of 
Fishery Measures (Council of the European Union); Council Directive 92/
43/EEC on the Conservation of Natural Habitats and of Wild Fauna and 
Flora (EC Habitats Directive); Food and Agricultural Organization (FAO) 
Technical Consultation on Sea Turtle-Fishery Interactions; Inter-
American Convention for the Protection and Conservation of Sea Turtles 
(IAC); MARPOL; Inter-American Tropical Tuna Convention (IATTC); IUCN; 
North American Agreement for Environmental Cooperation; Protocol 
Concerning Specially Protected Areas and Biological Diversity in the 
Mediterranean; Ramsar Convention on Wetlands; Regional Fishery 
Management Organizations (RFMOs); UN Convention on the Law of the Sea 
(UNCLOS); and UN Resolution 44/225 on Large-Scale Pelagic Driftnet 
Fishing. Although numerous conservation efforts apply to the turtles of 
this DPS, they do not adequately reduce its risk of extinction.
---------------------------------------------------------------------------

    \1\ For a related discussion of existing regulatory mechanisms 
to protect turtles, which are considered separately under Section 
4(a)(1)(D), see the discussion above at ``Inadequacy of Existing 
Regulatory Mechanisms.''
---------------------------------------------------------------------------

Extinction Risk Analysis

    After reviewing the best available information, the Team concluded 
that the NW Atlantic DPS is at high risk of extinction. The total index 
of nesting female abundance is 20,659 females at consistently monitored 
beaches, and the most recent annual rate of decline is estimated to be 
approximately nine percent (NW Atlantic Leatherback Working Group 
2018). The best available nest data reflect a steady decline for more 
than a decade, becoming more pronounced since 2008 (Eckert and Mitchell 
2018; NW Atlantic Leatherback Working Group 2018). This decreasing 
trend is observed when all available nest data are combined and at most 
nesting beaches (NW Atlantic Leatherback Working Group 2018), including 
the largest nesting aggregation in Trinidad (i.e., Grande Riviere, 
which is declining at 6.9 percent annually). In terms of productivity, 
the DPS exhibits low hatching success, while other key parameters such 
as clutch size, remigration interval, and clutch frequency are similar 
to species' averages. There are also indications of decreased 
productivity within the DPS at one of the most intensively monitored 
nesting beaches (i.e., Sandy Point, St. Croix; Garner et al. 2017). The 
declining region-wide nest trend and potential changes in productivity 
make the DPS highly vulnerable to threats.
    However, the DPS exhibits broad spatial distribution and some 
diversity. Based upon genetic data, as well as flipper tagging and 
satellite telemetry data, this DPS shows significant spatial structure 
with some connectivity among nesting and foraging areas. Further, 
nesting occurs in a variety of habitats,

[[Page 48361]]

including islands and mainland, as well as muddy, sandy, and shelly 
beaches. The DPS uses multiple, distant, and diverse foraging areas, 
including oceanic and coastal waters throughout the North Atlantic 
Ocean, Mediterranean Sea, and GOM, providing some resilience against 
reduced prey availability. While the numerous and diverse nesting and 
foraging locations, along with moderate levels of genetic diversity, 
provide some level of buffer to the DPS, the highest concentrations of 
nesting occur in Trinidad, French Guiana, and Panama, where a 
catastrophic event could have a disproportionate impact on the DPS.
    The primary threat to the NW Atlantic DPS is bycatch in commercial 
and artisanal, pelagic and coastal fisheries. Gillnet fisheries, in 
particular those off nesting beaches, are the greatest concern given 
the high mortality rate. In particular, the coastal surface drift 
gillnet fishery off Trinidad kills an estimated 1,000 adult leatherback 
turtles annually (Lee Lum 2006; Eckert et al. 2008; Eckert 2013). 
Bycatch, and subsequent mortality, in Trinidad bottom set gillnets and 
surface gillnets in Suriname and French Guiana are major threats to the 
NW Atlantic DPS. Trinidad and French Guiana host the highest number of 
nesting females in this DPS, so the continued mortality of adults in 
that area is of significant concern. Further, no adequate regulatory 
mechanism is currently in place (e.g., no gear modifications or 
closures) to address this incidental bycatch. These fisheries and the 
related mortality rates have been occurring for years (Lee Lum 2006; 
Eckert 2013). Longline fisheries are the most widespread threat, 
occurring throughout the Atlantic Ocean by fisheries from multiple 
nations, incidentally capturing thousands of leatherback turtles 
annually based on the best available data. Longline gear modifications 
(e.g., circle hooks) are sometimes, but not consistently, used. Fishery 
bycatch in pot/trap gear, especially off the northeastern U.S. coast 
and in Canadian waters, and trawls are also significant threats. 
Fisheries bycatch reduces abundance by removing individuals from the 
population; when those individuals are nesting females, it reduces 
productivity as well. Given the lack of observer coverage and 
reporting, cumulative mortality due to fisheries bycatch is likely 
higher than available estimates.
    Additional threats to the DPS include habitat loss, the legal and 
illegal harvest of turtles and eggs, predation, vessel strikes, 
pollution, climate change, oil and gas activities, and natural 
disasters. Coastal development and armoring, erosion (natural and 
anthropogenic), and artificial lighting are some of the most 
significant stressors on nesting beach habitat, reducing nesting and 
hatching success (i.e., productivity). Habitat loss and modification is 
also anticipated to increase over time with additional development and 
climate change. Legal and illegal harvest of turtles and eggs reduces 
abundance and productivity. Illegal egg poaching occurs in several 
nations, particularly Costa Rica, Dominican Republic, and Colombia. 
While reduced in some nations, illegal poaching still occurs on 
unmonitored beaches throughout most of the Caribbean, including 
Suriname and Trinidad. While leatherback eggs and hatchlings are preyed 
upon by many species, the biggest threat is from feral dogs. Egg 
predation by dogs occurs in many nations, but it is a particular 
concern in Colombia, French Guiana, Guyana, Panama, Puerto Rico, and 
Trinidad and Tobago. Intervention (e.g., screening) to reduce predation 
is not used in most places, partially due to the concern of attracting 
poachers as well as the infeasibility of implementing effective 
measures at high-density or remote beaches. Egg predation reduces 
productivity.
    Vessel strikes are also a threat, killing numerous leatherback 
turtles each year. While exposure to vessel strikes may be most severe 
in developed areas, the total impacts are high, affecting both 
abundance and productivity. Pollution, ingestion of plastics, and 
entanglement in marine debris are threats to all leatherback turtles, 
most likely resulting in injury and compromised health, and sometimes 
mortality. Exposure to pollution is widespread in the NW Atlantic 
Ocean, but effect data are limited. Oil and gas activities are threats 
with the potential to grow in some Caribbean areas. Natural disasters 
(hurricanes) and phenomenon (large Sargassum events) have an 
intermittent impact to the NW Atlantic DPS. Climate change is likely to 
result in reduced productivity due to greater rates of coastal erosion 
and sea level rise and subsequent nest inundation and habitat loss, 
reduced hatching success, changing sex ratios, and distributional 
changes. Although many international, national, and local regulatory 
mechanisms are in place, they do not adequately reduce the impact of 
these threats.
    The cumulative impact of these multiple threats is potentially 
large (Andersen et al. 2017). Innis et al. (2010) reported that many 
individuals are simultaneously exposed to multiple threats, including: 
entanglement, injury, plastic ingestion, adrenal gland parasitism, 
diverticulitis, and burdens of environmental toxins (Innis et al. 
2010). Such cumulative pressures affect individual survival and 
productivity. In some cases, it is possible to directly link individual 
threats to demographic reductions (e.g., high mortality in gillnets off 
nesting beaches reduces nesting female abundance). More often, however, 
several threats contribute synergistically to demographic reductions. 
For example, reductions in hatching success may be caused by one or 
more of the following threats alone or in combination: erosion, 
poaching, predation, climate change, and pollution.
    We find that the NW Atlantic DPS is affected by numerous severe 
threats. These present, ongoing threats injure or kill turtles and 
contribute to the declining nest trend. The Team evaluated whether the 
DPS is at risk of extinction currently or would become so within the 
foreseeable future. To answer this question, they asked how long it 
would take for the total index of nesting female abundance to be 
reduced by 50 percent, a drastic decline that would reduce abundance to 
a level where demographic risks would present an independent threat to 
the DPS's continued existence, and whether this time period places the 
DPS at risk currently or within the foreseeable future. Using estimates 
of the mean time to maturation for the population (approximately 19 
years; Avens et al. in review) and mean nesting longevity 
(approximately 11 years; Avens et al. in review) of the species, they 
estimated a generation time of approximately 30 years. They then 
considered three different scenarios. First, they calculated the time 
until 50 percent reduction in the total index of nesting female 
abundance using data on a significant and influential, well-documented, 
threat: Gillnet bycatch mortality of 1,000 adult turtles annually off 
the largest nesting aggregation, i.e., Trinidad. Assuming that half of 
the turtles at Trinidad killed are female, total index of nesting 
female abundance would decrease by 50 percent in 28 years, which is 
approximately one generation.
    Second, the Team used regional nest trend data from the NW Atlantic 
Leatherback Working Group (2018). Using the most recent trends as is 
typical for population projections (i.e., -9.32 percent per year from 
2008 to 2017), they found that the total index of nesting female 
abundance would fall by 50 percent within 8 years (95 percent CI: 6 to 
13 years). Using trends from the longer data set (-4.21 percent per 
year

[[Page 48362]]

from 1990 to 2017), the total index of nesting female abundance would 
fall by 50 percent within 17 years (95 percent CI: 11 to 31 years). 
Finally, using their calculation of nest trend for the highest 
abundance nesting area in the DPS, Trinidad (-7.3 percent per year, 95 
percent CI: -34 to 18 percent), the Team found that the total index of 
nesting female abundance would decrease by 50 percent within 10 years 
(95 percent CI: 3 years to ``never;'' however, ``never'' is highly 
unlikely, given that there is a 75 percent likelihood that the true 
value of the nest trend in Trinidad is negative (f = 0.754)). There are 
several caveats with using nest trend data: Adult females typically 
account for, at most, a small percentage of the population; trends in 
nesting female abundance may not be an index of the remainder of 
population; stable age distribution is assumed; and time series of 
available data do not always span one generation (let alone multiple 
generations required to reach stable age distribution). Despite these 
caveats, all scenarios resulted in a 50 percent reduction in the total 
index of nesting female abundance in less than one generation. While 
the first scenario did not involve the use of nest trend data, it did 
result in a 50 percent reduction within one generation when considering 
only one threat (albeit the most severe), and we know that the DPS 
faces many large-impact threats, (suggesting that the first scenario 
understates the DPS's degree of risk).
    For the purpose of the extinction risk analysis, the Team discussed 
whether the resulting range of time periods (8 to 28 years) suggests a 
present risk of extinction or a risk of extinction within the 
foreseeable future. The Team did not have a unanimous view. All but one 
Team member were present to vote on the level of extinction risk. Eight 
Team members concluded with moderate confidence that the DPS is at high 
extinction risk due to threats and the declining trend that has 
accelerated in recent years. Their confidence was moderate rather than 
high due to some resilience provided by the abundance, spatial 
distribution, and diversity for this DPS. Two Team members concluded 
with low confidence that the DPS is at moderate extinction risk. Their 
confidence in this conclusion was low due to the declining trend that 
has accelerated in recent years. The Terms of Reference called for a 
simple majority, and after voting, the Team concluded that the DPS met 
the definition for high risk of extinction. We agree with the Team's 
overall conclusion that a 50 percent decline in less than one 
generation equates to a current, high risk of extinction. We find 
support for this conclusion in well documented examples of other 
leatherback populations that have quickly declined despite larger 
abundances (e.g., the Mexico nesting aggregation declined from 70,000 
nesting females in 1982 to under 1,000 nesting females by 1994; Spotila 
et al. 2000).
    We conclude that the NW Atlantic DPS is presently in danger of 
extinction due to the number and magnitude of threats, of which 
fisheries bycatch is the greatest concern. These present and ongoing 
threats have resulted in imminent and substantial demographic risks 
(i.e., declining trends and reduced abundance). Although numerous 
conservation efforts apply to the turtles of this DPS, they do not 
adequately reduce the risk of extinction. We conclude that the NW 
Atlantic DPS is in danger of extinction throughout its range and 
therefore meets the definition of an endangered species. The threatened 
species definition does not apply because the DPS is currently at risk 
of extinction (i.e., at present), rather than on a trajectory to become 
so within the foreseeable future.

SW Atlantic DPS

    The Team defined the SW Atlantic DPS as leatherback turtles 
originating from the SW Atlantic Ocean, north of 47[deg] S, east of 
South America, and west of 20[deg] W; the northern boundary is a 
diagonal line between 5.377[deg] S, 35.321[deg] W and 12.084620[deg] N, 
20[deg] W. The southern boundary is based on the Antarctic circumpolar 
current which prevents sea turtles from nesting further south. The 
western end of the northern boundary is based at the ``elbow'' of the 
Brazilian coast, where the Brazilian Current begins and likely 
restricts the northern nesting range of this DPS. We placed the eastern 
boundary at the 20[deg] W meridian as an approximate midpoint between 
SW Atlantic and SE Atlantic (i.e., turtles that nest in western Africa) 
nesting beaches and to reflect both DPS's wide foraging range 
throughout the South Atlantic Ocean. However, due to its low abundance, 
the SW Atlantic DPS is less likely to be encountered compared to the 
more abundant SE Atlantic DPS.
    The SW Atlantic DPS only nests on the southeastern coast of Brazil, 
primarily in the state of Esp[iacute]rito Santo, on a continuous 
stretch of beach, less than 100 km in length, with concentrated nesting 
in Povoa[ccedil][atilde]o and Comboios. While there is occasional, 
limited nesting south of these primary nesting beaches, the sand 
becomes coarser further south, and the excavation of nests becomes more 
difficult because the sand falls back into the holes (Thom[eacute] et 
al. 2007).
    While nesting is limited geographically, the overall range of this 
DPS (i.e., all areas of occurrence) is extensive, as demonstrated by 
individuals tracked to numerous foraging areas. Leatherback turtles of 
this DPS use coastal waters off South America from the ``elbow'' of 
Brazil southwards to Uruguay and Argentina, where quality foraging 
areas allow for coastal foraging in addition to open-ocean foraging 
(Almeida et al. 2011). Individuals of this DPS are also known to 
migrate to the waters off western Africa and forage in the oceanic 
habitat in between South America and Africa (Almeida et al. 2011). 
Likewise, Prosdocimi et al. (2014) found 84 to 86 percent of 
leatherback turtles sampled from the foraging grounds off Argentina and 
Eleva[ccedil][atilde]o do Rio Grande (an elevated offshore area across 
from Brazil) to originate from western African beaches.

Abundance

    The total index of nesting female abundance for the SW Atlantic DPS 
is 27 females. We based this index on nest monitoring data from Projeto 
TAMAR, the Brazilian Sea Turtle Conservation Program, which has 
established an index nesting survey area along 47 km of beach (10 km 
along Povoa[ccedil][atilde]o and 37 km along Comboios; IAC Brazil 
Annual Report 2018), where complete daily surveys have been conducted 
during the primary nesting season from September through March, since 
the 1986/1987 nesting season. Some nesting occurs along the non-index 
stretches of Povoa[ccedil][atilde]o and the beaches to the northern 
part of the area, but it is minor relative to nesting on the index 
survey area (Thom[eacute] et al. 2007). To calculate the index of 
nesting female abundance (i.e., 27 nesting females) for the 
Esp[iacute]rito Santo index area, we divided the total number of nests 
between the 2014/2015 and 2016/2017 nesting seasons (i.e., a 3-year 
remigration interval; Thom[eacute] et al. 2007) by the clutch frequency 
(5 clutches/season; Thom[eacute] et al. 2007; Tiwari et al. 2013).
    Minimal, scattered nesting has been reported on beaches outside 
Esp[iacute]rito Santo (Barata and Fabiano 2002; Thom[eacute] et al. 
2007; Bezerra et al. 2014), but these beaches exhibit suboptimal sand 
characteristics for nesting, limiting the possibility of substantial 
nesting expansion into those areas (Thom[eacute] et al. 2007). 
Therefore, while the nest counts from the index beach surveys do not 
provide a full estimate of all nesting for

[[Page 48363]]

the DPS, they provide a high-quality dataset, account for the majority 
of the nests (approximately 80 percent; Colman et al. 2019), and are 
used for determining our index of nesting female abundance and the nest 
trend in the next section.
    Our total index of nesting female abundance is similar to the IUCN 
Red List assessment's estimate of 35 mature individuals (female and 
male, assuming a 3:1 sex ratio) based on nesting data through 2010 
(Tiwari et al. 2013).
    The total index of nesting female abundance (i.e., 27 nesting 
females at the index beach) places the DPS at risk for environmental 
variation, genetic complications, demographic stochasticity, negative 
ecological feedback, and catastrophes (McElhany et al. 2000; NMFS 
2017). These processes, working alone or in concert, place small 
populations at a greater extinction risk than large populations, which 
are better able to absorb losses in individuals. Due to its small size, 
the DPS has limited capacity to buffer such losses. Given the intrinsic 
problems of small population size, we conclude that the nesting female 
abundance is a major factor in the extinction risk of the SW Atlantic 
DPS.

Productivity

    The SW Atlantic DPS exhibits an increasing, although variable nest 
trend. Long-term monitoring data for this small DPS are limited to the 
index nesting survey area in southeastern Brazil, where data has been 
collected between the 1986/1987 and 2016/2017 nesting seasons. Over the 
31-year data collection period, the mean annual number of nests for 
these beaches was 35. While this is below the criterion of 50 annual 
nests for conducting a trend analysis, we determined that this site 
should nevertheless be included due to the high quality and consistency 
of the data, and the fact that these data accurately represent the low 
level of nesting of this DPS. The median increase in nest counts was 
4.8 percent annually (sd = 5.8 percent; 95 percent CI = -8.4 to 15.5 
percent; f = 0.832; mean annual nests = 35). As the index area hosts 
the majority of known nesting activity, these data are representative 
of the entire DPS. We conclude that nesting has increased from 1986 to 
2017. Our trend estimate is similar to that of the IUCN Red List 
assessment, which characterizes the population as increasing (Tiwari et 
al. 2013). It is also in agreement with the recent study by Colman et 
al. (2019), which describes the trend as increasing but variable, with 
the mean annual number of nests increasing from 25.6 nests in the first 
5 years to 89.8 nests in the last 5 years of monitoring (between 1988 
and 2017).
    While the long term trend indicates an increase in nesting, the 
most recent 3 years of data (i.e., 30, 64, and 38 nests from 2014 to 
2016) show a marked reduction in nests compared to the previous 3 years 
(i.e., 78, 124, and 102 nests from 2011 to 2013). The reason for this 
reduction is unknown. It could reflect declining nesting female 
abundance or changes in productivity metrics (i.e., a longer 
remigration interval or reduced clutch frequency) related to 
environmental shifts or prey availability. Therefore, there is 
uncertainty regarding whether the increasing trend will continue.
    The productivity parameters for this DPS are fairly typical for the 
species. In Brazil, the average clutch size appears to be on the lower 
end of the range for Atlantic populations; conversely, Brazilian nests 
tend to have a higher average number and percentage of eggs per clutch 
(Thom[eacute] et al. 2007). Therefore, the egg production of this DPS 
appears to be weighed more towards production of viable, hatchling-
producing eggs compared to other Atlantic populations (Thom[eacute] et 
al. 2007). Nesting females produced an average of 3,496 hatchlings 
annually over the past 10 years of nesting, which was calculated by 
multiplying 60.4 nests annually, 87.7 eggs per nest, and 66.0 percent 
hatching success (Colman et al. 2019). This estimate does not include 
the limited nesting outside the index area. The mean size of nesting 
females (CCL) has changed from 159.8 cm, with a range of 139 to 182 cm 
(Thom[eacute] et al. 2007) to 152.9 cm  10.0 SD, with a 
range of 124.7 to 182.0 cm; the decrease was statistically significant 
and may indicate recruitment (Colman et al. 2019). Hatching success has 
increased from a mean of 65.1 percent (with a range of 53.3 to 78 
percent; Thom[eacute] et al. 2007) to a mean of 66 percent (with a 
range of 38.8 to 82.4 percent; Colman et al. 2019).
    While the overall nest trend for this DPS is increasing, there is 
uncertainty regarding the continuation of this trend, given the data 
for the past 3 years. The population remains extremely small, and thus 
overall productivity is limited. Additionally, the potential for 
population growth is not clear, given the limited suitable nesting 
habitat available. We conclude that limited productivity places the DPS 
at risk of extinction.

Spatial Distribution

    The SW Atlantic DPS comprises a single, small nesting aggregation 
concentrated on the beaches of one state in Brazil (Esp[iacute]rito 
Santo). A tagging study has shown internesting movements along 300 km 
of the coast, including over 100 km on either side of known nesting 
beaches (Almeida et al. 2011), indicating connectivity throughout this 
area. The nesting spatial distribution is extremely restricted, with 
nesting constrained to a small area, with little suitable nesting 
habitat into which it can expand. Conversely, the DPS exhibits a broad 
foraging range, extending south to waters off Uruguay and Argentina, 
throughout the pelagic waters of the South Atlantic, and across to 
western Africa (Almeida et al. 2011).
    The wide distribution of foraging areas likely provides some level 
of buffer for the DPS against local catastrophes or environmental 
changes that could limit prey availability. However, the limited 
nesting range, and apparent lack of suitable nesting beaches into which 
to expand, renders the DPS highly susceptible to detrimental 
environmental impacts, both acute (e.g., storms and singular events) 
and chronic (e.g., sea level rise and temperature changes). Any such 
change would impact the entire extent of the DPS's nesting habitat. 
With no metapopulation structure, the DPS has reduced capacity to 
withstand other catastrophic events. Thus, despite widely distributed 
foraging areas, the extremely narrow nesting distribution and lack of 
population structure increases the extinction risk of the SW Atlantic 
DPS.

Diversity

    Despite its extremely low nesting female abundance, the Brazilian 
nesting aggregation has the second-highest haplotype diversity among 
all Atlantic populations (h = 0.498-0.532; Dutton et al. 2013; Vargas 
et al. 2017). According to Thom[eacute] et al. (2007), while most 
nesting occurs from September through March, sporadic nesting has been 
recorded throughout the year, which may provide temporal resilience if 
environmental conditions limit nesting during the primary nesting 
season. The use of estuarine waters (of the Rio de la Plata) as a year-
round foraging ground is an unusual characteristic shared with the SE 
Atlantic DPS (Lopez-Mendilaharsu et al. 2009; Prosdocimi et al. 2014). 
Despite genetic and foraging diversity, the limited size and range of 
the nesting aggregation reduces the resilience of this DPS.

[[Page 48364]]

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    Within the limited nesting range of the SW Atlantic DPS, habitat 
modification is a threat. The 2015 collapse of a tailings dam at an ore 
mine upstream of the index nesting survey area had an undetermined, but 
potentially long-term, impact on the nesting beach of the DPS. Tens of 
millions of cubic meters of heavy metal-laden mining waste entered the 
Doce River and ultimately passed through the mouth of the river, in the 
middle of the index nesting area. Nests laid near the river mouth were 
relocated to prevent hatchlings from entering polluted waters. Hatching 
success was not significantly different between years in the period of 
2012 to 2017, which include three seasons before (2012-2014) and three 
seasons after (2015-2017) the mining event (Colman et al. 2019). While 
no difference was noted in the distribution of nests following the dam 
breach, non-lethal impacts to individuals encountering the polluted 
waters, especially hatchlings, could not be measured. Such impacts may 
have occurred but may not be evident for decades following the spill. 
Projeto TAMAR is monitoring for heavy metals in eggs and nesting 
females and is closely watching for changes in fitness and reproductive 
parameters (Thom[eacute] et al. 2017). As a result of the dam's 
collapse, the Brazilian Federal government is implementing a marine 
protected area (APA-Area de Protecao Ambiental da Foz do Rio Doce), 
including about 100 kilometers of coastline, which should encompass the 
entire extension of the index nesting beaches, with both coastline and 
surrounding marine areas. Such a measure is an environmental 
compensation for the dam's collapse, and should be implemented with 
specific resources in the coming years (ICMBio, MMA, Brazil; J. 
Thom[eacute], Projeto TAMAR, pers. comm., 2019).
    Beach erosion and tidal flooding are also threats to this DPS. 
According to Thom[eacute] et al. (2007), occasional relocation of nests 
and nest protection occur when inundation or predation risk is 
considered high. The majority of nests are relocated when in danger of 
beach erosion or tidal flooding (J. Thom[eacute], Projeto TAMAR, pers. 
comm., 2019).
    Although coastal light pollution has been documented to be 
increasing in Brazil, nesting has not been notably impacted thus far 
(Colman et al. 2018). The lack of impact may be attributable to 
conservation strategies including the creation of protected areas and 
minimization of direct lighting on the nesting beaches. Nests are 
relocated from heavily lit areas. All light sources with a light 
intensity greater than 0 lux (lux = lumen per m\2\) on these beaches 
are prohibited by a Federal ordinance (Portaria IBAMA 11/1995). 
Construction, lighting, and poaching were not considered a significant 
problem at the leatherback nesting beaches by Thom[eacute] et al. 
(2007). However, such problems persist in several other turtle nesting 
beaches in Brazil (Mascarenhas et al. 2004; Lara et al. 2016). More 
recently, coastal development and artificial lighting have been 
identified as potential threats for leatherback turtles on the beaches 
of Esp[iacute]rito Santo (TAMAR/Unpublished data) and further research 
is needed to better understand these threats. Nests are relocated from 
heavily lit areas. Colman et al. (2018) found a negative relationship 
between nest density and light levels. Additionally, as oil industry 
and other economic developments are explored, the potential threat to 
the nesting habitat may increase (Thom[eacute] et al. 2007).
    A significant portion of the nesting beach is protected as a 
Federal reserve under Brazilian Decree no. 90222 (September, 25 1984), 
which covers 15 km of Comboios Beach, south of the mouth of the Doce 
River. An additional 22 km, south of the reserve, falls within 
indigenous land that has restricted access under Federal law. No 
Federally protected areas exist north of the Doce River mouth, where 
Povoa[ccedil][atilde]o Beach occurs. However, local, state, and Federal 
regulations provide some coastal zone protections in that area.

Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    Overutilization poses a threat to the SW Atlantic DPS. Though 
specific information on leatherback turtle harvests is not available, 
there was historically traditional harvest of sea turtles and eggs in 
Esp[iacute]rito Santo (Hartt 1941; Medeiros 1983). This harvest, 
however, has been largely curtailed through the work of Projeto TAMAR, 
which promoted other economic activities and hired ex-turtle hunters to 
protect nests (Marcovaldi et al. 2005; Almeida and Mendes 2007). The 
capture of leatherback turtles was banned in Brazil in 1968, and full 
protection for all sea turtles was enacted in 1986 (Marcovaldi and 
Marcovaldi 1999). At present, egg poaching has been reduced, and there 
is no known subsistence hunting for sea turtles of any species 
(Thom[eacute] et al. 2007). As previously noted, there is protection 
for or limited access to much of the nesting habitat south of the Doce 
River. However, this protection does not extend north of the river, 
where additional nesting occurs. Because of the very small size of the 
population, even very low levels of egg poaching have the potential to 
impact the population. Therefore, we conclude that overutilization 
poses a threat to the SW Atlantic DPS.

Disease or Predation

    While we could not find any information on disease for this DPS, 
predation is a threat to the SW Atlantic DPS. Invertebrates, reptiles, 
and mammals prey on eggs, while hatchlings fall prey to land, air, and 
marine predators. According to Thom[eacute] et al. (2007), relocation 
and protection of nests may be undertaken when inundation (primarily) 
or predation (secondarily) risk is considered high (J. Thom[eacute], 
Projeto TAMAR, pers. comm., 2019). Predators include foxes (Cerdocyon 
thous), raccoons (Procyon cancrivorus), and domestic dogs, although 
there are no quantitative estimates of predation for this DPS (J. 
Thom[eacute], Projeto TAMAR, pers. comm., 2019). Some predation of 
large juveniles and adults occurs in the marine environment, especially 
by sharks (Bornatowski et al. 2012), but the frequency and impact on 
those populations is not well understood. For this DPS, predation 
primarily impacts productivity (i.e., reduced egg and hatching 
success). We conclude that predation is a threat to the SW Atlantic 
DPS, but that there is insufficient information to classify disease as 
a threat.

Inadequacy of Existing Regulatory Mechanisms

    The SW Atlantic DPS is protected by several regulatory mechanisms. 
For each, the Team reviewed the objectives of the regulation and to 
what extent it adequately addresses the targeted threat.
    Beach habitat is protected throughout much of the nesting range of 
this DPS. The vast majority of nesting occurs in Esp[iacute]rito Santo, 
where beaches have been protected since 1982. All light sources with a 
light intensity greater than 0 lux (lux = lumen per m\2\) on these 
beaches are prohibited by a Federal ordinance (Portaria IBAMA 11/1995).
    The take of leatherback turtles is illegal throughout the SW 
Atlantic Ocean. Regional regulations include: Brazil Portaria, Manter 
proibida a captura de tartarugas marinhas das esp[eacute]cies Caretta, 
Dermochelys coriacea, Eretmochelys imbricata e Lepidochelys

[[Page 48365]]

olivacea \2\ No.27/1982; Uruguay Presidential Decree 144 and additional 
legislation to reduce bycatch and prevent habitat alteration, and to 
prevent the removal of individuals from their natural environment; 
Argentina National Decree 666 from 1997; and various laws prohibiting 
hunting and selling sea turtles. Harvest and consumption of sea turtles 
are illegal under Brazilian law (Law on Environmental Crimes N[deg] 
9605/1998). While these protections are mostly effective, very low 
levels of egg poaching still exist (Thom[eacute] et al. 2007).
---------------------------------------------------------------------------

    \2\ Prohibition of the capture of sea turtles of the species 
Caretta caretta, Dermochelys coriacea, Eretmochelys imbricata, and 
Lepidochelys olivacea.
---------------------------------------------------------------------------

    Fisheries bycatch is the primary threat to the SW Atlantic DPS. 
Although regulations address this issue to some extent, they do not do 
so adequately and it continues to be a threat. In 2001, Brazil 
established the National Plan for the Reduction of Incidental Capture 
of Sea Turtles in Fishing Activities (Marcovaldi et al. 2005). However, 
bycatch continues to be a major problem. In Brazil, the use of TEDs in 
trawl fisheries is mandatory (Instru[ccedil][atilde]o Normativa MMA No. 
31; December 13, 2004), but most fishermen nevertheless do not use such 
gear, and there is little or no enforcement by authorities (IAC Brazil 
Annual Report 2018). The UN established a worldwide moratorium on drift 
gillnet fishing effective in 1992, the General Fisheries Commission for 
the Mediterranean prohibited driftnet fishing in 1997, and the 
International Commission for the Conservation of Atlantic Tunas (ICCAT) 
banned driftnets in 2003. Despite these and other numerous regulations 
and international instruments to protect sea turtles, significant 
bycatch still occurs in artisanal and commercial fisheries operating in 
the territorial waters of Argentina, Uruguay, and Brazil and on the 
high seas (Gonz[aacute]lez et al. 2012).
    In summary, while numerous regulatory mechanisms have been enacted 
to provide some protections to leatherback turtles, their eggs, and 
nesting habitat throughout the range of this DPS, they have been 
inadequate. Many do not effectively reduce the threat that they were 
designed to address, generally as a result of limited implementation or 
enforcement. Fisheries bycatch, in particular, remains a major threat 
to the DPS despite regulatory mechanisms. We conclude that the failure 
to implement and enforce effective regulations is a threat to the DPS.

Fisheries Bycatch

    Fisheries bycatch is the primary threat to the SW Atlantic DPS. 
Leatherback turtles are captured as bycatch in commercial and artisanal 
fisheries, along coastal foraging and breeding areas, and on the high 
seas. The extensive foraging range of this DPS makes it vulnerable to 
interactions with fisheries off the coasts of Brazil, Uruguay, and 
Argentina, in the pelagic waters of the South Atlantic Ocean, and along 
the coastal waters off western Africa. Recoveries of females tagged in 
Esp[iacute]rito Santo are scarce, however. Three were found dead on the 
Brazilian coast (incidentally captured in fisheries around the Doce 
River mouth (TAMAR, unpublished data)), one in Argentina (Alvarez et 
al. 2009), and one in Namibia, West Africa (Almeida et al. 2014). 
Fisheries interaction information specific to this DPS is limited, 
because the data do not differentiate among individuals from this DPS 
and SE Atlantic individuals that forage within the same range. Because 
the SE Atlantic DPS is much more abundant than the SW Atlantic DPS, 
most fishery interactions likely involve SE Atlantic individuals. 
However, data about bycatch involving the SE Atlantic DPS is 
informative because the impact to the SW Atlantic DPS individuals is 
likely to be proportional to their relative presence in the area. 
Bycatch in gillnets; surface, deep-water longline hooks; and trawls are 
the principal causes of sea turtle deaths, with not only higher 
interaction numbers, but higher mortality rates than other fishery 
interactions (Kotas et al. 2004; Pinedo and Polacheck 2004; Tudela et 
al. 2005; Giffoni et al. 2013).
    Coastal gillnet fisheries interactions are one of the largest 
threats to the survival of the SW Atlantic DPS. In an analysis of 
Brazilian fishery data from 1990 to 2012, Giffoni et al. (2013) 
documented 237 leatherback turtle interactions, and 31 percent 
mortality, in coastal set, fixed, encircling, and pelagic drift 
gillnets. The actual number of interactions is likely substantially 
higher, as many interactions go unreported.
    Smaller scale artisanal gillnet fisheries occur in coastal waters 
that are used by SW Atlantic individuals for mating, access to nesting 
beaches, and foraging. Thom[eacute] et al. (2007) described the 
occurrence of artisanal gillnet fisheries close to the nesting beach 
but indicated that Brazil was investing resources in developing lower-
impact fishing techniques. However, such fisheries occur throughout 
other important coastal foraging areas off South America. Additionally, 
coastal artisanal gillnet fishery interactions with leatherback turtles 
are known to occur off the western coast of Africa, where some 
individuals from the SW Atlantic DPS forage (Riskas and Tiwari 2013). 
The Rio de la Plata estuary, an important foraging area off Uruguay, 
has numerous documented instances of leatherback turtle entanglements, 
including mortalities from coastal bottom-set gillnet fisheries 
(Fallabrino et al. 2006; Lopez-Mendilaharsu et al. 2009; Velez-Rubio et 
al. 2013).
    Larger-scale commercial ocean gillnet fisheries are also a 
significant threat for the SW Atlantic DPS, with high bycatch rates 
reported off Brazil in drift and set gillnets (Fiedler et al. 2012; 
Ramos and Vasconcellos 2013). Drift gillnet fishing off Brazil started 
in 1986, targeting hammerhead sharks (Domingo et al. 2006). Marcovaldi 
et al. (2006) reported that leatherback turtles comprised about 70 
percent of all sea turtles captured in Brazilian driftnet shark 
fisheries. From 2002 to 2008, 351 sea turtles were incidentally caught 
in 41 fishing trips and 371 sets. Leatherback turtles accounted for 
77.3 percent of the take (n = 252 turtles, capture rate = 0.1405 
turtles/km of net) with 22.2 to 29.4 percent of turtles dead upon 
retrieval and no estimate of post-release mortality for those released 
alive. The annual catch by this fishery ranged from 1,212 to 6,160 
leatherback turtles, as estimated based on bootstrap procedures under 
different fishing effort scenarios in the 1990s (Fiedler et al. 2012). 
In 1998, a Brazilian Federal ordinance limited the use and transport of 
bottom and drift gillnets over 2.5 km long. Such regulations were 
difficult to enforce, and vessels from the ports of Itaja[iacute], 
Navegantes and Porto Belo, Santa Catarina, Brazil, deployed nets up to 
7,846 m long between 2005 and 2006 (Kotas et al. 2008). In 2010 the 
ordinance was suspended, permitting unrestricted fishing with driftnets 
(Fiedler 2012). The shark drift gillnet fishery declined steeply in 
later years, with no vessels operating in 2009 (UNIVALI/CTTMar 2010) 
likely because of target species reduction, reduced profitability, and 
IBAMA Normative Instruction N166/2007 which temporarily stopped the 
issuance of new driftnet fishing licenses and established a 2-year 
deadline by which vessels were to replace driftnets with other gear. 
Various other gillnet fisheries, such as bottom gillnets for sharks and 
mollusks, have reported leatherback mortalities as well, such as that 
occurring off Uruguay (Fallabrino et al. 2006; Laporta et al.

[[Page 48366]]

2006; Eckert et al. 2009) and the western coast of Africa (Riskas and 
Tiwari 2013).
    Longline fisheries pose a significant threat to the SW Atlantic 
DPS, as the spatio-temporal distribution of leatherback turtles 
overlaps with longline fishing effort (Fossette et al. 2014). In a 
review of reported, observed takes in hook and line fishery (primarily 
longline) interactions with leatherback turtles in all of Brazil from 
1990 to 2012, 1061 takes were documented, with 3 percent of the taken 
turtles found dead on the line and another 37.5 percent of unknown 
condition after release (Giffoni et al. 2013). High frequencies of 
leatherback deaths from bycatch have been documented on longline 
fishing vessels from southern Brazil and Uruguay (Kotas et al. 2004; 
Pinedo and Polacheck 2004; Domingo et al. 2006; Giffoni et al. 2008; 
Monteiro 2008). Between 2004 and 2005, in a study off southern Brazil, 
eight leatherback turtles were captured, with a mean capture rate of 
0.03 turtles per 1,000 hooks (Monteiro 2008). In 1999, there were 70 
longliners in the Brazilian fleet, with 33 vessels operating out of 
southern Brazil and fishing a total of 13,598,260 hooks (ICCAT 2001). 
However, the overall effort in the area was much higher, as longliners 
from Uruguay, Chile, Japan, Taiwan, and Spain fish in this area (Folsom 
1997; Weidner and Arocha 1999; Weidner et al. 1999). Scientific 
observers documenting 10 trips by longline vessels from the Uruguayan 
fleet operating in the SW Atlantic Ocean between 26[deg] and 37[deg] S 
between April 1998 and November 2000 observed 27 incidentally caught 
leatherback turtles (Balestre et al. 2003). The prevalence of 
leatherback interactions in pelagic longline fisheries is likely a 
result of the longline fleet fishing the productive areas in the 
convergence zone of the Brazilian Current and the cold waters from the 
Falklands Current (Kotas et al. 2004), which coincides with important 
sea turtle foraging and developmental habitat. As with gillnets, the 
scope of the longline threat to the SW Atlantic DPS spans across the 
South Atlantic Ocean in both coastal and oceanic waters. In addition to 
exposure to longline fisheries off South America, coastal longline 
fisheries off Cameroon, Angola, and Namibia, and pelagic longlines in 
the Gulf of Guinea and the eastern portion of the South Atlantic Ocean 
have also been documented to take leatherback turtles (Honig et al. 
2007; Riskas and Tiwari 2013; Angel et al. 2014; Huang 2015; Gray and 
Diaz 2017). Additional evidence of longline interactions comes from the 
stranding data, where flipper injuries on some of the stranded 
leatherback turtles could have been caused by interactions with pelagic 
longlines. Onboard observers in longline fisheries off Brazil have 
reported that leatherback turtles tend to be foul-hooked in the flipper 
rather than the mouth (Kotas et al. 2004; Pinedo and Polacheck 2004; 
Lima 2007). In 2017, Brazil enacted a law (PORTARIA INTERMINISTERIAL No 
74, DE 1o- DE NOVEMBRO DE 2017) requiring the use of circle hooks in 
the pelagic longline fisheries as well as keeping specified dehooking 
and gear removal equipment on board any Brazilian longline vessel. 
Specifically, the Brazilian government required the use of 14/0 or 
larger circle hooks for all longline vessels targeting swordfish or 
tuna (https://www.jusbrasil.com.br/diarios/166677996/dou-secao-1-06-11-2017-pg-81).
    Trawl fisheries also impact the SW Atlantic DPS, mainly along 
coastal waters off southern Brazil, Argentina, and Uruguay (Gonzalez 
Carman et al. 2011; Velez Rubio et al. 2013; Monteiro et al. 2016). 
Although there are fewer interactions with trawl fisheries relative to 
other fisheries (i.e., gillnet and longline fisheries), mortality rates 
in trawl fisheries are far higher (Miller et al. 2006; Laporta et al. 
2013). Observation of the Uruguayan bottom trawl fishery, during a 
tagging and data collection program designed to increase the 
understanding of the fishery impacts on sea turtles, documented 17 
leatherback interactions from April 2002 to June 2005 (Laporta et al. 
2013). Coastal bottom trawl and artisanal gillnet fisheries were the 
main causes of death of leatherbacks found stranded in Uruguay (Velez 
Rubio et al. 2013). Recorded interactions in coastal trawl fisheries 
are also known from Gabon, Congo, and Namibia (Riskas and Tiwari 2013).
    Other fisheries such as corrals, pound nets, and pots appear to 
present a much lower risk for leatherback turtles than to other sea 
turtle species. From 1990 to 2012, Giffoni et al. (2013) documented 
only two leatherback turtles (both alive) of the 8,367 total sea 
turtles taken in those fisheries.
    While specific information is not available to permit calculating 
an estimate of overall bycatch and mortality rates of SW Atlantic 
leatherback turtles, it is clear that fisheries bycatch, especially in 
gillnets and longlines, is a major threat to the DPS. Immature and 
adult individuals are exposed to high fishing effort throughout their 
foraging range and in coastal waters near nesting beaches. Bycatch 
mortality is also high, with reported rates of up to 31 percent 
(Giffoni et al. 2013). Mortality reduces abundance, by removing 
individuals from the population; it also reduces productivity, when 
nesting females are incidentally captured and killed. Given the small 
size of the DPS, the loss of even a small number of individuals from 
fishery interactions has the potential to reduce abundance and 
productivity. Therefore, we conclude that fisheries bycatch is the 
primary threat to the SW Atlantic DPS.

Vessel Strikes

    There is little information regarding vessel strikes for the SW 
Atlantic DPS. Many of the primary foraging areas for this DPS off the 
coasts of Argentina, Uruguay, and Brazil are experiencing increased 
vessel traffic from fishing vessels, cargo transport, and tourism 
(L[oacute]pez-Mendilaharsu et al. 2009; Fossette et al. 2014), so 
leatherback turtle interactions with vessels may occur. Affected 
individuals likely include immature and mature turtles. Impacts range 
from injury to mortality. We conclude from the best available 
information that vessel strikes are likely a threat to the DPS.

Pollution

    As with all leatherback turtles, entanglement in and ingestion of 
marine debris and plastics is a threat that likely kills several 
individuals a year. Multiple studies have implicated the ingestion of 
marine debris and/or entanglement in cases of injury or death of 
turtles found in waters occupied by the SW Atlantic DPS (Bugoni et al. 
2001; Eckert et al. 2009; Schulyer et al. 2013; Scherer et al. 2014). 
However, no individuals were assigned to a particular population and 
could have been members of the more abundant SE Atlantic DPS, which is 
known to occupy the same waters.
    While there is no specific information on effects to leatherback 
turtles of this DPS, pollution from various economic activities 
including maritime transport, tourism, and domestic and industrial 
waste discharges that are known to occur within their range, may also 
have an impact (L[oacute]pez-Mendilaharsu et al. 2009; Fossette et al. 
2014). Events such as the failure of a mining tailings dam in 2015 that 
resulted in the discharge of tons of mining mud contaminated with heavy 
metals into the Doce River, and subsequently into the waters off 
Esp[iacute]rito Santo nesting beaches, are also a concern, though no 
specific impacts to leatherback turtles have so far been noted from 
that event (Garcia et al. 2017). There is also concern about the 
potential for increased oil and gas exploration activities 
(Thom[eacute] et al. 2007). The petroleum industry in Brazil

[[Page 48367]]

has implemented a beach monitoring program, along large stretches of 
the Brazilian coast, including Esp[iacute]rito Santo, to monitor for 
potential impacts to sea turtles and their nesting beaches from 
industry activities (Werneck et al. 2018)
    Assigning impacts of pollution specifically to individuals within 
the SW Atlantic DPS is difficult, and the best available information 
does not quantify such impacts. However, given its prevalence, we 
conclude that pollution poses a threat to the DPS.

Climate Change

    Climate change poses a threat to the SW Atlantic DPS. The impacts 
of climate change include: Increases in temperatures (air, sand, and 
sea surface); sea level rise; increased coastal erosion; more frequent 
and intense storm events; and changes in ocean currents.
    Because leatherback turtles nest lower on the beach than other sea 
turtles, their eggs are more at risk of being exposed and destroyed by 
increases in sea level and coastal erosion (Boyes et al. 2010). 
Additionally, given the limited availability of suitable nesting 
habitat, the loss of the current nesting habitat with no buffer area to 
move into would pose a major problem for the DPS. Thus, rising sea 
level and beach erosion are potential threats to the DPS.
    While we do not have specific information on pivotal temperatures 
and temperature thresholds for egg mortality for this DPS, sand 
temperatures influence egg viability and sex determination. Given the 
potential lack of suitable nesting habitat outside the area currently 
being utilized, there is little opportunity for a spatial shift in 
nesting in response to changing temperatures. This DPS exhibits some 
year-round nesting, which provides a small measure of resilience to 
counteract increasing temperatures. However, it is not likely to be 
sufficient to make up for the loss of nesting habitat and opportunity 
resulting from sea level rise and temperature increases. The impacts on 
productivity and survivorship for such shifts in nesting are unknown.
    The threat of climate change is likely to modify the nesting 
conditions for the DPS. Adverse impacts on turtles of the SW Atlantic 
DPS would be inescapable because the entire DPS is confined to a 
limited nesting area. Impacts are likely to range from small, temporal 
changes in nesting season to large losses of productivity. Therefore, 
we conclude that climate change is a threat to the DPS.

Channel Dredging

    There is evidence of interactions with hopper dredges associated 
with channel dredging and maintenance. Between 2008 and 2014, four 
leatherback turtles were killed by hopper dredges in Rio de Janeiro 
(Goldberg et al. 2015).

Conservation Efforts

    There are numerous efforts to conserve the leatherback turtle. The 
following conservation efforts apply turtles of the SW Atlantic DPS 
(for a description of each effort, please see the section on 
conservation efforts for the overall species): Southwest Atlantic Sea 
Turtle Network, Convention on the Conservation of Migratory Species of 
Wild Animals, Convention on Biological Diversity, Convention on 
International Trade in Endangered Species of Wild Fauna and Flora, 
Convention Concerning the Protection of the World Cultural and Natural 
Heritage (World Heritage Convention), FAO Technical Consultation on Sea 
Turtle-Fishery Interactions, IAC, MARPOL, IUCN, Ramsar Convention on 
Wetlands, RFMOs, South Atlantic Association, UNCLOS, and UN Resolution 
44/225 on Large-Scale Pelagic Driftnet Fishing. Although numerous 
conservation efforts apply to the turtles of this DPS, they do not 
adequately reduce its risk of extinction.

Extinction Risk Analysis

    After reviewing the best available information, the Team concluded 
that the SW Atlantic DPS is at ``high'' risk of extinction. The DPS 
exhibits a total index of nesting female abundance of 27 females at the 
index beach. Such a nesting population size places this DPS at risk of 
stochastic or catastrophic events that increase its extinction risk. 
Although there has been substantial variability in nesting at the index 
nesting beach since 1986, the nest trend shows a strong, nearly five 
percent annual increase through 2017, with the largest increase 
occurring in the past decade. However, nesting has declined in the past 
3 years. There is only one nesting aggregation, limited to a relatively 
small stretch (47 km) of beach along a single coast. Some nesting also 
occurs outside that area, but is mostly sporadic and limited by sand 
and temperatures unsuited for nesting. Thus, stochastic events have the 
potential to have catastrophic effects on the entire DPS, with no 
distant subpopulations serving as a buffer or source of additional 
individuals or diversity. Based on these factors, we find the DPS to be 
at risk of extinction as a result of its limited abundance, spatial 
structure, and resilience.
    Current threats place this DPS at further risk of extinction. The 
primary threat to this DPS is bycatch in commercial and artisanal, 
pelagic and coastal fisheries, especially gillnet and longline 
fisheries. Fisheries bycatch reduces abundance by removing individuals 
from the population. Because several fisheries operate near nesting 
beaches, productivity is also reduced when nesting females are 
prevented from returning to nesting beaches. Exposure to and impact of 
this threat are high. Additional threats include: Habitat modification, 
overutilization, predation, pollution, vessel strikes, and climate 
change. Habitat modification includes incidents such as the mining dam 
breach upstream of the Doce River, which flows into the ocean through 
the middle of the primary nesting beach. Overutilization and predation 
are threats for this DPS as well, though some protective measures 
exist. While many laws are in place to protect sea turtles from fishery 
impacts, the continued impact of bycatch indicates that regulatory 
mechanisms are inadequate to sufficiently address the threat. Pollution 
and vessel strikes are potentially increasing threats to the DPS. 
Climate change is another threat that is likely to increase, resulting 
in reduced productivity due to greater rates of coastal erosion and 
nest inundation, and in some areas, nest failure or skewed sex ratios 
due to increased sand temperatures.
    We conclude, consistent with the Team's findings, that the SW 
Atlantic DPS is currently in danger of extinction. The total index of 
nesting female abundance make the DPS highly vulnerable to threats 
despite the apparent increasing nesting trend. In addition, this DPS 
consists of only one small nesting aggregation with limited potential 
nesting beaches to the north and south for expansion. The limited 
nesting range and small size makes the DPS highly vulnerable to 
stochastic impacts in the natural environment as well as singular, 
large-scale, anthropogenic events such as oil spills. Some degree of 
resilience is provided by the use of multiple foraging areas across a 
vast geographic area. However, that expansive foraging range also 
exposes the DPS to numerous fisheries (which are coastal and on the 
high seas, artisanal and commercial, off both South America and western 
Africa), making fisheries bycatch by far the biggest threat to the DPS. 
Although numerous conservation efforts apply to the turtles of this 
DPS, they do not adequately reduce the risk of extinction.

[[Page 48368]]

We conclude that the SW Atlantic DPS is currently in danger of 
extinction throughout its range and thus meets the definition of an 
endangered species. The threatened species definition does not apply 
because the DPS is at risk of extinction now (i.e., at present), rather 
than on a trajectory to become so within the foreseeable future.

SE Atlantic DPS

    The Team defined the SE Atlantic DPS as leatherback turtles 
originating from the SE Atlantic Ocean, north of 47[deg] S, east of 
20[deg] W, and west of 20[deg] E; the NW boundary is a diagonal line 
between 12.084620[deg] N, 20[deg] W and 16.063[deg] N, 16.51[deg] W. 
The eastern boundary occurs at the southern tip of Africa, where the 
Agulhas and Benguela Currents meet. Along with the cold waters of the 
Antarctic Circumpolar Current, these currents likely restrict the 
nesting range of this DPS. We placed the western boundary at the 
20[deg] W meridian as an approximate midpoint between SE Atlantic and 
SW Atlantic (i.e., turtles that nest in Brazil) nesting beaches and to 
reflect the DPS's wide foraging range throughout the South Atlantic 
Ocean; this DPS is more likely to be encountered in these waters 
compared to individuals from the less abundant SW Atlantic DPS. The 
northern boundary is a diagonal line between the elbow of Brazil and 
the northern boundary of Senegal because the SE Atlantic DPS does not 
appear to nest above this boundary (Fretey et al. 2007).
    The range of the SE Atlantic DPS is extensive, mirroring that of 
the SW Atlantic DPS. While nesting occurs along the western coast of 
Africa, data indicate that foraging areas and migratory paths stretch 
along the Atlantic coast of Africa from Senegal to South Africa, across 
the South Atlantic Ocean, and into the coastal waters of Brazil, 
Uruguay, and Argentina. As with the SW Atlantic DPS, this DPS does not 
appear to forage in northern latitudes.
    All nesting for the SE Atlantic DPS occurs along the Atlantic coast 
of western Africa, from Senegal to Angola, a nesting range of over 
7,500 km. However, the vast majority of nesting occurs in Gabon, 
Equatorial Guinea (including Bioko Island), and the Republic of Congo 
(TEWG 2007; Fretey et al. 2007, Witt et al. 2009; Tiwari et al. 2013). 
Gabon may have once hosted the largest nesting aggregation in the world 
when it was discovered in the early 2000s (Witt et al. 2009), but 
current data indicate much lower levels of nesting (Formia et al. in 
prep) compared to those described in Witt et al. (2009).
    While nesting occurs along the western coast of Africa, foraging 
grounds and migratory paths stretch across the South Atlantic Ocean to 
the coastal waters of Brazil, Uruguay, and Argentina. Because of the 
greater abundance of this DPS, most individuals found in the western 
South Atlantic along the coast of South America, and on the high seas, 
belong to the SE Atlantic DPS. Prosdocimi et al. (2014) found 84 to 86 
percent of leatherback turtles sampled from the foraging grounds off 
Argentina and Eleva[ccedil][atilde]o do Rio Grande (an elevated 
offshore area across from Brazil) to originate from western African 
beaches.

Abundance

    The total index of nesting female abundance for the SE Atlantic DPS 
is 9,198 females. We based this total index on nine nesting 
aggregations in Gabon (n = 8,495 nesting females), Equatorial Guinea (n 
= 457), Republic of Congo (n = 69), Sierra Leone (n = 39), Liberia (n = 
45), Ivory Coast (n = 40), Ghana (n = 4), Cameroon (n = 3), and Sao 
Tome and Principe (n = 46). Our total index does not include 10 
unquantified nesting aggregations in Guinea-Bissau, Angola, and other 
nations. For more information on data sources and calculations, please 
see the Status Review Report.
    Our total index of nesting female abundance is an index because we 
do not have consistent data from much of the nesting range of the DPS, 
which extends from Senegal to Angola. However, the largest nesting 
aggregations occur in Gabon, Equatorial Guinea (including Bioko 
Island), and the Republic of Congo (TEWG 2007; Fretey et al. 2007; Witt 
et al. 2009; Tiwari et al. 2013), which are represented in our total 
index. The IUCN Red List assessment did not provide an estimate of 
population size but instead concluded that the subpopulation was ``data 
deficient'' (Tiwari et al. 2013).
    To calculate the index of nesting female abundance in Gabon, where 
annual aerial surveys of 600 km of nesting beaches gather emergence 
data, we used a remigration interval of 3 years, a clutch frequency of 
7.8 clutches per season per female, and estimated that 95 percent of 
emergences resulted in nesting (Formia et al. in prep). Our index of 
nesting female abundance for Gabon (i.e., 8,495 nesting females) is 
lower than previous estimates. According to Witt et al. (2009), Gabon 
once hosted the largest leatherback nesting aggregation in the world, 
with an estimated 36,185 to 126,480 clutches per year (approximately 
15,730 to 41,373 nesting females). These estimates were based on a 
combination of aerial surveys and ground-truthing surveys, conducted 
during the 2002/2003, 2005/2006, and 2006/2007 nesting seasons. More 
recent aerial surveys indicate a steep decline in nesting since the 
early 2000s, with a high of 108,588 estimated nests in 2002/03, a low 
of 4,275 estimated nests in 2009/10, and fewer than 25,000 nests in the 
final year of available data (2015/16; Formia et al. in prep).
    Nesting is scattered on continental Equatorial Guinea (Fretey 
2001), but it occurs on several beaches of Bioko Island and is 
monitored at the Gran Caldera Scientific Reserve (n = 457 nesting 
females, based on body pit data from the 2000/2001 through 2017/2018 
nesting seasons; D. Venditti et al., Drexel University, pers. comm., 
2018). Rader et al. (2006) documented an average of 3,896 nests 
annually between the 2000/2001 to 2004/2005 nesting seasons, which 
equates to approximately 2,338 nesting females (i.e., using a 3-year 
remigration interval and a clutch frequency of 5 nests annually). Based 
on the data available on nesting in the Republic of Congo from the 
2003/2004 to 2016/2017 nesting seasons (N. Breheret, SWOT, pers. comm., 
2018), we estimated 69 nesting females. In an analysis of older data 
(1999 to 2008), Girard et al. (2016) estimated 933 nests per year on 
the monitored beaches, which equates to approximately 560 nesting 
females.
    In Guinea-Bissau, only one beach is monitored regularly, in Orango 
National Park, Bijagos Archipelago, where occasional leatherback 
nesting tracks are recorded. Each season, a few nests are reported 
elsewhere throughout the nation (Barbosa et al. 1998; Fretey et al. 
2007).
    In the Ivory Coast (n = 40 nesting females), Gomez (2005) counted 
218 nests over 41 km of beach in February 2001. Pe[ntilde]ate et al. 
(2007) reported 189 nests reported from non-exhaustive surveys of 27 km 
of coastline during the 2001/2002 nesting season.
    In Ghana, nest monitoring occurs on three beaches: Mankoadze (n = 4 
nesting females), Ada, and Keta. We were unable to calculate the index 
for Ada and Keta beaches because we only received information on nest 
averages. From 2000 to 2017, an annual average of 34 nests were 
observed on Ada Beach (D. Agyeman, pers. comm., 2018). During the 2006/
2007 nesting season, 481 leatherback nests were counted on Ada Beach 
(Allman and Armah 2010). Over an unspecified time frame, an annual 
average of 80 nests were observed on Keta Beach (A. Fuseini, pers. 
comm., 2018).

[[Page 48369]]

    In Cameroon (n = 3 nesting females; Fretey and Nibam unpublished 
data 2018), Girard et al. (2016) estimated an average of 43 leatherback 
nests annually, which would equate to 26 nesting females, from 1999 to 
2008. In S[atilde]o Tom[eacute] and Principe (n = 46 nesting females), 
Girard et al. (2016) estimated an average of 78 nests annually from 
1999 to 2008, which is similar to our estimate.
    Nesting occurs on other beaches throughout western Africa. However, 
recent consistent and standardized monitoring data are not available. 
Sporadic nesting occurs in Senegal (Maigret 1978; Dupuy 1986), Republic 
of The Gambia (Barnett et al. 2004, Hawkes et al. 2006), Togo 
(Segniagbeto 2004), Nigeria (Fretey 2001; Mojisola et al. 2015), 
Democratic Republic of Congo, (OCPE-ONG 2006), and Angola (Carr and 
Carr 1991; Weir et al. 2007).
    The total index of nesting female abundance of the SE Atlantic DPS 
(9,198 females) does not reduce the risk for environmental variation, 
genetic complications, demographic stochasticity, negative ecological 
feedback, and catastrophes (McElhany et al. 2000; NMFS 2017). Such 
abundance provides little resilience to buffer losses of individuals. 
We conclude that the nesting female abundance, as estimated, does not 
reduce the extinction risk of this DPS.

Productivity

    Based on data collected from the largest nesting aggregation (i.e., 
Gabon), the SE Atlantic DPS exhibits a declining nesting trend. Data 
collected between the 2002/2003 and 2015/2016 nesting seasons (with two 
years of missing data) indicated a median trend in nesting activity of 
-8.6 percent annually (sd = 21.9 percent; 95 percent CI = -52.6 to 36.9 
percent; f = 0.676; mean annual nesting activities = 35,204). The trend 
in Gabon is likely representative of the entire DPS, because the 
majority of nesting occurs there. Additional nest trend data are 
available from the Gran Caldera Scientific Reserve of Bioko Island, 
where the number of body pits increased 2.8 percent annually (sd = 15.6 
percent; 95 percent CI = -27.2 to 36.0 percent) from 1996/1997 to 2017/
2018.
    Regarding productivity parameters, available information is often 
from a limited area and may not be representative of the entire DPS. 
However, based on available data, the size of nesting females, clutch 
size, hatching success, and incubation period appear to be similar to 
the species' averages. We conclude that the declining nesting trend 
contributes to the extinction risk of this DPS.

Spatial Distribution

    The SE Atlantic DPS has a broad spatial distribution. The nesting 
range is centered on Gabon, with nesting occurring from Senegal to 
Angola. Genetic data available for Gabon and Ghana indicate significant 
genetic differentiation based on mtDNA data, but weak differentiation 
based on analysis of nuclear DNA, likely indicating demographically 
independent subpopulations connected by limited gene flow (Dutton et 
al. 2013).
    In addition to the extensive nesting range, this DPS also has an 
expansive foraging and migratory range, from the coastal waters of 
Atlantic Africa, across the pelagic waters of the South Atlantic, and 
along the South American coast from Brazil to Argentina. While nesting 
along the coast of Africa extends only to Angola, recent tag returns 
and satellite telemetry indicate that turtles utilize the waters in 
Namibia as well (Almeida et al. 2014). Transatlantic movements were 
first recorded from tag returns of four leatherback turtles tagged on 
the nesting beaches of Gabon and recaptured in the waters of Argentina 
and Brazil (Billes et al. 2006). Satellite telemetry confirmed that 
nesting females from Gabon follow three different post-nesting movement 
trajectories towards the equatorial Atlantic Ocean, South America, or 
southern Africa (Witt et al. 2011). For combined foraging areas off 
Argentina and Eleva[ccedil][atilde]o do Rio Grande (an elevated 
offshore area across from Brazil), the mean estimate from western 
Africa was 84 to 86 percent (45 percent Gabon, 41 percent Ghana; 
Prosdocimi et al. 2014).
    The wide distribution of foraging areas likely buffers the DPS 
against local catastrophes or environmental changes that could limit 
prey availability. The expansive nesting range may buffer the DPS from 
acute environmental impacts (e.g., storms and singular events) and to 
some degree, chronic impacts (e.g., sea level rise and temperature 
changes). Thus, the combination of extensive nesting range, widely 
distributed foraging areas, and population structure reduces the 
extinction risk of the SE Atlantic DPS.

Diversity

    Genetic analyses for the SE Atlantic DPS are limited, but Dutton et 
al. (2013) found moderate genetic diversity in samples from Gabon and 
Ghana, including four new haplotypes unique to western African nesting 
females. Nesting occurs on continental and insular beaches. There are 
multiple foraging strategies, including pelagic and coastal, along 
either side of the Atlantic Ocean. The genetic diversity, along with 
multiple and diverse foraging sites (i.e., coastal and pelagic), and 
combination of insular and mainland nesting provide diversity and 
resilience that may reduce the extinction risk of this DPS.

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    Modification and loss of habitat is a threat to the SE Atlantic 
DPS. Present threats include obstructions, erosion, and light pollution 
at nesting beaches. Future threats include coastal construction and 
development in the region.
    Nesting beach obstruction due to logs is a problem in Gabon, 
Equatorial Guinea, and Cameroon (Formia et al. 2003). Logs that have 
broken loose from timber rafts of industrial logging operations wash up 
on the beaches of Gabon at densities of up to 247 logs/km; logs blocked 
30.5 percent of the beach in Pongara, Gabon, resulting in an estimated 
2,111 disrupted or aborted nesting attempts (Laurance et al. 2008). In 
addition, several leatherback turtles have died as result of being 
trapped by logs (Laurance et al. 2008). Pikesley et al. (2013) 
determined that between 1.6 percent and 4.4 percent of nesting females 
could be trapped at beaches with high log- and turtle-densities. 
However, Gabon has since banned the export of whole logs. The Gabon Sea 
Turtle Partnership has carried out log removal efforts for at least one 
high-density nesting beach in Pongara National Park (Kingere Beach), 
and a 3 km stretch of nesting beach is now virtually free of logs; at 
the other main monitored beaches in Gabon, such as Mayumba and Gamba, 
logs are not a major threat (A. Formia, WCS, pers. comm. 2019).
    Habitat loss from coastal erosion due to sand mining, harbor 
building, and irregular current flows has compromised the suitability 
of long stretches of coastal areas as nesting sites. This issue is 
especially prevalent between Ghana and Nigeria (Formia et al. 2003). 
Ikaran (2010) found low hatching/emergence success rates at three 
nesting sites in Gabon: Pointe Denis (17/16 percent), Mayumba (43/40 
percent), and Kingere (16/16 percent).In addition to predation, the 
main identified sources of egg mortality were beach erosion and 
inundation (Ikaran 2010).
    Light pollution modifies nesting beach habitat, deterring nesting 
females

[[Page 48370]]

and disorienting both hatchlings and nesting females. Bourgeois (2009) 
found that artificial lighting disoriented leatherback hatchlings in 
Pongara National Park, Gabon: Hatchlings in 27 of the 41 nests (66 
percent) studied crawled towards artificial lights. Deem et al. (2007) 
documented 71 disoriented females that crawled directly into the 
savannah behind the beach and towards the artificial lights. Bourgeois 
et al. (2009) concluded that light pollution from Libreville and Pointe 
Denis, Gabon is a major threat to nesting females and hatchlings, which 
become disoriented and die in the surrounding savannah.
    Urbanization and coastal development are rapidly growing threats at 
some nesting beaches (Girard and Honarvar 2017). There is a high 
potential for coastal development in Gabon, including the beaches near 
Pointe Denis, an important and growing tourist area (Ikaran 2010). 
Along with direct habitat loss from coastal development and 
urbanization, impacts from pollution and litter are expected to 
increase.
    In Gabon, a network of marine protected areas was created by decree 
00161/PR in 2017, covering 26 percent of Gabon's territorial seas, 
including a vast area in front of the most important nesting beach in 
Gabon (Mayumba National Park) that stretches to the outer limits of the 
EEZ.
    We conclude that a large portion of nesting females, hatchlings, 
and eggs are exposed to the reduction and modification of nesting 
habitat, as a result of logging, erosion, coastal development, and 
artificial lighting. These threats impact the DPS by reducing nesting 
and hatching success, thus lowering the productivity of the DPS. 
Logging also results in the death of nesting females, reducing the 
abundance of the population by removing its most reproductively 
important individuals. Based on the information presented above, we 
conclude that habitat loss and modification are major and increasing 
threats to the DPS.


Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    Overutilization is a threat to the SE Atlantic DPS. Although 
receiving some legal protections, eggs and turtles nevertheless are 
poached for consumption, traditional medicine, and religious practices.
    In Gabon, poaching is limited because 78 percent of nesting occurs 
within national parks and human population density along the coast is 
low (A. Formia, Gabon Sea Turtle Partnership, pers. comm., 2018). 
However, elsewhere in the region, poaching occurs at a high rate, or 
would be reasonably expected to return to high levels, if not limited 
by activities funded through the USFWS' Marine Turtle Conservation Fund 
enacted under the MTCA. These activities reduce poaching through 
increased presence on nesting beaches, beach monitoring, hiring of 
local citizens for participation in the projects, and raising awareness 
and providing education to local communities (M. Tiwari, NMFS, pers. 
comm. 2018).
    Conflicting beliefs about sea turtles exist throughout the region. 
In some communities sea turtles are considered divinely provided food, 
while in others they have been historically protected by indigenous 
custom, often based on stories passed down by ancestors (Barbosa and 
Regalla 2016; Alexander et al. 2017). In general, however, poaching is 
a significant problem throughout the region. Catry et al. (2009) 
concluded that, in addition to fisheries bycatch, poaching of eggs and 
nesting females is the main threat to sea turtles, including 
leatherback turtles, in Guinea-Bissau. In many cases ``few if any 
turtles or nests are left alone when found by locals'' (Catry et al. 
2009). The fat of leatherback turtles is often used for various 
purported medicinal applications, including: Treatment of convulsions 
and malaria (Togo), fever, fainting spells, liver problems, tetanus 
(Benin), and to induce vomiting (Togo, Benin). In one community in the 
Ivory Coast and parts of Cameroon, leatherback turtle fat is applied to 
wounds in the mouth and is used to massage into painful joints. In 
northwestern and southern Cameroon, it is applied to bruises (Fretey et 
al. 1999). In Togo, some mothers add turtle bones daily to the baby's 
bath water; some believe that the power of the turtle (especially the 
leatherback) will be transmitted to the child through this practice 
(Segniagbeto 2004).
    Turtles and eggs are poached throughout the nesting range of the 
DPS. Though most nesting females and eggs are protected in Gabon, 
poaching is widespread in other areas. Poaching of nesting females 
reduces both abundance (through loss of nesting females) and 
productivity (through loss of reproductive potential). Such impacts are 
high because they directly remove the most productive individuals from 
DPS, reducing current and/or future reproductive potential. Egg 
poaching reduces productivity. Given the moderate exposure and high 
impact, we conclude that the poaching of turtles and eggs poses a 
threat to the DPS.

Disease or Predation

    Information on diseases among leatherback turtles originating in 
the SE Atlantic is minimal, but an analysis of samples from nesting 
females in Gabon indicated normal blood chemistry parameters (Deem et 
al. 2006). Predation may occur at high rates in some areas, but 
information is limited.
    Predation of leatherback eggs and/or hatchlings has been documented 
for a variety of predators, including: Various ants, ghost crabs, 
monitor lizards (Varanus niloticius), crows (Corvus albus), mongoose, 
porcupine (Atherurus africanus), domestic dogs, African civet cat 
(Civettictis civetta and Viverra civetta), and drills (Mandrillus 
leucophaeus) (summarized from Eckert et al. 2012). In Kingere, Gabon, 
Ikaran (2010) noted high predation rates of eggs by crabs, lizards, 
mongooses, small cat species, and ants. Predation was the main source 
of egg mortality at three nesting sites in Gabon: Pointe Denis (43 
percent), Mayumba (44 percent), and Kingere (51 to 56 percent; Ikaran 
2010).
    As is common for all sea turtle species, leatherback hatchlings 
likely experience predation from various fish species as they enter the 
water and swim towards the open ocean. In-water predation of juveniles 
and adults is not well-documented, but there is evidence of shark and 
killer whale predation. Shark predation was determined to be the cause 
of one leatherback stranding reported from Central Africa (Parnell et 
al. 2007), while interactions between killer whales and leatherback 
turtles resulting in possible predation has been observed in Namibian 
waters (Elwen and Leeney 2011).
    While all eggs and hatchlings have some exposure to predation, the 
species compensates for a certain level of natural predation by 
producing a large number of eggs and hatchlings. For this DPS, the 
primary impact is to productivity (i.e., reduced egg and hatching 
success). We conclude that predation poses a threat to the SE Atlantic 
DPS.

Inadequacy of Existing Regulatory Mechanisms

    The SE Atlantic DPS is protected by various regulatory mechanisms. 
For each, the Team reviewed the objectives of the regulation and to 
what extent it adequately addresses the targeted threat.
    The harvest of turtles and eggs is illegal in most of the nations 
where the DPS nests. In some cases, however, these protective 
mechanisms are inadequate. In addition, many nesting beaches are not 
protected.
    In Gabon, the harvest of turtles and eggs is illegal (2011 decree 
0164/PR/

[[Page 48371]]

MEF) and much of the nesting beach habitat (and turtles utilizing that 
habitat) is protected because of inclusion in parks as well as being 
far from any city or town. However, low levels of poaching occurs, and 
the threats from encroaching development and associated impacts are 
increasing.
    In Congo, wildlife laws prohibit the hunting and collection of 
wildlife and their products, including eggs, between November 1 and 
April 31. Turtles are also protected in the Conkaouati-Douli National 
Park. However, in areas without permanent beach monitoring, almost all 
eggs and nesting individuals are collected and eaten (Bal et al. 2007).
    In the Democratic Republic of Congo, leatherback turtles are cited 
under the 1982 Hunting Act for protection. However, there is no post-
independence legislation protecting sea turtles, and there is little 
commitment to the legislated protections (Fretey 2001).
    Since 1988, Equatorial Guinea has protected all sea turtles under 
Law 8/1988 and Decree 183/87 on fishing (Tom[aacute]s et al. 2010). 
However, the poaching of eggs and females for local consumption and 
sale has occurred (Castroviejo et al. 1994).
    In Ghana, the Wildlife Regulations Act of 1974 prohibits all 
harvest of eggs and turtles. However, poverty is prevalent, and eggs 
and sea turtles are poached at nesting beaches (Tanner 2013). 
Enforcement is likely inadequate because of funding issues, the 
remoteness of some nesting beaches, and cultural practices.
    Fishery bycatch is the primary threat to this DPS. While most 
nations in the region have some form of legal protection for sea 
turtles, many leatherback turtles die from fisheries bycatch throughout 
the range of the DPS. Examples of fisheries legislation include 
Brazil's gear restrictions and Nigeria's requirement to use TEDs in 
bottom trawls.
    In summary, numerous regulatory mechanisms provide some protection 
to leatherback turtles, their eggs, and nesting habitat throughout the 
range of this DPS. Though the regulatory mechanisms provide some 
protection to the turtles, many do not adequately reduce the threat 
that they were designed to address, generally as a result of limited 
implementation or enforcement. Fisheries bycatch, poaching, and habitat 
loss remain major threats to the DPS despite regulatory mechanisms. We 
conclude that inadequacy of the regulatory mechanisms are a threat to 
the SE Atlantic DPS.

Fisheries Bycatch

    Fisheries bycatch is the primary threat to the SE Atlantic DPS. 
Leatherback turtles are captured as bycatch in commercial and artisanal 
fisheries along coastal foraging and breeding areas as well as on the 
high seas. Because of the overlapping range with the SW Atlantic DPS, 
this DPS is vulnerable to interactions with fisheries off the coasts of 
Brazil, Uruguay, and Argentina, in the pelagic waters of the South 
Atlantic Ocean, and along the coastal waters off western Africa. 
Therefore, the information presented on the fisheries bycatch for the 
SW Atlantic is applicable to this DPS.
    One of the biggest threats for leatherback turtles in Atlantic 
waters is bycatch in artisanal and commercial fisheries (Wallace et al. 
2010; Riskas and Tiwari 2013;). Lewison et al. (2004) estimated that 
30,000 to 60,000 leatherback turtles were taken as longline fisheries 
bycatch in the entire Atlantic Ocean in 2000. Stewart et al. (2010) 
estimated that in West Africa, Benin, Togo, and Cameroon had the 
highest average fishing densities, ranging from 11.1 to 6.5 boat-
meters/km\2\, and gillnet densities ranked among the highest on a 
global scale. Despite very active artisanal and industrial fisheries in 
the region, overall bycatch data are quite sparse or qualitative 
(rather than quantitative) in nature, and Africa still represents a 
significant gap in bycatch evaluation studies (Wallace et al. 2010, 
2013). Accurate and reliable bycatch data are difficult to achieve, as 
direct observation rates are low (<1 percent of total fleets) and 
statistics from the region's many small-scale fisheries are largely 
incomplete (Kelleher 2005; Moore et al. 2010; Wallace et al. 2010). 
However, several studies have concluded that bycatch rates in the 
region are high, given the degree of fishing activity near nesting and 
foraging areas (Lewison et al. 2004; Moore et al. 2010; Wallace et al. 
2010).
    Along the coasts of Angola, Namibia, and South Africa, Honig et al. 
(2007) evaluated turtle bycatch by longline fisheries in the Benguela 
Large Marine Ecosystem by using data from observer reports, surveys, 
and specialized trips from the coastal nations of South Africa, Namibia 
and Angola. They estimated bycatch at 672 leatherback turtles annually 
(based on an annual bycatch estimate of 4,200 turtles, of which 
approximately 16 percent are leatherback turtles) in the southern and 
central regions and as many as 5,600 leatherback turtles (based on an 
annual bycatch estimate of 35,000 turtles) for the entire Benguela 
Large Marine Ecosystem (Honig et al. 2007). Mortality rates were not 
provided in this study but may range from 25 to 75 percent (Aguilar et 
al. 1995). The estimates mostly include turtles from the SE Atlantic 
DPS, but telemetry studies indicate that the turtles of the much 
smaller SW Indian DPS also use this foraging area (Luschi et al. 2006; 
Robinson et al. 2016). Evaluating ICCAT data, Angel et al. (2014) 
confirm exposure to high longline fishing effort and some purse seine 
effort for the population originating from the SE Atlantic Ocean.
    The limited bycatch data available for waters of the western coast 
of Africa show that other fisheries interact with leatherback turtles. 
Between 2005 and 2015, artisanal fishing nets in Loango Bay in the 
Republic of Congo killed a total of 45 leatherback turtles; 0 to 628 
leatherback turtles were captured or recaptured annually over that time 
period (Br[eacute]heret et al. 2017). An assessment of bycatch in the 
trawling fisheries in Gabon found that leatherback turtles represented 
only 2 percent of the bycatch despite being the most abundant sea 
turtle species in Gabonese waters; the low rate is possibly because 
leatherback turtles do not occur in the section of the water column 
where the trawl net is towed (Casale et al. 2017). Trawl bycatch in the 
waters around S[atilde]o Tom[eacute] and Principe included 4 juvenile 
leatherback turtles (17 to 21 cm in carapace length) in March 1994 
(Fretey et al. 1999).
    While specific information to estimate overall capture and 
mortality rates of SE Atlantic leatherback turtles in fisheries is not 
available, it is clear that bycatch in fisheries, especially gillnets 
and longlines, are a threat to the DPS across its range. Immature and 
mature individuals are exposed to high fishing effort throughout their 
foraging range and in coastal waters near nesting beaches. Mortality is 
also high. Mortality reduces abundance, by removing individuals from 
the population; it also reduces productivity, when nesting females are 
incidentally captured and killed. We conclude that fisheries bycatch is 
a major, and the primary, threat to the SE Atlantic DPS.

Vessel Strikes

    There is little information regarding vessel strikes for the SE 
Atlantic DPS, but such interactions are a potential, and possibly 
increasing, threat across at least a portion of this DPS's range. In 
the western South Atlantic foraging grounds off Brazil, Uruguay, and 
Argentina, increasing vessel traffic from fishing vessels, cargo 
transport, and tourism has been noted (L[oacute]pez-Mendilaharsu et al.

[[Page 48372]]

2009; Fossette et al. 2014), potentially increasing the likelihood of 
vessel strikes on leatherback turtles. Although no specific information 
is available for the waters off western Africa, any economic 
development along the coast is likely to result in an increase in 
vessel traffic. We conclude that vessel strikes are a threat to the SE 
Atlantic DPS.

Pollution

    The SE Atlantic DPS faces the threat of pollution across its 
extensive range throughout the South Atlantic Ocean, from Africa to 
South America. As the ranges of the SW Atlantic and SE Atlantic DPSs 
overlap, they are exposed to the same pollutants, which include 
contaminants, marine debris, and ghost fishing gear. Throughout Africa, 
marine and coastal pollution is widespread in industrial and urban 
areas, and garbage litters many developed beaches (Formia et al. 2003; 
Agyekumhene et al. 2017). Off the coast of South America, the Argentine 
and Brazilian coastal waters are increasingly impacted by economic 
activities, such as maritime cargo transport, tourism, and the 
discharge of domestic and industrial waste (L[oacute]pez-Mendilaharsu 
et al. 2009; Fossette et al. 2014).
    The Gulf of Guinea has increasingly been the focus of extensive oil 
exploitation activities, following the discovery of large oil reserves. 
Drilling activities by large oil corporations, with associated 
pollution and habitat destruction, are threats to nesting aggregations 
in the area (Formia et al. 2003; Agyekumhene et al. 2017). In 2012/
2013, oil spills following the dredging of the Port of Pointe-Noire in 
the Republic of Congo significantly degraded the fauna and flora of 
Loango Bay, where leatherback turtles occur. However, the ecosystem is 
believed to be slowly recovering (Br[eacute]heret et al. 2017). In 
2005, a moderate slick of oil on the beaches of Mayumba National Park 
in Gabon was observed, although its impacts on turtles are unknown 
(Parnell et al. 2007).
    In Nigeria, the main sources of pollution include industrial waste, 
raw/untreated sewage, and pesticides. Oil exploration, exploitation, 
and transportation have a significant effect on the environment. Spills 
of crude and refined oil are frequent in the coastal and marine 
environment, especially during periods of very strong ocean currents, 
when they can spread to cover the entire 853 km coastline of Nigeria.
    It is clear that individuals from the SE Atlantic DPS have a high 
probability of encountering pollution across their range and throughout 
their lifecycle. Although the best available information does not 
quantify such impacts, ample information demonstrates that these 
threats are ongoing. We conclude that pollution is a threat to the DPS.

Climate Change

    Climate change is a threat to the SE Atlantic DPS. The impacts of 
climate change include: Increases in temperatures (air, sand, and sea 
surface); sea level rise; increased coastal erosion; more frequent and 
intense storm events; and changes in ocean currents.
    Sea level rise resulting from climate change negatively impacts sea 
turtle nesting. Erosion of important nesting beaches in Gabon may be at 
least partially attributable to sea level rise. From 1983 through the 
2000s, some areas have lost up to 100 m of beach width, reducing the 
availability of suitable nesting beach (Gabon Sea Turtle Partnership 
2018; http://www.seaturtle.org/groups/gabon/erosion.html). Because 
leatherback turtles nest lower on the beach than other sea turtles, 
their eggs are more at risk of being inundated and destroyed by 
increases in sea level and coastal erosion (Boyes et al. 2010).
    Changes in sand temperatures are likely to impact egg viability and 
sex determination. Ikaran (2010) found the thermal range of sand over 
the nesting season to be adequate for hatchling sex ratios to be mixed 
or even male dominated. In Gabon, the early rainy months tend to 
produce males, while the later, warmer months produce females, with a 
tendency towards a net higher production of males. Ikaran (2010) 
considered the nesting beaches of Gabon to be an important male 
producing area. However, based on predictions of warming trends, he 
found that within two decades the ratio could skew towards 100 percent 
female.
    The threat of climate change is likely to modify the nesting 
conditions for turtles of the DPS, and it is unclear whether they have 
or can develop the ability to nest in different locations along 
existing beaches, or on new beaches. Impacts from climate change are 
likely to range from small, temporal changes in nesting season to large 
losses of productivity. Therefore, we conclude that climate change is a 
threat to the DPS.

Conservation Efforts

    There are numerous efforts to conserve the leatherback turtle. The 
following conservation efforts apply within the range of the SE 
Atlantic DPS (for a description of each effort, please see the section 
on conservation efforts for the overall species): Convention on the 
Conservation of Migratory Species of Wild Animals, Convention on 
Biological Diversity, Convention on International Trade in Endangered 
Species of Wild Fauna and Flora, Convention Concerning the Protection 
of the World Cultural and Natural Heritage (World Heritage Convention), 
FAO Technical Consultation on Sea Turtle-Fishery Interactions, IAC, 
MARPOL, IUCN, Memorandum of Understanding Concerning Conservation 
Measures for Marine Turtles of the Atlantic Coast of Africa, Ramsar 
Convention on Wetlands, South-East Atlantic Fisheries Organization, 
UNCLOS, and UN Resolution 44/225 on Large-Scale Pelagic Driftnet 
Fishing. Although numerous conservation efforts apply to the turtles of 
this DPS, they do not adequately reduce its risk of extinction.

Extinction Risk Analysis

    After reviewing the best available information, the Team concluded 
overall that the SE Atlantic DPS is at high risk of extinction. The 
total index of nesting female abundance is 9,198 females. Since 2002, 
the first year that aerial survey data was collected, nesting activity 
has declined by -8.6 percent annually in Gabon, the largest nesting 
aggregation of the DPS, and what was, in 2002, the largest nesting 
aggregation in the world. This declining trend has the potential to 
further lower abundance and increase the risk of extinction. Nesting 
and foraging is broadly distributed; thus, the population is somewhat 
buffered from stochastic events that could otherwise have catastrophic 
effects on the entire DPS. There is a metapopulation structure within 
this DPS, with fine-scale genetic differentiation between Gabon and 
Ghana. Genetic diversity also appears to be moderate. Based on the 
reduced nesting female abundance and declining nest trend, we find the 
DPS to be at risk of extinction, likely as a result of past threats.
    Current threats place the DPS at further risk of extinction. The 
primary threat to this DPS is bycatch in commercial and artisanal, 
pelagic and coastal, fisheries, especially coastal gillnet and pelagic 
longline fisheries. Fisheries bycatch reduces abundance by removing 
individuals from the population. Because several fisheries operate near 
nesting beaches, productivity is also reduced when nesting females are 
prevented from returning to nesting beaches. Thus, exposure and impact 
of this threat are high. Habitat loss or modification is a threat that 
reduces abundance and productivity and includes the impacts of logs, 
which block access to the

[[Page 48373]]

beaches or trap nesting females and hatchlings. Poaching of turtles and 
eggs is also a threat to this DPS, although most nesting beaches in 
Gabon are somewhat protected because they occur in parks or are far 
from any towns. Many of the beaches outside Gabon (e.g., Guinea-Bissau) 
have limited or no protection. The degree of overutilization is highly 
varied across locations, but quite extensive in some areas. Funding 
from the MTCA has resulted in some reduction of this threat as 
conservation activities, research, and community involvement results in 
lower poaching on those beaches. However, poaching continues at high 
levels in other areas. Additional threats include: predation and 
disease, inadequate regulatory mechanisms, pollution, and climate 
change. Predation can be extensive at some specific beaches, but 
overall it does not occur at a high level. Pollution is a persistent 
and potentially increasing threat. Ingestion of plastics and 
entanglement in marine debris result in injury and reduced health, and 
sometimes mortality. Climate change is likely to result in reduced 
productivity due to greater rates of coastal erosion and nest 
inundation, and in some areas, nest failure or skewed sex ratios due to 
increased sand temperatures. Vessel strikes are a threat that is likely 
to increase over time as recreational and commercial vessel activity 
increases, resulting in more opportunity for interactions. Though many 
regulatory mechanisms are in place, they do not adequately reduce the 
impact of logs, poaching, and fisheries. Additionally, many areas in 
the region have little or no enforcement of laws protecting turtles or 
nests on the beach.
    The DPS is relatively data-poor, reducing our ability to quantify 
threats for more than a small portion of the population. For this 
reason, the Status Review Team did not come to consensus regarding the 
extinction risk analysis for the SE Atlantic DPS. All Team members were 
present to vote on the level of extinction risk. Nine Team members 
concluded with moderate confidence that the DPS is at high extinction 
risk due to threats and loss of abundance; their confidence was 
moderate due to the lack of data on this DPS. Two team members 
concluded with low confidence that the DPS is at moderate extinction 
risk; their confidence in this conclusion is low due to the lack of 
data on this DPS.
    We conclude, consistent with the Team's overall conclusion, that 
the SE Atlantic DPS is currently in danger of extinction. The 
decreasing nesting trend (i.e., 8.6 percent annually since 2002) is at 
or near a level that make the DPS highly vulnerable to threats, given 
the total index of nesting female abundance of 9,198 females. It faces 
present, ongoing threats that are likely to create imminent and 
substantial demographic risks (i.e., declining trends and reduced 
abundance). Though numerous conservation efforts apply within the range 
of this DPS, they do not adequately reduce the risk of extinction. We 
conclude that the SE Atlantic DPS is currently in danger of extinction 
throughout its range and therefore meets the definition of an 
endangered species. The threatened species definition does not apply 
because the DPS is at risk of extinction currently (i.e., at present), 
rather than on a trajectory to become so in the foreseeable future.

SW Indian DPS

    The Team defined the SW Indian DPS as leatherback turtles 
originating from the SW Indian Ocean, north of 47[deg] S, east of 
20[deg] E, and west of 61.577[deg] E. The western boundary occurs at 
the southern tip of Africa, approximately where the Agulhas and 
Benguela Currents meet. The eastern boundary occurs at the border 
between Iran and Pakistan, where the Somali Current begins. These 
currents, and the cold waters of the Antarctic Circumpolar Current, 
likely restrict the nesting range of this DPS.
    The range of the DPS (i.e., all documented areas of occurrence) 
extends into the SE Atlantic Ocean, where leatherback turtles forage in 
the highly productive Benguela Current Large Marine Ecosystem, which 
occurs along the western coast of Africa, from Angola to South Africa. 
Leatherback turtles also range throughout the waters of eastern Africa 
(Ross 1985) and possibly into the Red Sea (Gasparetti et al. 1993). 
Records indicate that the species has been observed in the waters of 
the following nations: Djibouti; Eritrea; French Territories (Reunion 
Island, Mayotte, and Iles Eparses); Kenya; Madagascar; Mozambique; 
Seychelles; Somalia; South Africa; Tanzania; and Yemen (Hamann et al. 
2006). Leatherback turtles may occur in the waters of the following 
nations: Bahrain, Kuwait; United Arab Emirates; Oman; and Sudan (Hamann 
et al. 2006).
    Leatherback turtles of the SW Indian DPS nest over a distance of 
approximately 900 km, from Cape Vidal, South Africa to Bazaruto 
Islands, Mozambique (Videira et al. 2011; Nel et al. 2015). The vast 
majority of nesting (80 to 90 percent) occurs in South Africa, between 
Bhanga Nek and Leifeld's Rock (Nel et al. 2015). In Mozambique, most 
nesting occurs from the southern border to Inhaca Island, Mozambique, 
with low levels of nesting farther north at Bilene Beach and Bazaruto 
Islands (Nel et al. 2015). This DPS nests at the highest latitude (and 
southernmost location) of all leatherback turtles (Saba et al. 2015).
    Nesting occurs on long (5 to 15 km), broad (50 to 100 m), silica 
sand beaches with little vegetation (Botha 2010; Nel et al. 2015; 
Robinson et al. 2017). The beaches are characterized by pristine, 
intact dunes that rise up to 100 m above sea level, interspersed with a 
few dynamic dunes and small, primary dunes (Nel et al. 2015). The 
beaches are separated by short rocky headlands (Robinson et al. 2017). 
Subtidal rock formations are dispersed throughout the high energy 
coastline. Nesting females approach the beach using strong rip-currents 
through obstruction-free areas (Hughes 1974; Hughes 1996; Botha 2010; 
Nel et al. 2015).
    Foraging areas of the SW Indian DPS include coastal and pelagic 
waters of the SW Indian Ocean and the SE Atlantic Ocean. The DPS is 
somewhat unique in that turtles forage in two ocean basins and do not 
need to undergo long migrations between nesting and foraging areas 
because highly productive foraging areas are available adjacent to 
nesting beaches or connected to nesting beaches via fast-moving 
currents. For example, the warm, fast-flowing Agulhas Current 
(Lutjeharms 2001; Nel et al. 2015) results in high productivity 
foraging areas near nesting beaches and provides a migratory corridor 
to distant foraging areas. As a result, the SW Indian turtles have the 
largest body size, largest clutch size, and highest reproductive output 
of all leatherback turtles (Saba et al. 2015).
    Satellite tracking of post-nesting females (n = 27) reveals the use 
of one of three post-nesting migratory corridors: north into the nearby 
coastal waters of the Mozambique channel; south and west (via the 
Agulhas and Benguela Currents) into the pelagic waters of the South 
Atlantic Ocean; or south and east (via the Agulhas Current and 
Retroflection) into the oceanic eddies in the SW Indian Ocean (Luschi 
et al. 2006; Robinson et al. 2016; Harris et al. 2018). Luschi et al. 
(2006) reviewed satellite telemetry data of 11 post-nesting females 
tagged between 1996 and 2003 (Hughes et al. 1998; Luschi et al. 2003; 
Sale et al. 2006); and Robinson et al. (2016) satellite tracked 16 
post-nesting females tagged between 2011 and 2013. Evaluating tracking 
data for 14 post-nesting females between 2006 and 2014, Harris et al. 
(2018) found that leatherback turtles equally used all three migration 
corridors. In the other studies, a total of 11 post-nesting

[[Page 48374]]

females migrated a relatively short distance (approximately 500 km) to 
the shallow (less than 50 m depth), coastal waters of the Sofala Banks 
(i.e., the Mozambique Channel), where net primary productivity and sea 
surface temperatures remain elevated year-round (n = 4, Sale et al. 
2006; n = 7, Robinson et al. 2016). One post-nesting female migrated to 
the similarly hospitable coastal waters of Madagascar (Robinson et al. 
2016). Ten post-nesting females tracked to pelagic waters of the 
Atlantic Ocean (n = 6, Sale et al. 2006; n = 4, Robinson et al. 2016). 
These waters are among the most productive in the world, as a result of 
strong upwelling (caused by the southeast trade winds) and the area's 
unique bathymetry, hydrography, chemistry, and trophodynamics (Honig et 
al. 2007). Five post-nesting females appeared to track oceanic eddies 
into the SW Indian Ocean (n = 1, Sale et al. 2006; n = 4, Robinson et 
al. 2016). Luschi et al. (2003 and 2006) characterized leatherback 
turtles using this latter strategy as ``wanderers, ranging over vast 
oceanic areas while searching for their planktonic prey.'' 
Opportunistically encountered and highly productive eddies likely 
shaped the circuitous routes of these foraging turtles, which resemble 
drifters more than active swimmers (Luschi et al. 2006; Robinson et al. 
2016; Harris et al. 2018). Thus, this DPS benefits from the use of 
three migratory corridors that all provide highly productive foraging 
opportunities, with minimal energetic cost required to return to waters 
off nesting beaches.

Abundance

    The total index of nesting female abundance of the SW Indian DPS is 
149 females. We based this index on two nesting aggregations: South 
Africa (Ezemvelo KwaZulu-Natal Wildlife (Ezemvelo), unpublished data, 
2018) and Mozambique (Centro Terra Viva Estudos e Advocacia Ambiental 
(CTV), unpublished data, 2018). Our total index does not include two 
unquantified nesting aggregations in Mozambique. To calculate the index 
of nesting female abundance (i.e., 134 females) for the South Africa 
``monitoring area'' (i.e., a 52.8 km stretch of beach that has been 
monitored for decades), we divided the total number of nests between 
the 2014/2015 and 2016/2017 nesting seasons (i.e., a 3-year remigration 
interval; Hughes 1996; Lambardi et al. 2008; Nel et al. 2013; Saba et 
al. 2015) by the clutch frequency (7 clutches/season; Nel et al. 2013; 
Saba et al. 2015). To calculate the index of nesting female abundance 
in Mozambique (i.e., 15 females), we divided the total number of nests 
between the 2015/2016 and 2017/2018 nesting seasons (i.e., a 3-year 
remigration interval) by the clutch frequency for South Africa (7 
clutches/season; Nel et al. 2013; Saba et al. 2015).
    This is an index for the DPS because it only includes available 
data from recently and consistently monitored nesting beaches. While 
nesting occurs on beaches that stretch across 900 km of South Africa 
and Mozambique, consistent and standardized monitoring occurs only 
across approximately 300 km of beaches across the two nations (Nel et 
al. 2013; Nel et al. 2015). Furthermore, while nesting is known to 
occur at low levels at Inhaca Island and Bazaruto Archipelago in 
Mozambique, we did not include these sites because we did not have data 
from the most recent 3 years.
    Other estimates of total or annual nesting female abundance have 
been published. The IUCN Red List assessment estimated the total number 
of mature individuals (males and females) at 148 individuals, based on 
an average of 259 annual nests (Nel et al. 2013), a 3-year remigration 
interval (Nel et al. 2013), and a 3:1 sex ratio (Wallace et al. 2013). 
Their estimates are based on nesting surveys conducted in South Africa, 
which hosts approximately 80 to 90 percent of nesting, and Mozambique 
(Wallace et al. 2013; Nel et al. 2015). Their estimate is less than our 
index, despite including mature males and females. The reason for this 
difference is because they used an average annual number of nests that 
was lower than recent nest counts over the 3-year remigration interval. 
Nel et al. (2015) estimated the size of the total nesting population at 
approximately 100 females per season (Nel et al. 2015), based on 2010 
data: 375 emergences and 336 nests in South Africa; and 61 emergences 
in Mozambique (Videira et al. 2011). This estimate (n = 300, based on a 
3 year remigration interval) is greater than our index because there 
were more nests in 2010 compared to more recent years (2014 to 2016). 
Hamann et al. (2006) estimated approximately 20 to 40 nesting females 
annually in South Africa and approximately 10 nesting females annually 
in southern Mozambique. This estimate (n = 90 to 150, based on a 3 year 
remigration interval) is less than our index, likely as a result of 
using data collected over a different time-frame. The difference in 
estimates likely results from using different methods of calculation 
and different time frames and reflects some uncertainty in the precise 
number of nesting females. Our total index of nesting female abundance 
falls within the range of other estimates and is based on the best 
available data for the DPS at this time.
    There are additional published estimates for the South Africa 
monitoring area. Nel et al. (2013) identified 2,578 nesting females 
over 45 years (1965 to 2009), with a mean of 69.4  38.1 
nesting females per season (or 209 total nesting females) in the 
monitoring area. Hughes (1996) reported an annual average of 24 nesting 
females in the first decade (1976 to 1985) and an annual average of 86 
nesting females in the second decade (1986 to 1995) in the monitoring 
area. Hughes (1996) also reported an annual average of 113 nesting 
females from 1986 to 1995 in an extended protected area that includes 
the monitoring area plus another 93 km in the St. Lucia Marine Reserve, 
which is surveyed periodically. The difference between these two 
averages reflects that most estimates of nesting female abundance in 
South Africa are minimum estimates because nesting occurs outside the 
monitoring area. Thorson et al. (2012) found that annual resightings 
for leatherback turtles decreased from the 1960s to 2009, and their 
modeling indicated that this decline was due to decreased detection 
probabilities (i.e., decreased probability of returning to the 
monitored portion of the KwaZulu-Natal nesting beach), rather than 
decreased survival. Based on satellite tracking of 17 post-nesting 
females, Harris et al. (2015) estimates that approximately 66 percent 
of leatherback nesting activity occurs outside the monitoring area. 
However, considerable inter-annual variability exists, ranging from 
less than 30 percent to over 80 percent, with a median of approximately 
49 percent (Harris et al. 2015). Thus, incomplete beach monitoring is a 
source of uncertainty for this DPS and for our total index of nesting 
female abundance.
    For Mozambique, our index of nesting females is similar to other 
published estimates, which are generally less than 20 nesting females 
(Hamann et al. 2006; Louro 2014; Pereira et al. 2014; Fernandes et al. 
2018). If we use the clutch frequency for Ponta Malongane (2.25 
clutches per season; Louro et al. 2006), which is low for the species, 
our index of nesting female abundance is 45 females. This clutch 
frequency may be underestimated due to females nesting in distant areas 
where monitoring does not regularly occur. If we use the clutch 
frequency for South Africa, (7 clutches/season; Nel et al. 2013; Saba 
et al. 2015), the resulting index of nesting female abundance for 
Mozambique (i.e., 15

[[Page 48375]]

nesting females) is closer to published estimates.
    The total index of nesting female abundance of 149 females places 
the DPS at risk for environmental variation, genetic complications, 
demographic stochasticity, negative ecological feedback, and 
catastrophes (McElhany et al. 2000; NMFS 2017). These processes, 
working alone or in concert, place small populations at a greater 
extinction risk than large populations, which are better able to absorb 
losses in individuals. Due to its small size, the DPS has restricted 
capacity to buffer such losses. Given the intrinsic problems of small 
population size, we conclude that the limited nesting female abundance 
is a major factor in the extinction risk of this DPS.

Productivity

    The SW Indian DPS exhibits a slightly decreasing nesting trend. We 
base our conclusion on data consistently collected in a standardized 
approach in the 56 km South African monitoring area (Ezemvelo, 
unpublished data, 2018), where nest counts decreased by -0.3 percent 
annually (sd = 2.1 percent; 95 percent CI = -4.5 to 4.1 percent; f = 
0.557; mean annual nests = 301) between the 1973/1974 and 2016/2017 
nesting seasons. The trend in South Africa is likely representative of 
the entire DPS, as 80 to 90 percent of nesting is estimated to occur 
there (Wallace et al. 2013; Nel et al. 2015) and the 44-year time 
series is quite robust.
    Our trend estimates yield similar results to other published 
findings for the population. The IUCN concluded that this population 
has declined slightly, by 5.6 percent over the past three generations, 
with an annual decline of -0.1 percent in South Africa and -0.7 percent 
in Mozambique (Wallace et al. 2013). Hamann et al. (2006) also 
identified a declining trend in the nesting population of the SW Indian 
Ocean. Studies focused on the South African monitoring area (i.e., the 
source of data for our trend analysis), however, disagree on the 
whether the trend has declined recently (Hamann et al. 2006; Nel et al. 
2013) or is stable (Nel et al. 2015; Saba et al. 2015). The nest trend 
may be stable if nesting in unmonitored areas has increased over time 
(Thorson et al. 2012; Harris et al. 2015). Different datasets lead to 
different conclusions due to different methods of calculation, 
different time frames, incomplete monitoring of all nesting areas, and 
therefore uncertainty in the precise number of nesting females. We find 
that Nel et al. (2013) provide the best available published data, which 
are based on the most recent, primary data, and we agree with their 
characterization of the trend as declining or recently declining.
    Despite the recent decline in nesting, productivity parameters 
remain relatively high for the SW Indian DPS, which has the largest 
body size, largest clutch size, and highest reproductive output of all 
leatherback turtles, likely due to the close proximity between their 
nesting beaches and highly productive foraging areas (Saba et al. 
2015). Nel et al. (2015) reports that most metrics (i.e., female size, 
egg size, incubation time, and hatching success) are above average for 
this DPS. Nesting females produced 1,171 to 53,139 hatchlings each 
season in the South Africa monitoring area between 1965 and 2009, with 
an average of 36,583 to 51,610 hatchlings per season, which was 
calculated by multiplying 480 hatchlings per nesting female by 69.4 
 38.1 nesting females per season (Nel et al. 2013).
    The recent nesting decline may reflect the effects of past and 
current threats that overwhelm the population's high productivity 
metrics. We conclude that the slightly declining nest trend places the 
DPS at risk of extinction, which is further exacerbated by the limited 
nesting female abundance.

Spatial Distribution

    The SW Indian DPS comprises, in essence, a single nesting 
aggregation, with nesting females moving freely between South African 
and Mozambican beaches (Hughes 1996; Luschi et al. 2006; Nel et al. 
2015). Nesting is limited to a total distance of approximately 900 km 
along South African and Mozambican coasts (Nel et al. 2015). While 80 
to 90 percent of nesting is concentrated in South Africa, nesting is 
somewhat concentrated in the southern section of the South African 
monitoring area, although most characterize nesting as low density 
throughout South Africa (Hughes 1974; Lambardi et al. 2008; Botha 2010; 
Nel et al. 2013; Harris et al. 2015; Nel et al. 2015).
    The DPS exhibits a broad foraging range that extends into coastal 
and pelagic waters of the eastern Atlantic and western Indian Oceans 
(Luschi et al. 2006; Lambardi et al. 2008; Girondot 2015). There is 
limited evidence that leatherback turtles may remain in South African 
waters throughout the year, as suggested by year-round fisheries 
bycatch records (Luschi et al. 2003, 2006; Petersen et al. 2009). Some 
forage off the coast of Madagascar (Robinson et al. 2016; Harris et al. 
2018). Some turtles follow the Agulhas and Benguela Currents into 
foraging areas in the southeast Atlantic Ocean, off the coasts of 
Angola and Namibia (Girondot 2015; Robinson et al. 2016; Harris et al. 
2018). Others follow the Agulhas Retroflection and deep-sea eddies into 
the SW Indian Ocean (Luschi et al. 2006; Lambardi et al. 2008; Robinson 
et al. 2016; Harris et al. 2018). Leatherback turtles, possibly from 
this DPS, have also been observed in the Red Sea, presumably foraging 
(Hamann et al. 2006). The use of various foraging areas may be 
influenced by the prevalent currents encountered off the nesting 
beaches (Luschi et al. 2006; Lambardi et al. 2008; Robinson et al. 
2016).
    The wide distribution of foraging areas likely buffers the DPS 
somewhat against local catastrophes or environmental changes that would 
limit prey availability. Nesting occurs along one coastline, which is 
3,000 km in length and may be similarly affected by environmental 
variation and directional changes (e.g., sea level rise). Because the 
DPS is essentially a single nesting aggregation, it has limited 
capacity to withstand other catastrophic events. Thus, spatial 
distribution likely has little net effect on the extinction risk of the 
SW Indian DPS.

Diversity

    Within the SW Indian DPS, genetic diversity is low, with only two 
mtDNA haplotypes found in 41 nesting females in South Africa (haplotype 
diversity = 0.298  0.078 and nucleotide diversity = 0.0004 
 0.0004; Dutton et al. 2013). Nesting habitat is mainly 
restricted to beaches along the same coast, with a few nests on 
Mozambican islands. The DPS does not exhibit temporal or seasonal 
nesting diversity, with most nesting occurring between October and 
March. The foraging strategies are diverse, however, with turtles using 
coastal and pelagic waters in the Atlantic and Indian Oceans. Diverse 
foraging strategies may provide some resilience against local 
reductions in prey availability or catastrophic events, such as oil 
spills, by limiting exposure. Low genetic diversity indicates the DPS 
may lack the raw material necessary for adapting to long-term 
environmental changes, such as cyclic or directional changes in ocean 
environments due to natural and human causes (McElhany et al. 2000; 
NMFS 2017). We conclude that limited overall diversity increases the 
extinction risk of this DPS by reducing its resilience to threats.

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    Coastal erosion, foot and vehicle traffic, and artificial lighting 
modify the available, suitable nesting habitat and

[[Page 48376]]

thus are threats to the SW Indian DPS. Angel et al. (2014) identifies 
coastal erosion as the main beach-based threat to this population and 
one that is likely to increase with climate change.
    Coastal erosion removes sand from nesting beaches, inundating nests 
and destroying eggs. Because leatherback turtles nest lower on the 
beach than other sea turtles, they have greater exposure to tidal 
erosion and deposition (Boyes et al. 2010). At South African nesting 
beaches over a duration of 70 days, Boyes et al. (2010) found an 
average of 0.62 m deposition (S.D. 0.15 m; range 0.34-0.85 m) and 0.42 
m erosion (S.D. 0.17 m; range 0.14- 0.71 m). Because the average depth 
of leatherback nests was 0.66 m (S.D. 0.19 m; range 0.15-1.07 m), eggs 
are at some risk of being exposed and destroyed (Boyes et al. 2010). 
Nel et al. (2006) concludes that coastal erosion is a threat in South 
Africa, where the high-energy coastline varies seasonally. During two 
nesting seasons (2009/2010 and 2010/2011), de Wet (2012) found that 6.3 
percent of nests in the South African monitoring area were destroyed by 
erosion. In Bazaruto Archipelago, Mozambique, coastal erosion and 
rising sea levels destroyed approximately 12 percent of nests over 10 
seasons of monitoring (Videira and Louro 2005; Louro 2006). Despite 
nest loss due to erosion, hatching success remains high in South Africa 
(70 to 80 percent; Nel et al. 2015; Santidri[aacute]n Tomillo et al. 
2015). Though the introduction of Casuarina trees do not necessarily 
increase the risk of erosion, they obstruct nesting females' access to 
and from beaches and alter nest incubation environments (de Vos et al. 
2019). Evolving in a high-energy coastline environment with seasonal 
variation has likely provided the DPS with some resilience to nesting 
losses due to coastal erosion. Sea level rise as a result of climate 
change, however, is likely to increase the rate and magnitude of this 
natural process.
    In Mozambique, Louro (2006) describes beach driving as a ``very 
serious problem.'' Tourism and beach driving are increasing in Ponta 
Malongane and Bazaruto Island, nesting areas in Mozambique, where there 
is no legislation regarding beach driving (Louro 2006). Foot and 
vehicular traffic, for tourism and recreational purposes, have been 
found to impact nesting beach habitat and turtles in several ways. 
Beach activities can deter females from using a nesting beach. Beach 
driving causes sand compaction, which may lower nest success. It also 
creates ruts that slow hatchlings' crawl to the surf, increasing their 
vulnerability to predators. Beach driving occurs to a lesser extent in 
South Africa. Recreational beach driving is allowed on a 1.5 km stretch 
of beach, and tourism driving (for concession, management, and media) 
involves a maximum of 10 vehicles per night across 40 km of beach (Nel 
2006).
    Artificial lighting modifies the quality of nesting beaches because 
lights over land disorient nesting females and hatchlings. Instead of 
crawling toward the surf and their marine habitat, they crawl further 
inland, where they may become dehydrated and die or become susceptible 
to predation. Within the 280 km of coastline within the iSimangaliso 
Wetland Park, South Africa, there are only four areas of less than 100 
m each that contain artificial lighting (Nel 2006). We were unable to 
find data on artificial lighting in Mozambique.
    The majority of nesting habitat occurs within the 280 km coastline 
of the iSimangaliso Wetland Park in South Africa, which has been a 
World Heritage Site since 1999 (UN Educational, Scientific and Cultural 
Organization 1999; Hughes 2010; Robinson et al. 2016). From 1979 to 
1999, much of the nesting habitat and nearshore marine habitat was 
protected, first as the St. Lucia Marine Reserve, then the Maputaland 
Marine Reserve (Hughes 1996). Such protections contributed to the 
prevention of dredging a deep water harbor through turtle nesting 
beaches and mining heavy minerals in the adjacent dunes (Hughes 2009, 
2010). In Mozambique, the Ponta do Ouro Partial Marine Reserve has 
provided beach and marine habitat protection since 2009. Additional 
protection is provided to Mozambican nesting beaches in: The Ponto du 
Ouro--Kosi Bay Transfrontier Marine Conservation Area; the Maputo 
Special Reserve; the Bazaruto Archipelago National Park; and the 
Quirimbas Archipelago National Park. However, nest protection only 
occurs over nine percent of the Mozambique coastline (Videira et al. 
2008; Garnier et al. 2012). Such protections have minimized vehicular 
traffic at nesting beaches in South Africa, but beach driving remains a 
threat in Mozambique. Erosion is a threat to nesting beaches in both 
South Africa and Mozambique. Thus, we conclude that the present 
modification of nesting habitat is a threat to the SW Indian DPS.

Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    Overutilization is a threat to the SW Indian DPS (Bourjea 2015; 
Williams et al. 2016; Williams 2017). Two of nine leatherback turtles 
equipped with satellite tags between 1996 and 2006 were incidentally or 
intentionally captured in Mozambique and Madagascar and likely retained 
for food or sale (de Wet 2012). In Mozambique, eggs and turtles were 
once legally harvested and are now illegally poached for consumption 
(Nel 2012; Wallace et al. 2013; Fernandes et al. 2018). Turtle poaching 
includes turtles taken on the beaches and at sea (Williams et al. 2016; 
Williams 2017). We do not have recent, quantitative estimates of egg or 
turtle poaching in Mozambique. However, significant usage has been 
documented at various points in time. Hughes (1995) reported that 
nearly every nesting female was killed during the civil war (1977 to 
1992). An estimated 32 loggerhead and leatherback turtles were killed 
at Ponta Malongane in 11 years (Louro 2006). Recent egg and turtle 
poaching rates in Mozambique have been qualitatively described as 
``alarming,'' ``significant,'' ``widespread,'' ``prominent,'' and 
``prevalent'' (Fernandes et al. 2015; Williams et al. 2016; Williams 
2017; Pereira and Louro 2017; Fernandes et al. 2017; Fernandes et al. 
2018). Nest monitoring programs in Mozambique have provided some 
protection since the 1990s (Garnier et al. 2012). Pereira et al. (2014) 
reports that as a result of the monitoring program at the Ponta do Ouro 
Partial Marine Reserve, where the majority of nesting in Mozambique 
occurs, turtle mortalities are very rare. Egg poaching has been reduced 
in the Bazaruto Archipelago, where it was previously prevalent (Louro 
2006). National legislation in Mozambique include: Diploma Legislativo 
2627 (7 August 1965), Forest and Wildlife Regulation (Decree 12/2002 of 
6 June 2002) and Conservation Law (Law 5/2017 of 11 May). These laws 
protect turtles and eggs and impose fines for poaching or possession. 
However, the laws are poorly implemented and enforced (Costa et al. 
2007; Louro 2006; Williams et al. 2016; Fernandes et al. 2018). We 
conclude that the poaching of turtles and eggs remains a significant 
threat in Mozambique.
    Poaching of turtles is also a threat in Madagascar, where 
leatherback turtles caught in gillnets are taken back to local villages 
and consumed, which is documented to have occurred twice in 2016 
(Williams 2017). Leatherback turtles were caught and consumed or sold 
in five of eight Malagasy villages surveyed between October 2004 and 
March 2004. Fishers reported that leatherback turtles were uncommon but 
large, possibly indicative of mature individuals (Walker and Roberts 
2005).

[[Page 48377]]

No leatherback turtles were reported caught during a 2007 Malagasy 
village survey (Humber et al. 2010). Although protected by Presidential 
Decree (2006-400), fishers target turtles at sea for consumption 
(Ratsimbazafy 2003; Epps 2006; Humber et al. 2010). Humber et al. 
(2010) report that the Malagasy law is not adequately implemented due 
to lack of enforcement, a reluctance to manage the local, cultural 
fishery, and the size of the coastline (Rakotonirina and Cooke 1994; 
Okemwa et al. 2005). We conclude that the poaching of turtles remains a 
significant threat in Madagascar.
    Egg and turtle poaching does not appear to be a significant threat 
in South Africa. Prior to the ban on egg harvest in 1963, substantial 
numbers of leatherback eggs in South Africa were harvested, likely 
contributing to the critically low number of nesting females at that 
time (Nel et al. 2015). Hughes et al. (1996) concluded that nesting 
females were not harvested. As a result of the ban, and with a 
lucrative tourism industry centered on the nesting turtles, egg and 
turtle harvest has been nearly eliminated (Hughes et al. 1996). Nesting 
females and hatchlings receive ``intensive and effective'' protection, 
as most nesting beaches fall within the iSimangaliso Wetland Park (Nel 
et al. 2015). Such beach protections have been key to recovering the 
number of nesting females to current levels (Hughes et al. 1996; Saba 
et al. 2015; Nel et al. 2015). We conclude that the poaching of turtles 
and eggs is not a significant threat in South Africa.
    Exposure to poaching is low in South Africa, where the majority of 
females nest. Few females nest in Mozambique, reducing the DPS's 
overall exposure to egg and nesting female poaching during nesting. 
However, turtles regularly forage in the Mozambique Channel, where they 
may be poached along the coasts of Mozambique and Madagascar. Poaching 
of nesting females or post-nesting females (i.e., on land or at sea) 
reduces both abundance (through loss of nesting females) and 
productivity (through loss of reproductive potential). Such impacts are 
high because they directly remove the most productive individuals from 
DPS, reducing current and/or future reproductive potential. Egg 
poaching reduces productivity. We conclude that overutilization, as a 
result of poaching of turtles and eggs, poses a threat to the DPS.

Disease or Predation

    While we could not find any information on disease for this DPS, 
predation is a threat to the SW Indian DPS. In South Africa, nest 
predators include feral dogs, side-striped jackals, honey badgers, and 
ghost crabs (Hughes 1996; Nel 2006). In the 1960s, the removal of feral 
dogs greatly reduced nest predation. Similarly, jackals were once a 
threat (Hughes 1996). However, nest predation by jackals has not been 
observed for 17 years (R. Nel, pers. comm. April 15, 2019). Nel (2006) 
reports current rates of predation as relatively low. Nel et al. (2013) 
reports that there is no evidence for significant beach predation on 
South African beaches. Describing nest predation as minimal in South 
Africa, de Wet (2012) found that 15.7 percent of nests were depredated 
in the 2009/2010 and 2010/2011 nesting seasons; ants and ghost crabs 
were the main cause of egg mortality. During the two seasons, ghost 
crabs consumed 3.2 percent of hatchlings as they made their way to the 
sea (de Wet 2012).
    While all eggs and hatchlings have some exposure to predation, the 
species compensated for a certain level of natural predation by 
producing a large number of eggs and hatchlings. For this DPS, the 
primary impact is to productivity (i.e., reduced egg and hatching 
success). We conclude that, though much reduced, predation still poses 
a threat to the SW Indian DPS.

Inadequacy of Existing Regulatory Mechanisms

    The SW Indian DPS is protected to some degree by several regulatory 
mechanisms. For each, we review the objectives of the regulation and to 
what extent it adequately addresses the targeted threat.
    Despite efforts to reduce impacts, fisheries bycatch continues to 
be the primary threat to this DPS (Petersen et al. 2009; Nel et al. 
2013; Wallace et al. 2013; Fossette et al. 2014; Angel et al. 2014; Nel 
et al. 2015; Harris et al. 2018). To minimize the impacts from longline 
fisheries, the FAO published guidelines for sea turtle protection, 
entitled Technical Consultation on Sea Turtle-Fishery Interactions (FAO 
2004; Huang and Liu 2010). The UN 1995 Code of Conduct for Responsible 
Fisheries (FAO 2004) provides guidelines for the development and 
implementation of national fisheries policies, including gear 
modification (e.g., circle hooks, fish bait, deeper sets, and reduced 
soak time), new technologies, and management of areas where fishery and 
sea turtle interactions are more severe. The guidelines stress the need 
for mitigation measures, data on all fisheries, fishing industry 
involvement, and education for fishers, observers, managers, and 
compliance officers (FAO 2004; Honig et al. 2007). These guidelines, 
however, are rarely enacted in full. The ICCAT has adopted a resolution 
for the reduction of sea turtle mortality (Resolution 03-11), 
encouraging States to submit data on sea turtle interactions, release 
sea turtles alive wherever possible, and conduct research on mitigation 
measures. The responsibility to implement mitigation measures remains 
within each nation, and many nations have not implemented such measures 
(Honig et al. 2007). South Africa, Namibia, and Angola signed the 
Memoranda of Understanding concerning Conservation Measures for Marine 
Turtles of the Atlantic Coast of Africa. Though South African vessels 
are required to carry a dehooker and line-cutter (Honig et al. 2007) 
and has instituted an observer program (Petersen et al. 2009), few 
other at-sea conservation measures have been implemented (Honig et al. 
2007). For Taiwanese fishing vessels operating within the range of this 
DPS, Taiwan has regulations to limit the number of vessels in the area 
and to require vessels to carry de-hookers. However, bycatch and 
mortality remain high (Huang and Liu 2010). Similarly, though the 
extent of shark nets off South African beaches has been reduced from 44 
km in the early 1990s to 23 km in 2007, bycatch and mortality continue 
to occur (Brazier et al. 2012), and Nel et al. (2015) identify bather 
protection nets, together with boat strikes, as the second greatest 
threat to the DPS, after longline fisheries. Regarding shark nets, 
Brazier et al. (2012) concludes that bycatch is low and rates are 
stable, but because the leatherback population is small, a further 
reduction in bycatch is desirable. Because the offshore longline 
fishery contributes more than the shark nets to leatherback mortality, 
Brazier et al. (2012) also recommends further introduction of bycatch 
reduction techniques in the longline fishery. Because longline threats 
are proportionally large and possibly increasing, Harris et al. (2018) 
concludes that bycatch mitigation measures in this industry remain 
first and most important management action. Thus, existing regulations 
have been inadequate to meet their objectives.
    Beach habitat is protected throughout a portion of the nesting 
range of this DPS. In South Africa, approximately 280 km of nesting 
beaches benefit from intensive and effective protection as part of the 
iSimangaliso Wetland Park, a World Heritage Site since 1999 (UN 
Educational, Scientific and Cultural Organization 1999; Nel et al. 
2015). iSimangaliso includes 280 km of beaches, rocky shores, 
mangroves, lakes,

[[Page 48378]]

estuaries, and coastal waters out to three nautical miles (5 km) and 
200 m depth. Regulations prevent coastal development and commercial 
fishing within this area. However, Harris et al. (2015) estimated that 
66 percent of leatherback turtles nest outside of the protected 
monitoring area (i.e., only 300 km of the 900 km nesting area is 
monitored and protected). In addition, leatherback turtles use coastal 
waters that are not protected under the marine reserve. In Mozambique, 
much of the nesting habitat is protected, including: The Ponto du 
Ouro--Kosi Bay Transfrontier Marine Conservation Area; the Maputo 
Special Reserve; the Bazaruto Archipelago National Park; and the 
Quirimbas Archipelago National Park. However, nest protection only 
occurs over nine percent of the Mozambique coastline (Videira et al. 
2008; Garnier et al. 2012). Thus, regulations to protect the nesting 
habitat of the DPS have been successful. However, leatherback turtles 
nesting outside these areas receive no protection.
    In addition, South Africa hosts several marine protected areas and 
has proposed to add 20 new marine protected areas to expand protection 
to five percent of its EEZ (https://www.marineprotectedareas.org.za/). 
Two of these were proposed in order to protect leatherback marine 
habitat: The 1200 km\2\ iSimangaliso Marine Protected Area (off nesting 
beaches); and the 6200 km\2\ Agulhas Front Marine Protected Area 
(encompassing core foraging habitat). These initiatives are likely to 
protect leatherback turtles within the proposed areas. However, the DPS 
has a large range that extends well beyond protected areas. Harris et 
al. (2018) identifies the Mozambique Channel as an additional key 
priority area to protect.
    In South Africa, a 1963 ban on egg and turtle harvest has been 
effective in virtually eliminating overutilization (Hughes 1996). The 
current law, Regulation 58(7) of the MLRA (1998), provides full 
protection to sea turtles and their products. In Mozambique, national 
legislation includes: Diploma Legislativo 2627 (7 August 1965), Forest 
and Wildlife Regulation (Decree 12/2002 of 6 June 2002) and 
Conservation Law (Law 5/2017 of 11 May). These laws protect turtles and 
eggs and impose fines for poaching or possession. For example, the 
Forest and Wildlife regulation prohibits the killing of turtles and the 
possession of their eggs, with fines up to US $1,000 (Decree 12/2002 of 
6 June 2002; Costa et al. 2007). In 2008, there were at least 13 
conservation programs focusing on protection and education. Despite 
these efforts, illegal poaching of eggs and turtles remains prevalent 
in Mozambique (Fernandes et al. 2014) due to limited implementation and 
enforcement of the environmental legislation (Costa et al. 2007; Louro 
2006; Williams et al. 2016; Fernandes et al. 2018). In Madagascar, all 
sea turtles are protected from exploitation by Presidential Decree 
(2006-400). However, fishers continue to target and consume turtles 
captured at sea (Ratsimbazafy 2003; Epps 2006; Humber et al. 2010). The 
effectiveness of the Malagasy law is limited due to lack of 
enforcement, a reluctance to manage the local, cultural fishery, and 
the size of the coastline (Rakotonirina and Cooke 1994; Okemwa et al. 
2005; Humber et al. 2010). Thus, while regulations to prevent the 
harvest of turtles and eggs have been adequate in South Africa, 
regulatory protections in Mozambique and Madagascar are inadequate.
    In summary, numerous regulatory mechanisms protect leatherback 
turtles, eggs, and nesting habitat throughout the range of this DPS. 
Though the regulatory mechanisms provide some protection to the 
species, many do not adequately reduce the threat that they were 
designed to address, generally as a result of limited implementation or 
enforcement. As a result, bycatch, incomplete nesting habitat 
protection, and poaching in Mozambique and Madagascar remain threats to 
the DPS. In summary, we consider the inadequacy of the regulatory 
mechanisms to be a threat to the SW Indian DPS.

Fisheries Bycatch

    Fisheries bycatch is the primary threat to the SW Indian DPS 
(Wallace et al. 2013; Fossette et al. 2014; Angel et al. 2014; Nel et 
al. 2015; Harris et al. 2018). Bycatch occurs in commercial and 
artisanal, coastal and pelagic fisheries. Gear types include: Longline, 
purse seine, pelagic trawl, shrimp trawl, gillnets, and beach seines 
(Honig et al. 2007; Petersen et al. 2009; Nel et al. 2013; Nel et al. 
2015).
    Of all gear types, longline fisheries likely have the largest 
impact on the DPS (Petersen et al. 2009; Nel et al. 2013; Angel et al. 
2014; Nel et al. 2015; Harris et al. 2018). Leatherback turtles are 
exposed to longline fisheries throughout their foraging range, 
including the Benguela Current in the Atlantic Ocean, the Agulhas 
Current in the Indian Ocean, and coastal waters off South Africa, 
Mozambique, and Madagascar (Honig et al. 2007; Peterson et al. 2009; 
Huang and Liu 2010; Harris et al. 2018). Flag states include: South 
Africa, Mozambique, Japan, and Taiwan (Honig et al. 2007; Peterson et 
al. 2009; Huang and Liu 2010).
    Harris et al. (2018) found a positive, significant relationship 
between the longline fisheries' extent of overlap with leatherback 
migratory corridors and threat intensity (F1,8 = 184.7, P 
<0.001, R2 = 0.95), which was defined as a product of the turtles 
utilization distribution and the normalized fishing effort. They 
concluded that incidental capture in longline fisheries was the most 
important offshore threat to leatherbacks and supports the hypothesis 
that longlining is suppressing growth of this DPS (Nel et al. 2013; 
Harris et al. 2018). Harris et al. (2018) calculated longline bycatch 
rates, around Southern Africa, to be 1,500 leatherback turtles 
annually. Though this estimate likely includes turtles from other DPSs 
(SE Atlantic and NE Indian), the authors concluded that even low 
absolute bycatch has a disproportionately large effect in slowing 
population growth rates, due to the small nesting female abundance of 
the SW Indian DPS (Harris et al. 2018). Additional reason for concern 
is that the threat intensity of longlining was especially high in the 
last 5 years of the study (ICCAT and IOTC data from 2004 to 2013), 
suggesting that the threat and its impacts on the DPS are increasing 
(Harris et al. 2018). Throughout the SE Atlantic and SW Indian Oceans, 
Harris et al. (2018), Wallace et al. (2013), deWet (2012), Thorson et 
al. (2012), and Peterson et al. (2009) analyze longline bycatch over a 
large portion of the DPS's foraging range. Wallace et al. (2013) 
categorize the longline fishing effort as medium to high and conclude 
that such effort leads to a high risk and high bycatch impact for the 
SW Indian DPS. Thorson et al. (2012) used data from the IOTC (1954 to 
2009) and South African fishery (2006 to 2009) in a model of 
leatherback turtle survival and availability. Their model did not find 
that leatherback survival declined during the period when longline 
fishing effort increase. However, the authors state that their null 
result could be explained by an imprecise index of longline effort or 
using newer bycatch rates for the South African longline fishery (i.e., 
Petersen et al. 2009). For example, based on fisheries data from 30 
South African and Asian pelagic longline vessels operating in the South 
African EEZ between 2006 and 2010, De Wet (2012) estimates the mean 
annual bycatch to be 7.8 (7.8 S.D.) leatherback turtles, 
based on 39 leatherback turtle captures reported over 5 years. Other 
studies estimate bycatch to be higher. Based on extrapolations from

[[Page 48379]]

independent observer bycatch reports from 1998 to 2005 (n = 2,256 
sets), Peterson et al. (2009) estimates that the South African pelagic 
longline fishery for tunas and swordfish captures 50 leatherback 
turtles annually, many of which likely belong to the SW Indian DPS (the 
remainder belong to the SE Atlantic DPS). Though most (84 percent) were 
caught alive, Peterson et al. (2009) estimates the long-term survival 
of affected turtles at 50 percent (based on an estimated range of 25 to 
75 percent; Aguilar et al. 1995). Peterson et al. (2009) thus estimates 
total mortality from the South African pelagic longline fishery to be 
25 turtles annually, or around two percent of the total population 
(based on a total population size of 1,200 leatherback turtles), which 
they conclude is enough to hamper recovery of the SW Indian population. 
Nel et al. (2013) agrees with this conclusion, citing a 30 year (1965 
to 1995) increasing trend in nesting female abundance that stalled as 
the longline fishery expanded from 1990 to 1995. Huang and Liu (2010) 
come to a similar conclusion. They report that the longline fishery 
operated at a relatively low level until 1995, when South Africa, 
Japan, and Taiwan started a joint venture fishing program.
    In the Indian Ocean, Huang and Liu (2010) evaluated the Taiwanese 
longline fishery bycatch, and Louro (2006) described illegal longlining 
in Mozambique waters. Huang and Liu (2010) evaluated observer data from 
77 trips (4,409 sets) on Taiwanese large-scale longline fishing 
vessels. They identified 84 leatherback turtles captured from 2004 to 
2008, with 48 mortalities (57 percent; Huang and Liu 2010). 
Extrapolating to the entire Taiwanese longline fishery in the Indian 
Ocean, they estimated an average bycatch of 173 leatherback turtles 
between 2004 and 2007. This number likely included individuals from the 
SW and NE Indian DPSs. In addition to commercial longlining, artisanal 
longlining also occurs in the SW Indian Ocean. Illegal longlining off 
Mozambique targets sharks and leatherback turtles. The level of take 
and mortality is unknown. A program called Eyes on the Horizon reports 
such events, when observed (Louro 2006).
    In the SE Atlantic Ocean, Honig et al. (2007) and Angel et al. 
(2014) evaluate longline bycatch. Honig et al. (2007) evaluated turtle 
bycatch by longline fisheries in the Benguela Large Marine Ecosystem by 
using data from observer reports, surveys, and specialized trips from 
the coastal nations of South Africa, Namibia and Angola. They estimated 
bycatch at 672 leatherback turtles annually (based on an annual bycatch 
estimate of 4,200 turtles, of which approximately 16 percent are 
leatherback turtles) in the southern and central regions and as many as 
5,600 leatherback turtles (based on an annual bycatch estimate of 
35,000 turtles) for the entire Benguela Large Marine Ecosystem (Honig 
et al. 2007). These estimates likely include many leatherback turtles 
from the much larger SE Atlantic DPS, but telemetry studies indicate 
that the turtles of the SW Indian DPS use this foraging area too 
(Luschi et al. 2006; Robinson et al. 2016). Evaluating ICCAT data, 
Angel et al. (2014) confirms exposure to high longline fishing effort 
but reports that bycatch of this population is low relative to other 
leatherback populations. Although Thorson et al. (2012) found that 
increased fishing effort had no explanatory power regarding changes in 
leatherback survival, other studies identify longline fisheries as the 
primary threat to the DPS (Petersen et al. 2009; Nel et al. 2013; Angel 
et al. 2014; Nel et al. 2015; Harris et al. 2018). Based on the weight 
of evidence, we agree with the latter and conclude that longline 
fisheries pose a major threat to the DPS throughout its foraging range.
    Other fisheries also impact the SW Indian DPS, possibly resulting 
in substantial mortalities. However, these fisheries are not as well 
studied, and mortality estimates are not available (Honig et al. 2007; 
Nel et al. 2013). Leatherback turtles are caught in artisanal and 
commercial shrimp trawl, pelagic trawl, gillnet, purse seine, and beach 
seine fisheries (Honig et al. 2007; Petersen et al. 2009; Nel et al. 
2013). Citing Walker (2005) and Rakotonirina (1994), Nel (2013) reports 
that the number of sea turtles (all species) caught in artisanal 
fisheries of the Mozambique Channel could exceed commercial fishery 
catches. Honig et al. (2007) echoes this concern for the Benguela 
Current Large Marine Ecosystem, citing high mortality rates for these 
fisheries in other regions. The Mozambican shrimp trawl fishery 
operates in the Sofala Bank of the Mozambique Channel, near leatherback 
nesting, migrating, and foraging areas (Luschi et al. 2006; Robinson et 
al. 2016). The fishery supports 50 to 96 vessels that employ standard 
otter trawl nets in a single or quad-net configuration with an average 
tow-time of three hours (Brito 2012). It does not employ TEDs and 
incidentally captures several (i.e., at least two to six but possibly 
many more) leatherback turtles annually (Louro 2006; Videira et al. 
2010; SWOT 2017). In 2001, one shrimp trawler captain reported 
capturing more than six leatherback turtles since fishing season 
opened; all were captured alive (Gove et al. 2001). Based on 39 
interviews with observers, enforcement officers, and vessel operators, 
the fleet (n = 50) captures approximately 56 (40) 
leatherback turtles; the overall estimated mortality rate for bycaught 
turtles is 14 percent (Brito 2012). Given the overlap between the 
fishery and an important foraging area, M. Pereira (CTV, pers. comm., 
2019) concludes that the Mozambican shrimp trawl fishery may be one of 
the main threats to this DPS. The South African shrimp trawl fishery 
has been reduced to two vessels, with an average annual bycatch of less 
than one leatherback (Honig et al. 2007; Petersen et al. 2009; Nel et 
al. 2013). Domestic shrimp trawling in Eritrea is considered a major 
threat to sea turtles, and bycatch is underreported. However, 
leatherback turtles are relatively rare in these waters, as 
demonstrated by the foreign trawl fleet, which has 100 percent observer 
coverage and bycatch records indicating 39 leatherback turtles between 
1996 and 2005 (Pilcher et al. 2006).
    During a small random sampling exercise in 2013 by onboard 
observers from the Research Division of Eritrea, one leatherback turtle 
(of 48 sea turtles total) was captured and released (Mebrahtu 2015). On 
June 20, 2019, the European Union passed a regulation (PE-CONS 59/1/19 
Rev 1) that requires shrimp trawl fisheries to use a turtle excluder 
device in European Union waters of the Indian and West Atlantic Oceans.
    Gillnets in Macaneta, Mozambique, killed two leatherback turtles 
during the 2010 nesting season (Videira et al. 2010) and captured one 
in the 2003 nesting season (Louro 2006). In Madagascar, leatherback 
turtles are a ``common'' bycatch of the set gillnet shark fishery 
(Robinson and Sauer 2013); mortality is likely high given the 24-hour 
soak time and propensity for consuming turtle meat. Purse seine 
fisheries have a much lower impact than longline fisheries (Angel et 
al. 2014); two leatherback turtles were captured (alive) between 1995 
and 2010 in the Indian Ocean (Clermont et al. 2012). In the EEZ of all 
Indian Ocean French Territories (mostly from the Mozambique Channel), 
40 leatherback turtles were captured in unspecified fisheries from 1996 
to 1999; 92 percent were released alive (Ciccione 2006).
    Shark or bather nets, which are gillnets installed off beaches in 
South Africa to limit human-shark interactions, incidentally capture

[[Page 48380]]

leatherback turtles. According to Nel et al. (2015), bather protection 
nets and boat strikes together present the second greatest threat to 
the DPS, after fisheries. Three of nine leatherback turtles equipped 
with satellite tags between 1996 and 2006 were caught in shark nets (de 
Wet 2012). Between 1981 and 2008, 150 leatherback turtles were captured 
(mean = 5.36; SE = 0.60), of which 20 were mature females and 39 were 
mature males (Brazier et al. 2012). Total mortality was 62.7 percent, 
with an annual range of 1 to 12 mortalities (mean = 3.4; SE = 0.47; 
Brazier et al. 2012). Most turtles were captured in December, the peak 
month for nesting, which together with the prevalence of mature 
individuals, suggests that bycatch is dominated by adults from nearby 
nesting and breeding areas (Brazier et al. 2012). Analyzing these data 
over an additional 2 years (1981 to 2010), de Wet (2012) found that 157 
leatherback turtles (mean = 5.26; SD = 2.7) were captured in the nets, 
with a 62.4 percent mortality rate (mean = 3.3; SD = 1.8).
    To reduce bycatch mortality in longlines, South African regulations 
require vessels to carry a dehooker and line cutter (Honig et al. 
2007). To reduce bycatch in the shark nets, effort was reduced from 44 
km of nets in the early 1990s to 23 km in 2007 (Brazier et al. 2012). 
Despite these efforts, a previously increasing trend in nesting female 
abundance has stalled and ``declined recently'' (Nel et al. 2013).
    Individuals (immature and adult turtles) of this DPS are exposed to 
high fishing effort throughout their foraging range. Estimates of 
bycatch rates, when available, range considerably. For example, Harris 
et al. (2018) estimated the annual longline bycatch rates around 
Southern Africa to be 1,500 leatherback turtles annually; whereas, de 
Wet (2012) estimated the mean annual bycatch to be 7.8 (7.8 
S.D.) leatherback turtles. We have annual mortality estimates for few 
individual fisheries: n = 25 for South African longline (Peterson et 
al. 2009); n = 12 for Taiwanese longline (Huang and Liu 2010); n = 1 to 
12 for shark nets (Brazier et al. 2012). Adding in other longline 
fisheries and additional gear types may result in more than 100 
mortalities annually. These estimates likely include individuals from 
other DPSs (i.e., the SE Atlantic and NE Indian). However, because of 
the small nesting population, even small levels of mortality have the 
potential to slow population growth (Harris et al. 2018). Mortality 
reduces abundance, by removing individuals from the population; it also 
reduces productivity, when potential nesting females are killed. 
Several studies conclude that bycatch has prevented continued 
population growth and/or contributed to the recent slight decline in 
nesting (Petersen et al. 2009; Huang and Liu 2010; Brazier et al. 2012; 
Nel et al. 2013; Harris et al. 2018). We conclude that fisheries 
bycatch is the primary threat to the SW Indian DPS.

Vessel Strikes

    Vessel strikes are a threat to the SW Indian DPS. According to Nel 
et al. (2015), vessel strikes and bather protection nets together 
present the second greatest threat to the DPS, after fisheries. 
Together these threats kill up to 10 leatherback turtles annually (Nel 
et al. 2015). One of 24 leatherback turtles stranded along the South 
African coastline between 1972 and 2010 was struck by a boat propeller 
(Nel 2008). However, additional mortalities or injuries may go 
unnoticed or unreported. Vessel strikes affect adult females returning 
to nest, removing individuals and their future reproductive potential. 
Thus, this threat reduces the abundance and productivity of the DPS. We 
conclude that vessel strikes pose a threat to the DPS.

Pollution

    Pollution includes contaminants, marine debris, and ghost fishing 
gear. As with all leatherback turtles, entanglement in and ingestion of 
marine debris and plastics are threats that likely kill several 
individuals a year. For six stranded hatchlings and 24 stranded adults 
over the past 40 years, the cause of death was generally unknown. 
However, fishery-related injuries, ghost-fishing (i.e., entanglement in 
discarded fishing gear), disease, or pollution may be responsible (de 
Wet 2012). Plastic pollution may be a main threat in the waters off 
Mozambique (M. Pereira, pers. comm., 2019). Outer accumulation of the 
Indian Ocean ``garbage patch'' (Cozar et al. 2014) overlaps with 
foraging areas in the Mozambique Channel and occurs in waters offshore 
from nesting areas in South Africa and Mozambique. Though we were 
unable to find ingestion or entanglement data for SW Indian leatherback 
turtles, 51.4 percent of gut and fecal samples from loggerhead turtles 
(n = 74) captured as bycatch in the Reunion Island longline fishery 
contained marine debris, of which plastic comprised 96.2 percent 
(Hoarau et al. 2014). Ryan et al. (2016) found that 24 of 40 loggerhead 
turtle post-hatchlings had ingested plastics or other anthropogenic 
debris. Based on the foraging behavior of leatherback turtles and the 
proximity of the ``garbage patch,'' we conclude that the ingestion and 
entanglement of marine debris are threats to this DPS.
    In addition, State of the World's Sea Turtles (SWOT 2017) 
identifies hydrocarbon extraction along the eastern African seaboard, 
including northern Mozambique, as the greatest emerging concern for 
this DPS. They report that the impact of such activities remain to be 
seen (SWOT 2017). However associated oil spills are likely to modify 
habitat off nesting beaches and reduce prey availability for all life 
stages. Harris et al. (2018) found that the hydrocarbon industry poses 
a moderate threat to the DPS because of its spatial overlap with 
migratory corridors (second in extent, after longline fisheries). They 
expressed concern over the expansion of the hydrocarbon extraction 
along the coasts of southern Mozambique and northeastern South African 
and the possibility of an oil spill in these areas (Harris et al. 
2018). Pretorius (2018) identified 28 significant impacts to sea 
turtles as a result of hydrocarbon exploration and production; these 
included: Potential water pollution, light pollution, noise pollution, 
and habitat destruction. However, Du Preez et al. (2018) reports that 
metal and metalloid contaminants do not appear to be a problem for this 
DPS. We conclude that pollution poses a threat to the DPS.

Climate Change

    Climate change is a threat to the SW Indian DPS. The impacts of 
climate change include: Increases in temperatures (air, sand, and sea 
surface); sea level rise; increased coastal erosion; more frequent and 
intense storm events; and changes in ocean currents.
    Angel et al. (2014) identifies coastal erosion as the main beach-
based threat to this population and one that is likely to increase with 
climate change. Though coastal erosion is a natural process, sea level 
rise (as a result of climate change) increases the rate of erosion and 
the amount of beach affected. In Bazaruto Archipelago, Mozambique, 
coastal erosion and rising sea levels destroyed approximately 12 
percent of nests over 10 seasons of monitoring (Videira and Louro 2005; 
Louro 2006). Because leatherback turtles nest lower on the beach than 
other sea turtles, their eggs are more at risk of being exposed and 
destroyed by increases in sea level and coastal erosion (Boyes et al. 
2010). Thus, erosion and rising sea level as a result of climate change 
are a threat to the DPS.
    Sand temperatures influence leatherbacks' egg viability and sex 
determination. Temperatures over 32 [deg]C

[[Page 48381]]

result in death and temperatures below 29.2 [deg]C produce only males 
(Rimblot et al. 1985; Rimblot-Baly et al. 1986). Temperature probes on 
South African beaches reveal that nests are already close to pivotal 
temperatures, with an average of 29.04 [deg]C (S.D. 0.86 [deg]C; range 
27.62 to 29.69 [deg]C; Boyes et al. 2010). A modeling study suggests 
that even if South African beaches experience a temperature increase of 
5 [deg]C, hatching success and emergence success may not be 
significantly reduced (Santidri[aacute]n Tomillo et al. 2015). Instead, 
nesting females may shift their nesting season to months (e.g., July 
through October) when temperature and precipitation would be similar to 
current conditions of the current nesting season (i.e., October through 
January). However, the authors cautioned that because nesting females 
do not change their nesting habits in response to oceanographic 
conditions, they may not change their nesting habits in response to 
climate change either (Santidri[aacute]n Tomillo et al. 2015). In 
addition, a shift in the nesting season could have impacts beyond 
hatching success, such as reduced post-hatchling survival and 
suboptimal foraging conditions for post-nesting females. We therefore 
conclude that increased temperatures may be a threat to the DPS, and 
will likely result in impacts ranging from nesting season shifts to 
significant nest losses.
    The threat of climate change may modify the nesting conditions for 
the entire DPS. Impacts likely range from small, temporal changes in 
nesting season to large losses of productivity. Because we are already 
seeing small impacts due to coastal erosion and sea level rise, we 
conclude that climate change is a threat to the SW Indian DPS.

Conservation Efforts

    There are numerous efforts to conserve the leatherback turtle. The 
following conservation efforts apply to the SW Indian DPS (for a 
description of each effort, please see the section on conservation 
efforts for the overall taxonomic species): African Convention on the 
Conservation of Nature and Natural Resources (Algiers Convention), 
Convention on the Conservation of Migratory Species of Wild Animals, 
Convention on Biological Diversity, Convention on International Trade 
in Endangered Species of Wild Fauna and Flora, Convention on the 
Conservation of European Wildlife and Natural Habitats, Convention for 
the Co-operation in the Protection and Development of the Marine and 
Coastal Environment of the West and Central African Region (Abidjan 
Convention) and Memorandum of Understanding Concerning Conservation 
Measures for Marine Turtles of the Atlantic Coast of Africa (Abidjan 
Memorandum), Convention Concerning the Protection of the World Cultural 
and Natural Heritage (World Heritage Convention), FAO Technical 
Consultation on Sea Turtle-Fishery Interactions, Indian Ocean Tuna 
Commission, The Indian Ocean Tuna Commission, Indian Ocean--South-East 
Asian Marine Turtle Memorandum of Understanding, MARPOL, IUCN, Nairobi 
Convention for the Protection, Management and Development of the Marine 
and Coastal Environment of the Eastern African Region, Ramsar 
Convention on Wetlands, UNCLOS, and UN Resolution 44/225 on Large-Scale 
Pelagic Driftnet Fishing. Although numerous conservation efforts apply 
to the turtles of this DPS, they do not adequately reduce its risk of 
extinction.

Extinction Risk Analysis

    After reviewing the best available information, the Team concluded 
that the SW Indian DPS is at high risk of extinction. The DPS exhibits 
a total index of nesting female abundance of 149 females. Such a 
limited nesting population size places this DPS in danger of stochastic 
or catastrophic events that increase its extinction risk. This DPS 
exhibits a slightly decreasing nest trend at monitored nesting beaches 
in South Africa. This declining trend has the potential to further 
lower abundance and thereby increase the risk of extinction. With only 
one nesting aggregation, the DPS lacks spatial structure, and its 
genetic diversity is low. Thus, stochastic events could have 
catastrophic effects on nesting for the entire DPS, with no potential 
source subpopulations to buffer losses or provide additional diversity. 
However, the DPS uses multiple, distant, and diverse foraging areas, 
providing some resilience against reduced prey availability. Based on 
these factors, we find the DPS to be at risk of extinction, likely as a 
result of past threats.
    Current (ongoing) threats further contribute the risk of extinction 
of this DPS. The primary threat to this DPS is bycatch in commercial 
and artisanal, pelagic and coastal, fisheries. Longline fisheries 
constitute the greatest threat. Though poorly studied, other fisheries 
together may have overall mortality rates for affected turtles from 
this DPS that rival those from longline fisheries. Fisheries bycatch 
reduces abundance by removing individuals from the population. Because 
several fisheries operate near nesting beaches, productivity is also 
reduced when nesting females are prevented from returning to nesting 
beaches. Exposure and impact of this threat are high. Poaching is also 
a threat to the DPS. Egg and turtle poaching, while no longer a threat 
in South Africa, likely continues in Mozambique. In Madagascar, turtles 
are illegally captured at sea and consumed in local villages. Vessel 
strikes also pose a threat. Vessel strikes kill several leatherback 
turtles each year, including females returning to beaches to nest. 
While exposure is low, impacts are high, affecting both abundance and 
productivity. Coastal erosion and beach driving in Mozambique modify 
nesting habitat and are believed to result in minor reductions in 
productivity currently. However, these threats are likely to increase 
over time as climate change and tourism increases. Climate change is 
likely to result in reduced productivity due to greater rates of 
coastal erosion and nest inundation. Predation of eggs and hatchlings 
is also a threat. However, although predation has the potential to 
reduce productivity, the DPS has likely adapted to predation by native 
species, which account for most of the predation at present. Ingestion 
of plastics and entanglement in marine debris are threats to all 
leatherback turtles, most likely resulting in injury and reduced 
health, though sometimes mortality. Though many regulatory mechanisms 
are in place, they do not reduce the impact of these threats to levels 
that allow the DPS to continue its previous increasing nesting trend.
    Thus, the Team unanimously concluded, that the SW Indian DPS is at 
high risk of extinction. The total index of nesting female abundance of 
149 females makes the DPS highly vulnerable to threats. We determine, 
consistent with the team's findings, that the DPS is currently ``in 
danger of extinction.'' The slightly declining nest trend and lack of 
spatial structure and diversity further contribute to its risk of 
extinction. While this small population had an increasing or stable 
nesting trend for decades, the lack of continued population growth and 
recent decline may indicate that threats have outpaced productivity. 
Past egg and turtle harvest initially reduced the nesting female 
abundance of this DPS and likely confined its nesting habitat to a 
relatively small geographic area, with little diversity or spatial 
structure. Currently, fisheries bycatch is the primary present, ongoing 
threat. It reduces abundance and productivity (i.e., imminent and 
substantial demographic risks) by removing mature and immature 
individuals from the

[[Page 48382]]

population at rates exceeding replacement. Though numerous conservation 
efforts apply to this DPS, they do not adequately reduce the risk of 
extinction. We conclude that the SW Indian DPS is in danger of 
extinction throughout its range and therefore meets the definition of 
an endangered species. The threatened species definition does not apply 
because the DPS is at risk of extinction currently (i.e., at present), 
rather than on a trajectory to become so within the foreseeable future.

NE Indian DPS

    The Team defined the NE Indian DPS as leatherback turtles 
originating from the NE Indian Ocean, south of 71[deg] N, east of 
61.577[deg] E, and west of 120[deg] E. The western boundary occurs at 
the border between Iran and Pakistan, where the Somali Current begins. 
This current, and the cold waters of the Antarctic Circumpolar Current, 
likely restrict the nesting range of this DPS. We placed the eastern 
boundary at 120[deg] E to approximate the Wallace and Huxley lines, 
which are established biogeographic barriers to gene flow between 
Indian and Pacific Ocean populations of numerous species. While the 
genetic differences between the NE Indian and West Pacific DPSs 
demonstrate discreteness, genetic sampling is unavailable from areas 
where the nesting range of the DPSs likely meet, preventing us from 
defining the boundary more specifically.
    The range of the DPS (i.e., all areas of documented occurrence) 
extends throughout the Indian Ocean and possibly into the Pacific 
Ocean. Records indicate that the species occurs in the waters of the 
following nations: India, Sri Lanka, Bangladesh, Myanmar, Thailand, 
Malaysia, Indonesia, Vietnam, China, and Philippines (Hamann et al. 
2006). Given the range of the DPS, leatherbacks may also occur in the 
waters of Pakistan, Australia, Brunei, Cambodia, Philippines, and 
Taiwan.
    Leatherback turtles of the NE Indian DPS nest on beaches scattered 
throughout the NE Indian Ocean. The largest abundance of nesting occurs 
on beaches of the Andaman and Nicobar Islands in India. The sandy 
beaches of the Andaman and Nicobar Islands consist of soft limestone 
formed of coral and shell (Lal 1976; Bandopadhyay and Carter 2017). A 
moderate amount of nesting occurs in Sri Lanka, and even less occurs in 
Thailand and Sumatra, Indonesia (Hamann et al. 2006; Nel 2015).
    Information on this DPS is limited, but foraging appears to occur 
throughout the Indian Ocean (Andrews et al. 2006; Hamann et al. 2006). 
The foraging range extends throughout the Bay of Bengal, south of Sri 
Lanka, and along the west coast of Sumatra, Indonesia, as indicated by 
satellite telemetry data and fisheries reports (NMFS and FWS 2013). 
Nesting females at Little Andaman Island likely use a variety of 
foraging areas and have been tracked to: South and east of the Andaman 
and Nicobar Islands; along the coast of Sumatra; beyond Cocos (Keeling) 
Island towards Western Australia; and across the Indian Ocean towards 
Madagascar and the African continent (Namboothri et al. 2012; 
Swaminathan et al. 2017; Swaminathan et al. 2019). Stranding data also 
indicate the use of diverse foraging areas: 15 individuals stranded or 
were caught in fishing gear along the mainland coast of India (Shanker 
2013). Leatherback turtles have also stranded along the coasts of 
Mindanao, Philippines and Pakistan (Firdous 2006; Lucero et al. 2011).

Abundance

    The total index of nesting female abundance of the NE Indian DPS is 
109 females. We based this total index on the nesting aggregations at 
South and West Bays, Little Andaman Island, India (K. Shanker pers. 
comm., 2018). Our total index does not include 14 unquantified nesting 
aggregations in Bangladesh, India, Indonesia, Malaysia, Myanmar, Sri 
Lanka, Thailand, Philippines, and Vietnam. To calculate the index of 
nesting female abundance, we divided the total number of nests at South 
and West Bays, Little Andaman Island between the 2015/2016 and 2017/
2018 nesting seasons (i.e., a 3-year remigration interval; Andrews 
2002) by the clutch frequency (3.8 clutches/season; Andrews 2002; 
Eckert et al. 2015). This number represents an index of abundance for 
this DPS, and is likely to be an underestimate, because it only 
includes available data from recently and consistently monitored 
nesting beaches. Additional nesting occurs at other locations but is 
unquantified.
    Published estimates of total nesting female abundance are not 
available for this DPS. The IUCN Red List assessment did not provide an 
estimate of the total number of mature individuals because monitoring 
was not sufficient (Tiwari et al. 2013). Currently, the largest nesting 
aggregations occur in the Andaman and Nicobar Islands of India. Nesting 
in Sri Lanka may consist of about 100 to 200 nesting females per year, 
and low levels of nesting occur in Thailand and Sumatra, Indonesia 
(Hamann et al. 2006; Nel 2012). Low and scattered nesting occurs in 
Indonesia: 1 to 14 nesting females annually at Alas Purwo in East Java; 
and one to three nesting females annually on three beaches in Bali. 
There are also rare reports of nesting in the Philippines (Lucero et 
al. 2011; Arguelles 2013), Vietnam, and Malaysia. In Myanmar, nesting 
is rare, and only one confirmed nesting event has been recorded in 
recent years (i.e., December 2016; Platt et al. 2017). Historically, 
there may have been nesting in Bangladesh, but no current reports exist 
(Hamann et al. 2006).
    Malaysia once hosted the DPS's largest nesting aggregation (Chan 
and Liew 1996). It is now considered functionally extinct or extirpated 
(Pilcher et al. 2013), as a result of continuous, large-scale egg 
harvest and fisheries bycatch (Chan and Liew 1996; Eckert et al. 2012). 
The major nesting site in Malaysia, Rantau Bang in Terengganu, 
decreased drastically from 10,000 nests in the 1950s to 10 or fewer 
nests in the 2010s (reviewed by Eckert et al. 2012), and to one or no 
nests annually, more recently. The number of nesting females in Vietnam 
has also decreased dramatically, from approximately 500 nesting females 
in the 1960s to two to three nests in 2005 and 2007 (The Chu and Nguyen 
2015). In the late 1970s, females nested in multiple locations of 
Thailand, including: along the airport beach in Changwat Phuket; in the 
Laem Phan Wa marine reserve; and in coastal Changwan Phangnga (Bain and 
Humphry 1980). Settle (1995) recorded about 30 nests on the Phuket and 
Phangnga coastlines from 1992 to 1993. Aureggi et al. (1999) found nine 
nests between 1997 and 1998, during a survey of Phra Thong Island in 
the south.
    Our total index of nesting female abundance (109 females) places 
the DPS at risk for environmental variation, genetic complications, 
demographic stochasticity, negative ecological feedback, and 
catastrophes (McElhany et al. 2000; NMFS 2017). These processes, 
working alone or in concert, place small populations at a greater 
extinction risk than large populations, which are better able to absorb 
losses in individuals. Due to its small size, the DPS has restricted 
capacity to buffer such losses. Historic abundance estimates indicate 
that the DPS was once much larger. The current abundance is likely a 
result of past and current threats, which we describe below. Given the 
intrinsic problems of small population size, we conclude that the 
limited nesting female abundance is a major factor in the extinction 
risk of this DPS.

[[Page 48383]]

Productivity

    The NE Indian DPS has exhibited a drastic population decline with 
extirpation of its largest nesting aggregation in Malaysia. Nest counts 
from Malaysia exhibited a steep decline of 17.9 percent annually (sd = 
4.2 percent; 95 percent CI = -25.5 to -8.4 percent; f = 0.998; mean 
annual nests = 1,166) over the 44-year period of data collection (1967 
to 2010). The drastic decline of nests observed in Malaysia is likely 
representative of the overall trend for the DPS given the magnitude of 
historical abundance for that site and the high confidence in the trend 
estimate.
    Despite the dramatic population decline, driven by the extirpation 
of the largest nesting aggregation (i.e., Malaysia), productivity 
parameters are similar to the species averages. However, we have a low 
degree of confidence in these estimates due to limited monitoring of 
existing nesting aggregations. We conclude that the NE Indian DPS 
exhibits a declining nesting trend, which increases its extinction 
risk.

Spatial Distribution

    For the NE Indian DPS, nesting is limited to a few, scattered 
nesting beaches. Currently, the majority of the nesting occurs on 
beaches of the Andaman and Nicobar Islands and Sri Lanka, with small 
numbers of nests on the western coast of Thailand, Sumatra, and Java 
(Nel et al. 2015).
    Spatial structure is unknown but presumed to be low. There are no 
estimates of genetic population structure within this DPS because 
published genotypes only exist for Malaysia (Dutton et al. 1999, 2007). 
Genetic samples were taken from nesting females at Little Andaman 
Island from 2008 through 2010, but the results are not yet available 
(Namboothri et al. 2010).
    The wide distribution of foraging areas likely buffers the DPS 
somewhat against local catastrophes or environmental changes that would 
limit prey availability. Remaining nesting is limited to a few, 
scattered but broadly distributed nesting sites. The largest nesting 
aggregations are clustered, thus rendering the DPS susceptible to 
environmental catastrophes (e.g., tsunamis), and directional changes 
(e.g., sea level rise). Thus, despite widely distributed foraging 
areas, the somewhat limited nesting distribution increases the 
extinction risk of the NE Indian DPS.

Diversity

    Genetic diversity of the NE Indian DPS is potentially relatively 
high, based on analyses of samples collected from the previously large, 
but now functionally extinct, nesting aggregation in Malaysia (Dutton 
et al. 1999, 2007); genetic diversity has not been assessed at other 
nesting sites. The diversity of nesting sites is low, given that the 
majority of the nesting currently occurs on islands (Sivasundar and 
Prasad 1996). We conclude that existing diversity provides little 
resilience to the DPS.

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    Erosion, coastal development, and artificial lighting have 
destroyed or modified the available, suitable nesting habitat and thus 
are threats to the NE Indian DPS.
    Erosion reduces the available nesting habitat for the DPS. Some 
erosion occurs as a result of natural disasters. In 2004, a major 
earthquake occurred off the west coast of Sumatra, Indonesia, resulting 
in the 2004 tsunami, which destroyed many of the beaches that once 
hosted over 1,000 nests (Subramaniam et al. 2009). As a result of the 
tsunami, the width of the coastline was reduced by one meter, severely 
modifying the beaches of South Bay, Little Andaman Island, and 
resulting in very low leatherback nesting in 2005 and 2006 (Namboothri 
2010). The tsunami also caused drastic changes at other leatherback 
nesting beaches (Alfred et al. 2005; Ramachandran et al. 2005; Murugan 
2005; Andrews et al. 2006). It caused erosion at some beaches and 
accretion at others, especially in the Andaman and Nicobar Islands, 
which lie closest to the epi-center of the earthquake and host the 
largest numbers of nesting females in the DPS (Subramaniam et al. 
2009). In addition, the beaches in Indonesia are being lost due to 
erosion from high tides and monsoons (Obermeier 2002).
    Sand mining and tourism-related development are the main threats to 
nesting habitat (Fatima et al. 2011). While we were unable to find 
specific information regarding sand mining, coastal development is 
increasing in Sri Lanka, India, and Bangladesh. The beaches of Sri 
Lanka are under high threat from tourism development (e.g., large 
hotels and restaurants), urban and industrial development, and 
artificial lighting (Kapurisinghe 2006). Along the mainland of India, 
granite blocks and embankments prevent turtles from approaching many 
beaches (Andrews et al. 2006). Intense coastal development, stemming 
from the tourism industry, occurs in Bangladesh without environmental 
review (Pilcher 2006), resulting in the alteration of sand dunes and 
other activities that reduce the quality of nesting habitat (Islam 
2002; Islam et al. 2011). In Vietnam, increasing tourism is expected to 
result in coastal development on the beaches of Son Tra Peninsula, 
QuanLan, and Minh Chau (Ministry of Fisheries 2003). In addition, most 
Vietnam beaches are affected by a large amount of marine debris (e.g., 
glass, plastics, polystyrenes, floats, nets, and light bulbs), which 
can entrap turtles and impede nesting activity.
    Artificial lighting modifies the quality of nesting beaches because 
lights over land disorient nesting females and hatchlings. Instead of 
crawling toward the surf and their marine habitat, they crawl further 
inland, where they may become dehydrated and die or are susceptible to 
predation. Nests moved to hatcheries as part of conservation efforts 
may be subject to inadequate hatchery practices, which have resulted in 
skewed sex ratios and low hatching success (Chan and Liew 1996; 
Kapurisinghe 2006; Rajakaruna et al. 2013; Phillott et al. 2018).
    Some areas are protected. Of the 306 islands in the Andaman and 
Nicobar Islands of India, 94 are designated as wildlife sanctuaries, 
six of which are national parks, and two of which are marine national 
parks (Andrews et al. 2006). In Sri Lanka, in 2006, sea turtle 
sanctuaries were established at Rekawa (4.5 km stretch) and Godawaya 
(3.8 km stretch), where high frequency leatherback nesting is observed; 
the area is bounded 500 meters towards the sea and 100 meters towards 
the land from the high tide level in both sites (Phillott et al. 2018). 
Although laws protect sea turtles throughout Sri Lanka, most nesting 
areas are not protected and hence, local communities can disturb 
nesting beaches by removing sand, lighting the beaches, and cutting the 
beach vegetation (Phillott et al. 2018). In Malaysia, turtle 
sanctuaries have been established in Terengganu, Sabah, and Sarawak. 
However, nesting habitat modification and destruction continue in many 
areas.
    We conclude that nesting females, hatchlings, and eggs are exposed 
to the reduction and modification of nesting habitat, as a result of 
erosion, coastal development, and artificial lighting. These threats 
impact the DPS by reducing nesting and hatching success, thus lowering 
its productivity. The most abundant remaining nesting aggregations are 
protected from

[[Page 48384]]

development, but they experience high rates of erosion; other nesting 
beaches are subject to anthropogenic threats. Thus, we conclude that 
habitat loss and modification pose a threat to the NE Indian DPS.

Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    Overutilization is a threat to the NE Indian DPS. The harvest of 
turtles and eggs led to the historical decline of the DPS, and poaching 
continues in several areas (Phillott et al. 2018).
    Regular, nearly complete egg harvest caused the functional 
extinction of the once large nesting aggregation in Malaysia (Chan and 
Liew 1996). In the early 1960s, the Terengganu, Malaysia nesting 
beaches were leased to the highest bidder, and nearly all leatherback 
eggs were harvested. In the 1980s, the State Fisheries Department tried 
to buy back about 10 percent of the harvested eggs to be incubated in a 
hatchery (Siow and Moll 1982; Chan and Liew 1996; Stiles 2009). 
However, such efforts could not prevent the extirpation. Excessive egg 
harvest, both legal and illegal, also caused declines in India, Sri 
Lanka, and Thailand (Ross 1982).
    The harvest of turtles and eggs continues but has not been 
quantified (Nel 2012). In Sri Lanka, almost all eggs are taken from the 
beach and sold at markets or to hatcheries for ecotourism (Kapurusinghe 
2000, 2006; Rajakaruna et al. 2013; Phillott et al. 2018). The 
conservation benefit provided by hatcheries in Sri Lanka is debatable 
(Phillott et al. 2018) because they do not follow the hatchery 
practices established by the IUCN (Hewavisenthi 1994; IUCN 2005; 
Namboothri et al. 2012; Rajakaruna et al. 2013; Phillott et al. 2018). 
Egg harvest also continues in Thailand. Commercial egg harvest is 
illegal in the Andaman and Nicobar Islands, and in the Andaman Islands, 
a ban on hunting and harvesting of turtles came into force in 1977. 
However, the original inhabitants of the Andaman and Nicobar Islands 
are exempt from the Indian Wildlife Protection Act (Shanker and Andrews 
2004), and Namboothri et al. (2012) observed intense egg harvest at 
Delgarno, Trilby, and East Turtle Islands. In Myanmar, despite 
regulations prohibiting the consumption of turtle meat and eggs (Hamann 
et al. 2006), there is illegal trade of turtles caught at sea, 
including leatherback turtles (Murugan 2007). In Sri Lanka, the 
historically high direct take of turtles at sea has been reduced 
(Kapurushinghe 2006). Records indicate that turtle meat and parts were 
once regularly exported from Tamil Nadu, India, to Sri Lanka, and then 
to other nations such as the United States, Singapore, and Belgium 
(Kuriyan 1950; Chari 1964; Shanmughasundarun 1968 as cited in 
Agastheesapillai and Thiagarajan 1979).
    Exposure to egg and turtle poaching remains high throughout the 
range of the DPS. Poaching of nesting females or post-nesting females 
at sea reduces both abundance (through loss of nesting females) and 
productivity (through loss of reproductive potential). Such impacts are 
high because they directly remove the most productive individuals from 
the DPS, reducing current and future reproductive potential. Egg 
harvest reduces productivity only, but, as previously demonstrated 
within this DPS, can have devastating population-level impacts. We 
conclude that overutilization, as a result of egg and turtle harvest, 
poses a major threat to the NE Indian DPS.

Disease or Predation

    While we could not find any information on disease for this DPS, 
the best available data indicate that predation is a threat to the NE 
Indian DPS. Multiple predators prey on eggs and hatchlings at several 
nesting beaches (Andrews 2000). During a 2016 survey of the Nicobar 
Islands, approximately 57 percent (n = 1,223) of leatherback nests were 
lost to depredation by feral dogs, water monitor lizards, or feral pigs 
(Sus domesticus; Swaminathan et al. 2017). In the South Bay of Great 
Nicobar Island, wild boars and dogs prey on eggs, while fiddler crabs, 
dogs, and raptors prey on hatchlings (Sivakumar 2002). Sivasundar and 
Prasad (1996) documented that Asian water monitor lizards took 68.6 
percent of leatherback nests in the Andaman Islands. In Sri Lanka, egg 
predators include feral dogs, water and land monitor lizards, jackals, 
wild boars, mongooses, and ants. Egg predation by feral pigs is a major 
threat in Indonesia (Maturbongs et al. 1993; Maturbongs 1995, 1996; 
Sivasundar and Prasad 1996).
    A large number of eggs and hatchlings are exposed to predation. 
Though leatherback turtles produce a large number of eggs and 
hatchlings, published rates of predation (57 to 69 percent) are high. 
The predation of eggs and hatchlings mainly impacts productivity. We 
conclude that predation poses a threat to the NE Indian DPS.

Inadequacy of Existing Regulatory Mechanisms

    Turtles of the NE Indian DPS are protected by several regulatory 
mechanisms. For each, we review the objectives of the regulation and to 
what extent it adequately addresses the targeted threat. Nearly all 
nations that host nesting aggregations have legislation to protect sea 
turtles.
    In India, the leatherback turtle is included on Schedule I, Part II 
of the Wildlife (Protection) Act, 1972 (Entry No. 11) updated by Wild 
Life (Protection) Amendment Act, 2002 (No. 16 of 2003). India also bans 
the hunting and trade of wild animals (India National Report to CMS, 
1991 and 1994). However, the indigenous people of the Andaman and 
Nicobar Islands are exempt from these laws. India has regulations to 
require TEDs and minimize fisheries interactions; and much of the 
Andaman and Nicobar Islands are protected as wildlife sanctuaries, 
including two marine national parks (Andrews et al. 2006).
    In Indonesia, Order No. 301/1991 lists leatherback turtles as a 
protected species. Pursuant to the Act of 10 August 1990 on the 
Conservation of Living Resources and Their Ecosystems, it is prohibited 
to kill, capture, possess, transport, trade in or export protected 
animals whether alive or dead, or parts of such animals. The taking, 
destruction, trade or possession of the eggs or nests of protected 
animals are also prohibited (ECOLEX 2003). There are no habitat 
protections and no regulations to minimize fisheries interactions or 
require TEDs in Indonesia.
    In Sabah, Malaysia, the leatherback turtle is not listed as a 
totally protected or partially protected species in the Wildlife 
Conservation Enactment (No. 6 of 1997). In Sarawak, Malaysia, 
leatherback turtles have been fully protected since 1958. Under the 
Wildlife Protection Ordinance 1998, all marine turtles in Malaysia are 
protected from hunting, killing, capture, sale, import, export, 
possession of any animal, recognizable part or derivative or any nest, 
except in accordance with the permission in writing of the Controller 
of Wildlife for scientific or educational purposes or for the 
protection or conservation of a species (Tisen and Bali 2002). The 
nesting beach at Rantau Abang, Terengganu is protected. However, the 
nesting aggregation that once used this beach has been extirpated. In 
1994, the waters surrounding 38 offshore islands of Peninsular Malaysia 
and Labuan became protected as marine parks. In

[[Page 48385]]

addition, one national park in Sarawak, three in Sabah, and one state 
park in Terengganu protect coastal and marine ecosystems (Malaysia 
National Biodiversity Policy 1998). Additional habitat protections 
include: The Turtle Trust Ordinance 1957; Land Code 1958; Turtle 
Protection Rules 1962; Fisheries Prohibited Areas under section 61 of 
the Fisheries Act 1985; and the Wildlife Protection Ordinance 1998 
(Tisen and Bali 2002). The use of TEDs will be required in Malaysia by 
2020.
    In Myanmar, the Burma Wildlife Protection Act 1936 (Act No. VII of 
1936) requires licenses to hunt, possess, sell, or buy wild animals 
with closed hunting seasons (FAOLEX 2003). The Burma Wildlife 
Protection Rules of 1941 states that the import or export of any 
reptile (including parts or products) into or from Myanmar is 
prohibited.
    In Pakistan, the leatherback turtle is protected in Baluchistan, 
Azad Kashmir and Sind (Baluchistan Wildlife Protection Act 1974 No.19/
1974; The Azad Jammu and Kashmir Wildlife Act 1975 No.23/1975; The 
Sindh Wildlife Protection Ordinance 1972 No.5/1972). Possession, 
transport, and/or national trade are prohibited or regulated (ECOLEX 
2003).
    In Sri Lanka, the leatherback turtle is protected under the Fauna 
and Flora Protection Ordinance (Sri Lanka National Report to CMS 1994), 
which makes it an offense to kill, wound, harm or take a turtle, or to 
use a noose, net, trap, explosive or any other device for those 
purposes, to keep in possession a turtle (dead or alive) or any part of 
a turtle, to sell or expose for sale a turtle or part of a turtle, or 
to destroy or take turtle eggs. The minister of Fisheries and Aquatic 
Resources may also prohibit or regulate the import and export of 
turtles or their derivatives (Parliament of the Democratic Socialist 
Republic of Sri Lanka 1993). The nesting beach in Yala Reserve is also 
protected.
    In Thailand, the Leatherback Turtle is protected under the Animals 
Protection Act B.D 2535 (The Zoological Park Organization 2003).
    In summary, numerous regulatory mechanisms protect leatherback 
turtles, their eggs, and nesting habitat throughout the range of this 
DPS. Although these regulatory mechanisms provide some protection, many 
do not adequately reduce the threat that they were designed to address, 
generally as a result of limited implementation or enforcement. As a 
result, bycatch, nesting habitat protection, and legal and illegal 
harvest remain threats to the DPS. We conclude that the inadequacy of 
the regulatory mechanisms is a threat to the NE Indian DPS.

Fisheries Bycatch

    Fisheries bycatch is a threat to the NE Indian DPS. Capture in 
gillnet, trawl, purse seine, and longline fisheries is a significant 
cause of leatherback mortality for this DPS (Wright and Mohanty 2002; 
Hamann et al. 2006; Project GloBAL 2007; Bourjea et al. 2008; 
Abdulqader 2010; Wallace et al. 2010).
    Gillnet fisheries pose a major threat to the DPS. A survey 
conducted at 16 main fishing ports in Sri Lanka estimated that 431 
leatherback turtles were caught in gillnets between 1999 and 2000 
(Kapurusinghe and Cooray 2002). In Malaysia, Chan et al. (1988) 
reported an average of 742 and 422 sea turtles, most of which were 
leatherback turtles, caught in drift gillnets and bottom longlines, 
respectively. In Bangladesh, gillnets, set bag nets, trawl nets, seine 
nets, hook and line and other net types of gear capture turtles 
(Hossain and Hoq 2010). Gillnet and purse seine fisheries are common 
off the coasts of the Andaman and Nicobar Islands, where the largest 
nesting aggregations occur (Shanker and Pilcher 2003; Chandi et al. 
2012).
    Trawl fisheries also pose a large threat to the DPS. In India, TEDs 
are required for trawl nets. However, fishers are reluctant to use them 
(Murugan 2007). Trawl fishing is also common in Bangladesh, and the use 
of TEDs is not required (Ahmed et al. 2006).
    Longline fisheries occur in coastal and pelagic waters. Huang and 
Liu (2010) evaluated observer data from 77 trips (4,409 sets) on 
Taiwanese large-scale longline fishing vessels in the Indian Ocean. 
They identified 84 leatherback turtles captured from 2004 to 2008, with 
48 mortalities (57 percent; Huang and Liu 2010). Extrapolating to the 
entire Taiwanese longline fishery in the Indian Ocean, they estimated 
an average bycatch of 173 leatherback turtles between 2004 and 2007. 
This number likely includes individuals from both the SW and NE Indian 
DPSs (Louro 2006). In Vietnam, longline fisheries continue to capture 
leatherback turtles. However, a circle hook program has been 
implemented to minimize the impact (WWF 2013).
    Purse seine fisheries have a much lower impact than longline 
fisheries (Angel et al. 2014); two leatherback turtles were captured 
(alive) between 1995 and 2010 in the Indian Ocean (Clermont et al. 
2012). In the EEZ of all Indian Ocean French Territories (mostly from 
the Mozambique Channel), 40 leatherback turtles were captured in 
unspecified fisheries from 1996 to 1999; 92 percent were released alive 
(Ciccione 2006).
    In Thailand, one of the main causes of decline in the turtle 
population is bycatch in trawl, drift gillnet, and purse seine 
fisheries. The rapid expansion of fishing operations is largely 
responsible for the increase in adult turtle mortality due to bycatch 
(Settle 1995).
    In Malaysia, the Fisheries Act of 1985 prohibited capture of sea 
turtles by any type of fishery. However, this merely reduced the 
reporting of interactions (Yeo et al. 2011 in Dutton et al. 2011). The 
1991 Regulations prohibit fishing in waters adjacent to Rantau Abang 
during the leatherback nesting season (Chan 1993).
    We conclude that juveniles and adults are exposed to high fishing 
effort throughout their foraging range and in coastal waters near 
nesting beaches. Mortality rates are likely high, especially in areas 
where turtle meat is consumed. Mortality reduces abundance, by removing 
individuals from the population. It also reduces productivity, when 
nesting females are incidentally captured and killed. We conclude that 
fisheries bycatch is a major threat to the NE Indian DPS.

Pollution

    Pollution includes contaminants, marine debris, and ghost fishing 
gear. Ghost fishing gear can drift in the ocean and fish unattended for 
decades and kill numerous individuals (Wilcox et al. 2013). The main 
sources of ghost fishing gear are gillnet, purse seine, and trawl 
fisheries (Stelfox et al. 2016). In one collection event, volunteers 
collected over 600 nets, ropes, and buoys from India, Maldives, Oman, 
Pakistan, Sri Lanka, and Thailand (Stelfox et al. 2016). Though 
educational programs created in 2014 focus on reusing and recycling 
fishing gear, the threat continues throughout the range of the DPS. 
Ghost nets in the Maldives primarily drift from fisheries in the Bay of 
Bengal (e.g., Sri Lanka and India; Stelfox et al. 2016). Around the 
Andaman and Nicobar Islands and Sri Lanka, plastics and other garbage 
are washed from polluted beaches and inland waters to the sea, where 
they can kill or harm sea turtles through ingestion or entanglement 
(Kapurusinghe 2006; Das et al. 2016). Pollution has been identified as 
a main threat to sea turtles in Iran (Mobaraki 2007) and Pakistan 
(Firdous 2001). However, no specific information about the type of 
pollution was provided. In Gujarat, India, increased port and shipping 
traffic have resulted in oil spills and the release of other

[[Page 48386]]

pollutants, such as fertilizers and cement (Sunderraj et al. 2006). 
Heavy metals and E. coli were found at relatively high levels in the 
waters of Malaysia (including Terengganu) and in the pancreases and 
livers of leatherback turtles (Caurant et al. 1999; Ngah et al. 2012). 
It is not known how these pollutants affect leatherback physiology 
(Jakimska et al. 2011).
    As with all leatherback turtles, entanglement in and ingestion of 
marine debris and plastics are threats that likely kill several 
individuals a year. However, data specific to this DPS were not 
available. We conclude that pollution is a threat to the NE Indian DPS, 
albeit with effects that are unquantifiable on the basis of the best 
available information.

Climate Change

    Climate change is a threat to the NE Indian DPS. A significant rise 
in sea level would further reduce nesting habitat, which is already 
affected by erosion. The DPS is also likely to be affected by increases 
in sand temperatures (Hawkes et al. 2009; Poloczanska et al. 2009). 
Sand temperatures prevailing during the middle third of the incubation 
period determine the sex of hatchling sea turtles (Mrosovsky and Yntema 
1980). Incubation temperatures near the upper end of the tolerable 
range produce only female hatchlings, while incubation temperatures 
near the lower end of the tolerable range produce only males. As 
temperatures increase, incubation temperatures may exceed the thermal 
tolerance for embryonic development, thus increasing embryo and 
hatchling mortality.
    In addition, the frequency and intensity of severe storm events and 
cyclones in the Bay of Bengal are predicted to increase with climate 
change (Balaguru et al. 2014).
    Climate change is likely to modify nesting conditions for the 
entire DPS. Impacts likely range from small changes in nesting metrics 
to large losses of productivity. As the DPS is already experiencing 
nesting habitat loss due to coastal erosion, we conclude that climate 
change is a threat to the NE Indian DPS.

Conservation Efforts

    There are numerous efforts to conserve the leatherback turtle. The 
following conservation efforts apply to the NE Indian DPS (for a 
description of each effort, please see the section on conservation 
efforts for the overall species): Association of Southeast Asian 
Nations Ministers on Agriculture and Forestry, Andaman and Nicobar 
Island Environmental Team, The Centre for Herpetology/Madras Crocodile 
Bank Trust, Convention on the Conservation of Migratory Species of Wild 
Animals, Convention on Biological Diversity, Convention on 
International Trade in Endangered Species of Wild Fauna and Flora, 
Convention Concerning the Protection of the World Cultural and Natural 
Heritage (World Heritage Convention), FAO Technical Consultation on Sea 
Turtle-Fishery Interactions, The Indian Ocean Tuna Commission, Indian 
Ocean--South-East Asian Marine Turtle Memorandum of Understanding, 
MARPOL, IUCN, Memorandum of Agreement between the Government of the 
Republic of the Philippines and the Government of Malaysia on the 
Establishment of the Turtle Island Heritage Protected Area, Memorandum 
of Understanding on Association of South East Asian Nations Sea Turtle 
Conservation and Protection, The Memorandum of Understanding of a Tri-
National Partnership between the Government of the Republic of 
Indonesia, the Independent State of Papua New Guinea and the Government 
of Solomon Islands, National Sea Turtle Conservation Project in India, 
Ramsar Convention on Wetlands, UNCLOS, and UN Resolution 44/225 on 
Large-Scale Pelagic Driftnet Fishing. Although numerous conservation 
efforts apply to the turtles of this DPS, they do not adequately reduce 
its risk of extinction.

Extinction Risk Analysis

    After reviewing the best available information, the Team concluded 
that the NE Indian DPS is at high risk of extinction. The once large 
nesting aggregation in Malaysia is now functionally extirpated. The 
total index of nesting female abundance is 109 females at all monitored 
beaches. This estimate is likely low because several nesting sites were 
not included in the calculation due to lack of consistent, standardized 
monitoring over multiple and entire nesting seasons. Still, the low 
nesting female abundance places this DPS at risk of stochastic or 
catastrophic events that increase its extinction risk. The DPS once 
exhibited much greater nesting female abundance, which has dramatically 
declined in recent decades. It currently exhibits a slightly declining 
nest trend at monitored nesting beaches in India. The DPS exhibits 
average productivity metrics, such as body size, clutch size and 
frequency. Though it exhibits some spatial distribution and diversity, 
with multiple foraging sites and relatively high genetic diversity at 
the sampled locations, nesting only occurs on islands. Based on these 
factors, we find the DPS to be at risk of extinction as a result of 
past threats.
    Current threats further contribute to the risk of extinction of 
this DPS. Major threats to the DPS include fisheries bycatch and the 
harvest of turtles and eggs. There are not many nests to exploit, but 
evidence suggests that if such nests are found by humans, the eggs are 
at risk of being harvested. Egg harvest led to the extirpation of the 
largest nesting aggregation (i.e., Malaysia), and current 
overexploitation occurs in Thailand, Vietnam, and Sri Lanka. The 
poaching of turtles is also a threat in Myanmar. Fisheries bycatch is a 
major threat, with turtles being captured in trawl and gillnet 
fisheries in Malaysia, India, Thailand, Sri Lanka, Bangladesh, and 
Indonesia. Erosion on the Andaman and Nicobar Islands, as a result of 
tsunami damage, has significantly reduced available nesting habitat. 
Additional habitat modifications include coastal development and 
artificial lighting, as a result of increases in tourism. Pollution and 
climate change are threats that likely affect the DPS by reducing 
abundance and productivity, though the best available data do not allow 
for quantification of those effects. Though many regulatory mechanisms 
are in place, they do not reduce the impact of threats to levels that 
ensure the continued existence of the DPS.
    We conclude, consistent with the team's findings, that the NE 
Indian DPS is currently in danger of extinction. Its low nesting female 
abundance makes the DPS highly vulnerable to threats. Dramatic declines 
in over the past several decades contribute to our concern over the 
continued persistence of the DPS. Past egg and turtle harvest initially 
reduced the nesting female abundance of this DPS and likely confined 
its nesting habitat to a few island beaches, with little diversity and 
reduced spatial distribution. The present, ongoing threats include: 
overutilization (i.e., turtle and egg harvest); fisheries bycatch; loss 
of habitat; and predation. Overutilization and fisheries bycatch 
reduces abundance and productivity (i.e., imminent and substantial 
demographic risks) by removing mature and immature individuals from the 
population at rates exceeding replacement. The loss of nesting habitat 
and predation (of eggs) reduces productivity and the DPS's ability to 
recover to its previous abundance. Though numerous conservation efforts 
apply to this DPS, they do not adequately reduce the risk of 
extinction. We conclude that the NE Indian DPS is in danger of 
extinction throughout its

[[Page 48387]]

range and therefore meets the definition of an endangered species. The 
threatened species definition does not apply because the DPS is at risk 
of extinction currently (i.e., at present), rather than on a trajectory 
to become so within the foreseeable future.

West Pacific DPS

    The Team defined the West Pacific DPS as leatherback turtles 
originating from the West Pacific Ocean, south of 71[deg] N, north of 
47[deg] S, east of 120[deg] E, and west of 117.124[deg] W. The northern 
and southern boundaries reflect the highest latitude occurrences of 
leatherback turtles in each hemisphere (Goff and Lien 1988; Carriol and 
Vader 2002; McMahon and Hayes 2006; Shillinger et al. 2008; Benson et 
al. 2011; Eckert et al. 2012). We placed the western boundary at 
120[deg] E to approximate the Wallace and Huxley lines, which are 
established biogeographic barriers to gene flow between Indian and 
Pacific Ocean populations of numerous species. While the genetic 
differences between the Northeast Indian and West Pacific DPSs 
demonstrate discreteness, genetic sampling is unavailable from areas 
where the nesting ranges of the DPSs likely meet, preventing us from 
defining the boundary more specifically. We placed the eastern boundary 
at the border between the United States and Mexico to reflect the DPS's 
wide foraging range throughout the Pacific Ocean. We chose this border 
because the West Pacific DPS crosses the ocean to forage in the eastern 
Pacific Ocean, including in waters of the United States, whereas the 
East Pacific DPS forages primarily off the coasts of Central and South 
America. The two DPSs overlap in foraging habitats off waters of Chile 
and Peru (Donoso and Dutton 2010).
    The range of the DPS (i.e., all areas of occurrence) extends 
throughout the Pacific Ocean with specific coastal and pelagic areas in 
the Indo-Pacific basin providing important foraging and migratory 
habitats. Documented nesting occurs on beaches of the following 
nations: Indonesia, Papua New Guinea, Solomon Islands, and Vanuatu. 
Leatherback turtles of the West Pacific DPS migrate through the EEZs of 
at least 32 nations including in the U.S. EEZs of California and 
Hawaii, spending between 45 and 78 percent of the year on the high seas 
(Harrison et al. 2018). Of the 32 nations, the West Pacific DPS 
migrates through at least 18 nations or territories of the western and 
southwestern Pacific Ocean: Indonesia, Papua New Guinea, Solomon 
Islands, Philippines, Malaysia, Vietnam, Japan, Palau, Micronesia, 
Marshall Islands, Northern Mariana Islands and Guam, Fiji, Vanuatu, 
Australia, New Caledonia, New Zealand, Line Islands, and Kiribati 
(Harrison et al. 2018). Foraging occurs in seven ecoregions: South 
China/Sulu and Sulawesi Seas, Indonesian Seas, East Australian Current 
Extension, Tasman Front, Kuroshio Extension of the Central North 
Pacific, equatorial Eastern Pacific, and California Current Extension 
(Benson et al. 2011). Individuals demonstrate fidelity to these 
foraging areas, likely as a result of their post-hatchling dispersal 
patterns and nesting season (Benson et al. 2011; Gaspar et al. 2012; 
Gaspar and Lalire 2017; Harrison et al. 2018).
    Leatherback turtles of the West Pacific DPS nest in tropical and 
subtropical latitudes primarily in Indonesia, Papua New Guinea, and 
Solomon Islands, and a lesser extent in Vanuatu (Dutton et al. 2007; 
Benson et al. 2007a; Benson et al. 2007b; Benson et al. 2011). The 
majority of nesting occurs along the north coast of the Bird's Head 
Peninsula, Papua Barat, Indonesia at Jamursba-Medi and Wermon Beaches 
(Dutton et al. 2007). A recent discovery of a previously undocumented 
nesting area on Buru Island, Maluku Province, Indonesia (WWF 2018) 
suggests that additional undocumented nesting habitats may exist on 
other remote or infrequently surveyed islands of the western Pacific 
Ocean. This DPS nests year round, and exhibits a bimodal nesting 
strategy whereby a proportion of females nest during November through 
February (i.e., ``winter'' nesting females) and other females nest May 
through September (i.e., ``summer'' nesting females; Benson et al. 
2007a; Benson et al. 2007b; Dutton et al. 2007; Tapilatu and Tiwari 
2007; Benson et al. 2011).
    Nesting beach habitats throughout the West Pacific are generally 
dynamic, high profile beaches associated with deep water approaches and 
strong waves. Beaches can be quite narrow as in parts of the Solomon 
Islands or Papua New Guinea, or broad as in the case of Jamursba-Medi, 
Indonesia during the summer months. Nesting females appear to prefer 
coarse-grained sand free of rocks, coral, or other abrasive substrates 
(reviewed by Eckert et al. 2012).
    While West Pacific leatherback turtles do not have distinct 
``migratory corridors,'' several areas are considered ``areas of 
passage'' used by turtles traveling between nesting and foraging 
locations, and there is clear separation of migratory and foraging 
destinations based on nesting season (Benson et al. 2007a, b; Benson et 
al. 2011; Harrison et al. 2018). Post-nesting, winter nesting females 
from Papua New Guinea, Indonesia, and Solomon Islands migrate through 
the Halmahera, Bismarck, Solomon, and Coral Seas, towards Southern 
Hemisphere temperate and tropical foraging areas in the Tasman Sea, 
East Australian Current, and western South Pacific Ocean (Benson et al. 
2011; Harrison et al. 2018; Jino et al. 2018). Genetic analyses of 
leatherback turtles caught in fisheries off Peru and Chile indicates 
that approximately 15 percent of sampled individuals originate from the 
West Pacific DPS, likely winter nesting females that have migrated 
across the Southern Hemisphere to the productive waters off South 
America (Donoso and Dutton 2010; NMFS unpublished data 2018). It is 
unclear what proportion of the West Pacific DPS might utilize this area 
and how important it might be to this DPS.
    Leatherback turtles migrate through and forage in the waters of the 
Philippines (Benson et al. 2007a, 2011; MRF 2010, 2014). In 2005, 
Salinas et al. (2009) found a female in San Fernando (close to El Nido) 
that had been previously tagged at Jamursba-Medi in July 2003. The 
Marine Research Foundation (MRF) utilized aerial transects to assess 
leatherback foraging area use in Palawan waters and off the coast of 
Borneo (MRF 2010, 2014). They found leatherback turtles (n = 28 in 2010 
and 2013/2014) foraging in nearshore waters around the NE and SE coasts 
of Palawan, potentially linked to large jellyfish aggregations from 
February to May, and overlapping with high density fishing activity in 
Taytay Bay, off NE Palawan (MRF 2010, 2014). Additionally, numerous 
leatherback turtle marine sightings, strandings, and fishery bycatch 
(typically entangled in gillnet gear) exist for locations throughout 
the Philippines including Marine Wildlife Watch of the local NGO, 
Marine Wildlife Watch of the Philippines, from 2010 to 2018 (Bagarinao 
2011; Cruz 2006; MRF 2010; MWWP unpublished data 2018).

Abundance

    The total index of nesting female abundance of the West Pacific DPS 
is 1,277 females. We based this total index on two nesting aggregations 
in Jamursba-Medi and Wermon, Indonesia (Tapilatu et al. 2013; Tiwari et 
al. in prep). Our total index does not include 18 unquantified nesting 
aggregations in Indonesia, Papua New Guinea, Solomon Islands, and 
Vanuatu. To calculate the index of nesting female abundance (723 
females) for Jamursba-Medi (i.e., a 18 km stretch of beach that has 
been monitored since 1981), we divided the total number of nests 
between the 2015/2016 and 2017/2018 nesting seasons (i.e., a 3-year 
remigration interval) by the clutch frequency (5.5 clutches per

[[Page 48388]]

season; Tapilatu et al. 2013). We performed a similar analysis for data 
from Wermon (index = 554 females), a 6 km beach that has been monitored 
since 2002.
    Based on the Tapilatu et al. (2013) study, the IUCN Red List 
assessment estimated the total number of mature individuals (including 
females and males) utilizing Jamursba-Medi and Wermon beaches to be 
1,438 leatherback turtles (Tiwari et al. 2013). The IUCN estimate 
includes males and thus is higher than ours. Curtis et al. (2015) 
provided a minimum annual nesting female estimate of 318 females (or 
954 total nesting female abundance over a 3-year remigration interval). 
Dutton et al. (2007) estimated that 1,113 females may have nested 
annually, or conservatively 2,700 total nesting females, in the entire 
western Pacific population. At that time, they estimated 75 percent of 
the population originated from Bird's Head Peninsula (or approximately 
2,025 females; Dutton et al. 2007). Our total index is within the range 
of published estimates of abundance for this DPS, taking into account 
differences in survey methods over time, and is based on the best 
available data for the DPS at this time.
    Within the nesting range of this DPS, nest monitoring activities 
have occurred relatively recently, with standardized methods in Papua 
Barat first implemented in 2002 (Hitipeuw et al. 2007; Tapilatu et al. 
2013). Outside the Bird's Head Peninsula, monitoring has been sporadic, 
opportunistic, and spatially limited because the region is vast, 
remote, and logistically challenging to access. Often nesting beaches 
are located far from towns or cities, where there are no roads to, or 
electricity in, adjacent villages. Cultural and socio-economic dynamics 
confound monitoring programs, which are dependent upon fiscal 
sponsorship, incentives, community buy-in, and the degree of 
familiarity of local communities with concepts of sustainability or 
conservation (Kinch 2006; Gjersten and Pakiding 2012). While Jamursba-
Medi and Wermon beaches have been monitored fairly consistently over 
time, less is known about the status and trends of nesting beaches in 
Papua New Guinea, Solomon Islands, and Vanuatu. Records are further 
confounded by changes in place names and jurisdictional boundaries over 
recent decades (e.g. the Indonesian province formerly known as Irian 
Jaya is currently two provinces of Papua and Papua Barat). Village 
names or location descriptions have also changed over time, and 
geographic coordinates were not recorded historically. Therefore, all 
estimates of abundance in this DPS carry substantial uncertainty.
    In Indonesia, aerial surveys provided the first indication of 
leatherback nesting in Papua (i.e., Irian Jaya; Salm 1982). At that 
time, Salm (1982) did not provide location details out of concern that 
public disclosure prior to protection would be detrimental. Follow-up 
studies during the 1980s and 1990s indicated that a large nesting 
population was located along the coastal beaches of northern Papua or 
Papua Barat, Bird's Head Peninsula (Bhaskar 1985). Systematic 
monitoring of leatherback turtles began during the early 1990s, 
primarily in the form of annual nest counts (Hitipeuw et al. 2007). On 
the Bird's Head Peninsula of Papua Barat, nesting occurs mainly at 
Jamursba-Medi and Wermon, where a total of 1,371 nesting females were 
tagged between 2002 and 2011 (Tapilatu et al. 2013). The primary 
nesting season at Jamursba-Medi occurs during the summer (May to 
September), while nesting occurs year round at Wermon with a small peak 
in July and primary nesting activity during the winter, between 
November and February (Hitipeuw et al. 2007). Historically, 
approximately 60 percent of nesting activity occurs at Jamursba-Medi 
with 40 percent of activity at Wermon (Tapilatu et al. 2013). While a 
few females have been documented nesting at both beaches during a 
nesting season (Tapilatu et al. 2013), the vast majority of females do 
not appear to utilize both Jamursba-Medi and Wermon Beaches during a 
single nesting season (Tapilatu and Tiwari 2007; Tapilatu et al. 2013; 
Lontoh 2014). Based on nest counts and clutch frequency per season 
(mean = 5.5 +/- 1.6 nests per female), approximately 464 to 612 females 
nested at Jamursba-Medi and Wermon in 2011 (Tapilatu et al. 2013). 
Additional low-level nesting activity in Indonesia occurs in the 
Manokawari region of the Bird's Head Peninsula to the east of the 
Jamursba-Medi and Wermon Beaches (Suganuma et al. 2012). Between 2008 
and 2011, 84 to 135 nests were recorded, or a mean of about 117 nests 
annually (Suganuma et al. 2012). However, survey effort was limited and 
not consistent across years and may underestimate total nesting 
activity. Further it is unknown whether interchange occurs between 
turtles nesting in the Manokawari region and those of the Bird's Head 
Peninsula index beaches. In 2016, nesting activity was identified in 
Central Maluku at Buru Island, west of Bird's Head Peninsula. In 2017, 
a monitoring program to quantify nesting activity was initiated on 
three north coast beaches of Buru Island (totaling 10 km) which 
documented 203 nests, and preliminary data indicates that there might 
be two nesting peaks: May through July and November through February 
(WWF 2018). Nesting activity in other areas of Indonesia are known or 
suspected, but unquantified (Dutton et al. 2007; Tapilatu 2017).
    In Papua New Guinea, the majority of known nesting activity occurs 
during the winter months (November to February) along the Huon Coast on 
the northeastern coast of the Morobe Province, where 576 females were 
tagged between 1999 and 2013 (Pilcher 2006, 2008, 2009, 2010, 2011, 
2012, 2013; Pilcher and Chaloupka 2013). Aerial surveys along the Huon 
Coast in January and December between 2004 and 2006 documented 276 
nests, with an estimate of 500 nests per season (Benson et al. 2007b; 
Dutton et al. 2007). During the Huon Coast Leatherback Turtle Project, 
which took place between 2005 and 2012, an average of 258 nests were 
laid per season (range: 193 to 527) at seven beaches which comprised 
approximately 35 km of nesting habitat along the Huon Coast (Pilcher 
2013; WPRFMC 2015). One challenge in estimating nesting activity in 
Papua New Guinea is that leatherback site fidelity appears to be 
variable, with some satellite tagged animals seen visiting a number of 
areas during one nesting season (Benson et al. 2007b). For example, a 
number of Huon Coast nesting females visited other nearby beaches and 
east-facing beaches of the Huon Peninsula, including Bougainville and 
Woodlark Islands during a single nesting season (Benson et al. 2007b). 
Therefore, for assessment purposes, we consider the Huon Coast to be 
one nesting beach complex.
    Additional nesting activity occurs in other areas of Papua New 
Guinea, such as along the north coast of the Madang Province and on 
several islands including Manus, Long, New Britain, Bougainville, New 
Ireland, and Normanby (Prichard 1982; Spring 1982; Benson et al. 2007b; 
Dutton et al. 2007). In these areas nesting activity has not been 
quantified via standardized or consistent methods, but information has 
been obtained via community surveys, aerial surveys, or rapid 
assessments. Nesting occurs primarily in the winter months, although 
low-level year-round nesting may also occur (Spring 1982; Dutton et al. 
2007). Approximately 50 nests may be laid annually along the north 
coast of the Madang Province (Benson et al. 2007b; TIRN 2017). The 
Islands of New Britain and Bougainville may host approximately 140 to 
160

[[Page 48389]]

nests per year, respectively (Benson et al. 2007b; Dutton et al. 2007; 
Kinch et al. 2009). On Bougainville Island, aerial surveys conducted 
during the 2005 and 2007 nesting seasons documented a mean of 68 nests 
(range: 41 to 107 nests) or an extrapolated estimate of 160 to 415 
nests per year (Dutton et al. 2007; Benson et al. 2007b). In 2009, a 
one week full-island ground survey (conducted by boat and foot) 
recorded 46 leatherback nests (Kinch et al. 2009).
    In the Solomon Islands nesting activity is distributed throughout 
the country with the majority of nesting activity at Sasakolo and 
Litogarhira beaches on Isabel Island, and on Rendova and Tetepare 
Islands in the Western Province (Pita 2005; Dutton et al. 2007; Benson 
et al. 2018a). The nesting season occurs primarily during winter 
(November through February), although some year-round nesting has been 
documented (Pilcher 2010b; Williams et al. 2014; Jino et al. 2018; TNC-
Solomon Islands 2018 unpublished). Leatherback turtle monitoring was 
begun by the Solomon Island Department of Fisheries in 1989 (Pita 
2005). Between 1999 and 2006, an estimated 640 to 700 nests were laid 
annually in the Solomon Islands, representing approximately eight 
percent of the total western Pacific leatherback nesting at that time 
(Dutton et al. 2007). At Sasokolo Beach, Isabel Island, during a 54 day 
monitoring period between November 28, 2000 and January 21, 2001, 132 
nests were documented with an additional 35 nests present when 
monitoring began (Ramohia et al. 2001). Between December 27, 2006 and 
January 2, 2007, aerial surveys provided seasonal estimates of 207 
nests laid on Isabel Island, and an additional 312 nests on other 
islands (Benson et al. 2018a). A January 2011 site visit resulted in 
315 nests identified at Sasakolo and Litogahira (Tiwari 2011 
unpublished). Recently, nesting activity has also been documented at 
the southeastern side of Isabel, where approximately 52 females may 
nest annually (TNC-Solomons 2018 unpublished). Since 2002, the Tetepare 
Descendants' Association (TDA) has monitored nesting activity 
opportunistically in the Solomon Islands, where approximately 30 to 50 
leatherback nests are laid seasonally on two beaches (Goby et al. 
2010). Between July 1, 2012 and April 30, 2013, TDA undertook 257 beach 
surveys and found 44 leatherback nests (TDA 2013). While monitoring 
efforts may be ongoing, data management and analysis remains a key 
challenge for these isolated communities (TDA 2013; Pilcher 2010b). At 
Rendova Island during the 2003/2004 winter nesting season, 235 
leatherback turtle nests were recorded, and during the 2009/2010 
season, 79 nests were laid (Pilcher 2010b; Goby et al. 2010). Likely 
the most comprehensive surveys occurred from September 1, 2012 to April 
30, 2013 (91 patrols, 3 days per week), which documented a total of 74 
nests (TDA 2013). During the 2017/2018 winter nesting season, 29 nests 
were documented (Solomon Islands Community Conservation Partnership 
2018 unpublished data). The community on Vangunu Island documented a 
total of 23 nests and 11 females between June 2011 and July 2014 (Jino 
et al. 2018). Nesting occurred during two distinct seasons from May to 
July and from November to January, and of the females tagged, one 
nested successfully six times and another nested five times (Jino et 
al. 2018). The other nine turtles were only observed nesting once or 
twice, and it is likely that either some nesting events were not 
recorded or the females nested on surrounding unmonitored beaches (Jino 
et al. 2018). On Malaita Island at Waisurione beach, nesting activity 
occurs during the summer (June to August), but only a few females were 
determined to use the area, with five and seven nests documented in 
2014 and 2015, respectively (Williams et al. 2014).
    Nesting occurs in low numbers at other islands in the western 
Pacific Ocean. In Vanuatu, 30 to 40 nests are laid annually on Epi and 
Ambrym Islands (Dutton et al. 2007; Petro et al. 2007; WSB 2011), 
although fewer nests (n = 15) were documented during the 2014/2015 
nesting season (WSB 2016). Leatherback turtles have been reported in 
Fiji (Rupeni et al. 2002; NMFS and USFWS 2013; Jino et al. 2018), but 
these accounts involved foraging or in-water capture of animals, and it 
is unclear if historic reports included nesting activity (Guinea 1993; 
Benson et al. 2013). Historical nesting records also exist for the 
eastern coast of Queensland, in New South Wales, and in the Northern 
Territories from December to February (Dobbs 2002; Limpus 2009). 
However, current information was not available at the time of the 
study, and no nests have been observed since 1995 despite regular 
monitoring (Flint et al. 2012). Since the 1980s, there have also been 
reports of leatherback turtles nesting in the Philippines (Cruz 2006; 
MRF 2010). Of recent reports, two documented cases have been confirmed 
by sea turtle experts (i.e., staff of the Marine Wildlife Watch of the 
Philippines). On July 15, 2013, at Barangay Yawah, Legazpi City, Albay, 
NAVFORSOL (the Philippines Naval facility) personnel observed a 
leatherback nesting, but the eggs failed to hatch. On August 6, 2013 at 
Camp Picardo beach, Barangay, Eastern Samar, a nesting event was 
aborted due to disturbance on the beach, but according to the social 
media report (i.e., a Facebook post), the female was tagged and led 
back to sea (MWWP unpublished 2018). Given the low-site fidelity of the 
turtles in this DPS (Benson et al. 2007b), it is not surprising that 
leatherbacks might distribute nests among various areas throughout the 
region.
    The total index of nesting female abundance of the West Pacific DPS 
(i.e., 1,277 females) places it at risk for environmental variation, 
genetic complications, demographic stochasticity, negative ecological 
feedback, and catastrophes (McElhany et al. 2000; NMFS 2017). These 
processes, working alone or in concert, place small populations at a 
greater extinction risk than large populations, which are better able 
to absorb impacts to habitat or losses in individuals. Due to its small 
size, the DPS has restricted capacity to buffer such losses. Given the 
intrinsic problems of small population size, we conclude that the 
nesting female abundance is a major factor in the extinction risk of 
this DPS.

Productivity

    The West Pacific DPS exhibits a declining nesting trend. We 
conducted trend analyses for the two index beaches in Indonesia, which 
were the only two beaches with 9 or more recent years of standardized 
data, with the most recent data collection in 2014 or more recently 
(the standards for conducting a trend analysis in this report). The 
median trend in annual nest counts estimated for Jamursba-Medi (data 
collected from 2001 to 2017) was -5.7 percent annually (sd = 5.4 
percent; 95 percent CI = -16.2 to 5.3 percent; f = 0.867; mean annual 
nests = 2,063). While data are available for the period starting in 
1999, the best available information indicates that beach monitoring 
and nest protection practices improved in 2001; therefore, we used the 
time series starting in 2001. For Wermon (data collected from 2006 to 
2017, excluding 2002-2005 and 2013-2015 due to low or insufficient 
effort), the median trend was -2.3 percent annually (sd = 8.4 percent; 
95 percent CI = -19.8 to 14.9 percent; f = 0.643; mean annual nests = 
1,010). As Jamursba-Medi and Wermon currently represent approximately 
75 percent of nesting for this DPS, we

[[Page 48390]]

consider these declining trends to be representative of the entire DPS.
    Our trend data for Indonesia yield similar results to other 
published findings. The IUCN Red List assessment found a decreasing 
trend of -7 percent annually (Tiwari et al. 2013). Tapilatu et al. 
(2013) identified a -5.5 percent annual rate of decline at Jamursba-
Medi between 1984 and 2011 and a -11.6 percent annual rate of decline 
at Wermon between 2002 and 2011. Between 1986 and 2010, Benson et al. 
(2013) highlighted drastic declines in the annual number of nests at 
Jamursba-Medi and Wermon. Additionally, a 27-year aerial survey study 
indicates a decline in the number of leatherback turtles foraging off 
central California (Benson et al. 2018b). From 1995 to 2003, an 
estimated 12 to 379 individuals (mean = 178) foraged in this area 
(Benson et al. 2007), while from 2004 to 2017, an estimated 23 to 112 
individuals foraged in this area, representing a decline of 5.6 percent 
annually (Benson et al. 2018b).
    At Jamursba-Medi, nesting data have been collected for some years 
since 1981. However, no data were collected during many years in the 
mid-1980s and late 1990s (Tapilatu et al. 2013). There is considerable 
uncertainty in the early estimates, with over 4,000 nests estimated in 
1981, 14,522 nests in 1984, and a dramatic drop to 3,261 nests in 1985 
(Tapilatu et al. 2013). It is unclear if there was sampling 
inconsistency between years or if there was an actual decline in 
nesting activity. However, if analyses are based on the 1984 data, 
during which the greatest number of nests was recorded, there was a 
78.3 percent decline over the past 27 years (1984 to 2011), or 5.5 
percent annual rate of decline (Tapilatu et al. 2013). Alternatively, 
if analysis is based on 2005 to 2011 when the Tapilatu et al. (2013) 
study ensued, nesting activity declined 29 percent from 2,626 nests (in 
2005) to 1,596 nests (in 2011; Tapilatu et al. 2013). Since the 
Tapilatu et al. (2013) study, University of Papua scientists have 
continued to engage with local communities to monitor nesting activity. 
The overall nesting trend has continued to decline by 5.6 percent per 
year between 2003 and 2017. However, there appears to be an increase in 
nesting since 2013 (Tiwari et al. in prep).
    The first comprehensive surveys at Wermon beach in 2002 found 
almost as many nests laid on Wermon as on Jamursba-Medi (Hitipeuw et 
al. 2007). At that time, it was hypothesized that the decline at 
Jamursba-Medi may have been offset by an increase at Wermon (Hitipeuw 
et al. 2007). However, Tapilatu et al. (2013) found a significant 
decline in nesting at Wermon from 2,994 nests in 2002 to 1,096 nests in 
2011 (62.8 percent total or 11.6 percent annual rate of decline). 
Unfortunately, no monitoring activities occurred at Wermon between 2013 
and 2015 due to community discord, which prevented beach access. 
Between 2006 and 2017, nesting has continued to decline at 
approximately 2.3 percent (Tiwari et al. in prep). However, there may 
have been a slight increase in recent nesting, similar to Jamursba-Medi 
(Tiwari et al. in prep).
    Local residents stated that leatherback turtles were the dominant 
sea turtle species nesting in Maokawari prior to the 1980s, but that 
the population has declined significantly since the 1990s due to 
village development and exploitation of turtles and eggs (Tapilatu et 
al. 2017).
    Data collection in Papua New Guinea spanned 8 years and ended prior 
to 2014. Because these data did not meet our criteria for ``recent,'' 
we did not perform a trend analysis, but included a bar graph in the 
Status Review Report. In Papua New Guinea, nesting activity along the 
Huon Coast was relatively stable between 2005 and 2013, with 193 to 527 
nests per year (mean annual nests = 258) and with most nesting activity 
occurring at two primary areas, Busama and Kamiali (Pilcher 2013; 
Benson et al. 2015; WPRFMC 2015). Given the exchange of females and 
evidence of multiple beach use among females in Papua New Guinea 
(Benson et al. 2007b), we consider the Huon Coast to be one nesting 
area and not individual nesting beaches. Though there have been several 
independent studies of abundance over time, we determined that these 
data are inadequate to incorporate into a trend analysis because these 
data do not meet our criteria (i.e., nest count data consistently 
collected in a standardized approach for at least 9 years). For 
historical perspective, leatherback turtle nesting along the Huon Coast 
was first identified south of the city of Lae near the Buang River, at 
an area likely between Labu Tale and Busama villages (i.e., Maus Buang 
or Buang-Buassi; Bedding and Lockhart 1989; Quinn and Kojis 1985; Hirth 
et al. 1993). Estimates of leatherback turtle nesting at Maus Buang 
during the 1980s ranged from five to 10 turtles per night from November 
to January (Quinn and Kojis 1985) or 300 nests laid annually (Bedding 
and Lockhart 1989). Quinn and Kojis (1985) estimated that 300 to 500 
females may nest annually in Papua New Guinea, although it is unclear 
if estimates were for the Maus Buang area specifically or the Huon 
Coast at large. Hirth et al. (1993) undertook the most standardized 
survey at that time and recorded 76 nests and 34 females nesting at 
``Piguwa'' (i.e., Maus Buang) on 725 meters of beach during a 15-day 
period in December 1989. During the Huon Coast leatherback turtle 
nesting beach program, an average of 35 and 114 nests were laid 
annually during the 4-month nesting season in this similar area at Labu 
Tale and Busama beaches, respectively (Pilcher 2013; WPRFMC 2015). 
Kamiali Beach lies approximately 30 km south of the city of Lae. In 
1996, the Kamiali Wildlife Management Area was declared a protected 
area for leatherback turtles, and the harvest of nests was prohibited 
along 2 km of beach. In 1999, village rangers began opportunistic 
tagging of nesting females at Kamiali. A community-based nesting beach 
monitoring program was established in 2003, which soon grew into the 
Huon Coast Leatherback Turtle Conservation Program (Benson et al. 
2007b; Pilcher and Chaloupka 2013; Kinch 2006). By 2005, monitoring 
activities expanded from Kamiali Beach (approximately 7 km) to seven 
beaches encompassing approximately 35 km of nesting beaches, which 
included an agreement by participating villages to no longer harvest 
eggs (Kinch 2006; Pilcher 2013). Of these seven beaches, Kamiali was 
the nesting beach with the longest running, most consistent monitoring 
within the Huon Coast nesting beach complex. At Kamiali, 194 females 
were tagged between 1999 and 2012, and an average of 77 nests laid per 
winter nesting season between 2005/2006 and 2012/2013 (Pilcher 2010, 
2011, 2012, 2013; Pilcher and Chaloupka 2013). While we are unable to 
interpret an overall trend from these studies, anecdotal reports from 
villagers and historic information indicates that leatherback nesting 
activity was significantly greater in past decades (Benson et al. 
2007b, 2015; Hirth et al. 1993; Kinch 2006; Bellagio Sea Turtle 
Conservation Initiative 2008).
    In the Solomon Islands, it is not possible to estimate nesting 
trends due to non-standardized methods and opportunistic monitoring 
efforts over time. Available datasets cannot be compared due to 
differences in methodology and do not meet our criteria (i.e., nest 
count data consistently collected in a standardized approach for at 
least 9 years). Historically, nesting was reported at more than 15 
beaches in the Solomon Islands, which may have totaled several hundred 
nests per season (McKeown 1977; Vaughan 1981). Currently, nesting 
activity occurs

[[Page 48391]]

primarily in eight locations (Pita 2005; Dutton et al. 2007; Benson et 
al. 2018a; Jino et al. 2018). However, due to the remoteness of these 
areas and lack of systematic surveys, and likely additional 
undocumented nesting beaches, additional low numbers of nesting 
leatherback turtles are likely to exist in Solomon Islands. For 
example, nesting activity was recently identified on Vanugnu Island, 
where 23 nests were recorded and 11 females nested between 2011 and 
2014 (Jino et al. 2018). Additionally, it is unknown to what extent 
females use multiple beaches throughout the Solomon Islands, or those 
in Papua New Guinea, and what proportion of females nest in the summer 
versus winter (Benson et al. 2007b; Jino et al. 2018; TNC-Solomons 2018 
unpublished). While we are unable to interpret an overall trend, local 
villagers indicate that leatherback nesting was greater in past decades 
(Bellagio Sea Turtle Conservation Initiative 2008; Benson et al. 2007b; 
Benson et al. 2015).
    In Vanuatu, anecdotal information suggests that nesting has 
declined over time (Petro et al. 2007). During the 2010/2011 winter 
nesting season, 41 nests were laid at Votlo Beach, Epi Island, and, 
during the 2014/2015 nesting season, three females laid 15 nests (WSB 
2011, 2016). It is not possible to estimate nest trends due to non-
standardized methods and opportunistic monitoring efforts over time, 
which render existing data incomparable and do not meet our criteria 
(i.e., nest count data consistently collected in a standardized 
approach for at least 9 years).
    In addition to an overall declining nest trend, the West Pacific 
DPS exhibits low reproductive output (i.e., low hatching success), due 
in part to a combination of past and current threats (e.g., beach 
erosion, predation, and beach temperatures).
    The DPS exhibits low productivity (i.e., low hatching success), and 
the overall nest trend is declining, likely due to anthropogenic and 
environmental impacts at nesting beaches and in foraging habitats 
(Tiwari et al. 2013). We conclude that the declining nest trend and low 
reproductive output place the DPS at elevated extinction risk, 
especially given the low nesting female abundance.

Spatial Distribution

    The West Pacific DPS nests throughout four countries with a broad, 
diverse foraging range. It exhibits metapopulation dynamics and fine-
scale population structure.
    Aerial surveys conducted between 2004 and 2007 identified 
Indonesia, Papua New Guinea and Solomon Islands as the core nesting 
areas for the DPS (Benson et al. 2007a; Benson et al. 2007b; Benson et 
al. 2011; Benson et al. 2018b). During the nesting season, nesting 
females generally stayed within 300 km or less of these nesting 
beaches, although a few females were documented visiting multiple 
beaches during a nesting season (Benson et al. 2007b). Distributing 
nesting activity among various habitats may help to buffer some of the 
population from impacts at a single nesting area, but the majority of 
females utilize one nesting area during a nesting season (Benson et al. 
2011).
    Migration and foraging strategies vary based on nesting season, 
likely due to prevailing offshore currents and seasonal monsoon-related 
effects experienced by the turtles as hatchlings (Gaspar et al. 2012). 
The lack of crossover among seasonal nesting populations suggests that 
leatherback turtles develop fidelity for specific foraging regions, 
likely based on juvenile dispersal patterns (Benson et al. 2011; Gaspar 
et al. 2012; Gaspar and Lalire 2017). Oceanic currents help to 
structure the spatial and temporal distribution of juveniles and lead 
them to foraging and developmental habitats (e.g., the North Pacific 
Transition Zone) and to undertake seasonal migrations seeking favorable 
oceanic habitats/temperatures and abundant foraging resources, such as 
the central California ecoregion (Gaspar and Lalire 2017). Inter-annual 
or long-term variability in dispersal patterns can influence population 
impacts or resilience to regional or Pacific Ocean perturbations (e.g., 
exposure to fisheries, ENSO events, etc.). Stable isotopes, linked to 
particular foraging regions, confirm nesting season fidelity to 
specific foraging regions (Seminoff et al. 2012). Size differences are 
also apparent, with slightly larger adults appearing to exploit distant 
temperate foraging habitats regardless of nesting season (Benson et al. 
2011; Lontoh 2014).
    Summer nesting females forage in Northern Hemisphere habitats in 
Asia and the Central North Pacific Ocean, while winter nesting females 
forage in tropical waters of the Southern Hemisphere in the South 
Pacific Ocean (Benson et al. 2011; Harrison et al. 2018). This variance 
in foraging strategy results in a foraging range that covers much of 
the Pacific Ocean: Tasman Sea; East Australian Current; eastern and 
western South Pacific Ocean; Indonesian, Sulu and Sulawesi, and South 
China Seas; North Pacific Transition Zone; equatorial currents; and 
central California ecoregion (Benson et al. 2011; Lontoh 2014; Harrison 
et al. 2018; Jino et al. 2018). Different strategies result in 
demographic differences within the DPS which may affect productivity 
and reproductive output. For example, leatherback turtles that exploit 
distant temperate foraging habitats (e.g., central California) may 
require multiple years of seasonal foraging before returning to nesting 
beaches, due to greater energetic demands. In contrast, leatherback 
turtles exploiting geographically closer, year-round prey resources in 
more tropical habitats (e.g., Sulu Sulawesi and South China Seas) may 
remigrate annually (Lontoh 2014).
    The DPS also exhibits genetic population structure. While mtDNA 
analyses of 106 samples from Indonesia, Papua New Guinea, and Solomon 
Islands did not detect genetic differentiation among nesting 
aggregations (Dutton et al. 2007), microsatellite DNA analyses indicate 
fine-scale genetic structure (Dutton 2019; NMFS SWFSC unpublished 
data).
    The wide distribution and variance in foraging strategies likely 
buffers the DPS to some degree against local catastrophes or 
environmental changes that would limit prey availability. The 
distribution of nesting beaches throughout four countries, although 
primarily concentrated in three, helps to buffer the entire DPS from 
major environmental catastrophes, because disturbances are not likely 
to similarly affect all countries during the same seasons. 
Additionally, the fine-scale genetic structure among nesting 
aggregations is indicative of metapopulation dynamics, which may also 
provide the DPS with some resilience.

Diversity

    The West Pacific DPS exhibits genetic diversity, with six 
haplotypes identified in 106 samples from Solomon Islands, Papua Barat 
Indonesia, and Papua New Guinea (Dutton 2006; Dutton et al. 2007; 
Dutton and Squires 2008). This may provide the DPS with the raw 
material necessary for adapting to long-term environmental changes, 
such as cyclic or directional changes in ocean environments due to 
natural and human causes (McElhany et al. 2000; NMFS 2017). The 
population also exhibits temporal nesting diversity, with various 
proportions of the population nesting during different times of the 
year (summer versus winter) which helps to increase resilience to 
environmental impacts. The foraging strategies are also diverse, with 
turtles using seven

[[Page 48392]]

ecoregions of the Pacific Ocean. Diverse foraging strategies likely 
provide some resilience against local reductions in prey availability 
or catastrophic events, such as oil spills or typhoons, by limiting 
exposure from a single event to only a portion of the DPS. We conclude 
that diversity within the DPS provides it with some level of resilience 
to threats.

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    The destruction or modification of habitat is a threat to this DPS. 
Primary impacts to nesting beaches include erosion and ocean 
inundation, which may be caused by natural processes.
    Nesting beaches of the West Pacific DPS are dynamic, high profile 
beaches that are subject to erosion, such as during King Tides 
(naturally occurring, predictable highest tides), which are common 
seasonal occurrences. In Indonesia, the Bird's Head Peninsula beaches 
are also subject to seasonal patterns of erosion and accretion. Changes 
in the currents brought on by monsoons beginning in September cause 
major erosion at Jamursba-Medi that often removes the entire beach, 
making the habitat unsuitable for nesting until accretion begins again 
in March (Hitipieuw et al. 2007). This natural erosion has been 
documented to impact many nests at Jamursba-Medi (Hitipeuw et al. 
2007). Arguably, western Pacific leatherbacks have been dealing with 
such changes in beach habitats over time, and a turtle's long 
reproductive lifespan in general is designed to sustain nest loss 
during a few bad years or seasons. For example, during the 2003/2004 
nesting season, 80 percent of marked nests at Jamursba-Medi (Warmamedi 
beach) washed away before they hatched (Hitipeuw et al. 2007). However, 
given the low abundance of the population, the loss (or continued loss 
over time) of nests is a concern.
    At Wermon, the inundation of nests from high tides is a threat 
during the winter months. During the 2008/2009 winter nesting season, 
26 percent of nests laid at Wermon were inundated by tidal activity 
(Wurlianty and Hitipeuw 2009). During the 2004/2005 nesting season, 23 
percent of nests were lost to inundation (Wurlianty and Hitipeuw 2005). 
During the 2003/2004 nesting season, 10.7 percent of all nests at 
Wermon were below the high water mark and were subsequently washed away 
by high tides (Hitipeuw et al. 2007). Tapilatu and Tiwari (2007) 
stressed that any management plan developed for Papua will need to 
address the impact of inundation and beach erosion.
    Beach erosion is also a threat to nests in Papua New Guinea, where 
strong storms and tidal surges result in substantial erosion and 
changes to beaches throughout the Huon Coast. For example, much of the 
Labu Tale nesting beach was lost to erosion during the 2012/2013 
nesting season (Pilcher 2013). The differences in beach width along the 
Huon Coast place some beaches at more risk of inundation and erosion, 
such as Kamiali Beach, which is half the width and significantly 
narrower than Busama Beach (Pilcher 2008). At Kamiali, the average 
distance of nests to the sea was 3.2 m, compared to 6.2 m at Busama; 
the distances to the vegetation line were comparable across sites (1.3 
m and 1.7 m, respectively; Pilcher 2013).
    In Vanuatu, there has been low hatching success in some years due 
to storms, floods, and high water (Petro et al. 2007; WSB 2016).
    In recent years, management and conservation practices have 
included relocating erosion-prone nests to bolster hatchling 
production. However, these projects are funding-dependent throughout 
the range of the West Pacific DPS. At Jamursba-Medi, ``doomed'' nests 
(i.e., those that are likely to be lost to erosion or inundation) are 
sometimes relocated to a more stable section of beach; 15 nests were 
relocated during the 2017 summer nesting season (Tiwari et al. in 
prep.). At Wermon, nests are relocated to avoid erosion and tidal 
inundation, and increasingly due to Ipomea root invasion (Tiwari et al. 
in prep), but beach management activities are project-dependent. At 
Wermon during the 2017/18 winter nesting season, nests could not be 
relocated because of the lack of permission from the beach owners, and 
all but three nests washed away (Tiwari et al. in prep). In Papua New 
Guinea, 22 of 47 nests (47 percent) at Kamiali beach were relocated to 
protect them from storm surge and erosion during the 2011/2012 nesting 
season, and 41 percent of nests were relocated during the 2009/2010 
season (Pilcher 2012). In the Solomon Islands, efforts to relocate 
``doomed'' nests is an ongoing and necessary management strategy to 
help bolster hatchling production, given that a large proportion of 
nests are inundated or have very low hatching success (Goby et al. 
2010; TDA 2013; Jino et al. 2018).
    A large, significant portion of nests (i.e., 10.7 percent to nearly 
all) are exposed to the reduction and modification of nesting habitat, 
as a result of erosion and inundation. This threat impacts the DPS by 
reducing nesting and hatching success, which has been documented 
throughout the nesting range of the DPS (NMFS and USFWS 2013; Bellagio 
Sea Turtle Conservation Initiative 2008). While West Pacific 
leatherback turtles have undoubtedly evolved to sustain changes in 
beach habitats given their proclivity to select highly dynamic and 
typically narrow beach habitats, and therefore at the population level 
can sustain some level (albeit unquantified level) of nest loss. 
However, the increasing frequency of storms and high water events 
perhaps as a result of climate change can result in increased and 
perhaps unnatural loss of nests. Such impacts may lower the 
productivity of the DPS. Based on the information presented above, we 
conclude that habitat loss and modification is a threat to the DPS.

Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    The primary threat to the West Pacific DPS is the harvest (both 
legal and illegal) of leatherback turtles and their eggs. Leatherback 
turtles are protected by regulatory mechanisms in all four nations 
where the DPS nests, but laws are largely ignored and not consistently 
enforced. This is due to the extreme remoteness of beaches, customary 
and traditional community-based ownership of natural resources (which 
includes sea turtles), and overall lack of institutional capacity and 
funding for enforcement. Furthermore, the cultural and socio-economic 
dynamics in these nations confound community buy-in and conservation 
efforts (Kinch 2006; Gjersten and Pakiding 2012; von Essen et al. 
2014). Additionally, there are nuances related to indigenous harvest 
(and the definition thereof), some of which is permitted in these 
nations.
    Turtle poaching affects both nesting females on beaches and turtles 
in their foraging habitats (Bellagio Sea Turtle Conservation Initiative 
2008; Kinch 2009; Suarez and Starbird 1996; Tiwari et al. 2013; WWF 
2018). Turtle poaching has been documented in all four countries where 
this DPS nests. Egg poaching is a well-documented threat (past and 
current) and is widespread throughout the range of the DPS (Bellagio 
Sea Turtle Conservation Initiative 2008; NMFS and USFWS 2013; Tiwari et 
al. 2013; Tapilatu et al. 2017).
    In Indonesia, the poaching of turtles and eggs continues to occur, 
though egg harvest and exploitation of females has been minimized at 
Jamursba-Medi and Wermon beaches due to the presence of monitoring 
programs and educational outreach. Large-scale egg poaching

[[Page 48393]]

occurred at Jamursba-Medi between 1980 and 1993, whereby approximately 
4 to 5 boats per week (from May to August) collected 10,000 to 15,000 
eggs per boat (Tapilatu et al. 2013). Commercial egg harvest has been 
effectively eliminated since beach monitoring was established at that 
beach in 1993 (Hitipeuw et al. 2007). However, recent survey efforts 
indicate that most, if not all, sea turtle eggs (including leatherback 
turtles) are poached at other Bird's Head Peninsula beaches and sold in 
local markets (Tapilatu et al. 2017). At Buru Island, Indonesia, 
between 2016 and 2017, eight females were poached (WWF 2018), and over 
the past 20+ years, three to five nesting females have likely been 
taken annually (J. Wang, NMFS, pers. comm., 2018). In 2017, 114 of 203 
leatherback nests were harvested at Buru Island (WWF 2018). In 2018, 
due to education provided by the newly established WWF program on Buru 
Island, local community-based efforts in four villages now prohibit 
female and egg harvest. While protective laws exist in Indonesia, 
enforcement is largely lacking in areas where monitoring programs do 
not exist.
    In Indonesia, foraging leatherback turtles are also harvested in 
the waters of the Kei Islands, Maluku Province, where a recognized 
indigenous subsistence harvest of immature and adult turtles (average 
size 145 to 170 cm; range 52 to 203 cm) occurs and has likely been a 
key feature of the local traditional culture for centuries (Compost 
1980; Hamman et al. 2006; Hitipeuw and Lawalata 2006, 2008). Within the 
Kei Islands, customary law (``hak adat'') authorizes a ritual 
leatherback turtle hunt in the nine villages of the traditional kingdom 
of the Nufit people. Starbird and Suarez (1994) brought attention to 
this hunt when they reported that approximately 200 turtles were 
harpooned in three months (October to December) of 1994, with as many 
as 13 taken in one day. Over the past three decades, sporadic 
monitoring efforts have estimated that up to 100 individuals are 
harvested annually (Suarez and Starbird 1996; Hitipeuw and Lawalata 
2008; WWF 2018). At one point, it was assumed that harvest pressure had 
declined and was no longer an issue (NMFS and USFWS 2013). However, 
recent surveys indicate that harvest continues, with conservative 
estimates of 431 turtles killed over an 8-year period (an average of 
53.9 turtles annually), typically between August to February (Hitipeuw 
and Lawalata 2008), and at least 103 turtles harvested in 2017 (WWF 
2018). Most concerning perhaps is that some of the turtle meat 
harvested may be commercially sold as dried meat (i.e., leatherback 
``jerky'' locally known as dendeng), which is illegal to sell and 
inconsistent with indigenous traditional practices. Of four genetic 
samples acquired in 1995 from turtles harvested in the Kei Islands, 
three were assigned to Birds Head Indonesian region and the fourth 
sample was not definitive (66 percent probability, with 34 percent 
probability to Solomon Islands), although it could also be from the 
Indian Ocean or from an undetermined location (NMFS SWFSC unpublished 
data 2018).
    In Papua New Guinea, turtle and egg poaching is a major threat 
despite the fact that leatherback turtles have been protected since the 
1976 Fauna (Protection and Control) Act. The illegal take of both eggs 
and turtles likely continues throughout the country due to lack of 
community-based awareness, reliance on traditional community-based 
practices, institutional capacity, and law enforcement (Bellagio Sea 
Turtle Conservation Initiative, 2008). The killing of nesting females 
has also been well documented throughout Papua New Guinea (Bellagio Sea 
Turtle Conservation Initiative 2008; Kinch 2009; Pilcher 2013). For 
example, at Bougainville Island, surveys of community members 
identified that 21 nesting females were poached during the last decade 
(Kinch 2009). However, the harvest of eggs is likely the most prolific 
threat in Papua New Guinea. If unprotected, egg harvest (compounded by 
intense dog predation described below) resulted in the loss of 70 to 
100 percent of nests (Quinn and Kojis 1985; Hirth 1993; Bellagio Sea 
Turtle Conservation Initiative 2008; Pilcher 2013). For example, during 
a one week survey in January 2009 at Bougainville Island, almost 100 
percent of the 46 documented nests were poached (Kinch 2009). It is 
likely that near total egg collection occurred throughout the Huon 
Coast between World War II and the establishment of the Huon Coast 
Leatherback Turtle Monitoring and Conservation Program in 2003 
(Bellagio Sea Turtle Conservation Initiative 2008; Pilcher and 
Chaloupka 2013; Pilcher 2013). The Huon Coast Project, which operated 
between 2003 and 2013, helped to reduce egg and turtle harvest due to 
program involvement and community incentive funds received in exchange 
for non-harvest agreements (Pilcher 2013). As a result of the program, 
hatchling production (i.e., percent of eggs yielding hatchlings) 
increased from zero to approximately 60 percent (Pilcher 2009, 2011, 
2013; WPRFMC 2015). The Project ended in 2013, and unfortunately egg 
harvest resumed since there was no incentive for communities to 
maintain their no-harvest agreements (John Ben, Huon Coast Leatherback 
Turtle Project, pers. comm., 2020).
    In Vanuatu and the Solomon Islands, the poaching of females and 
collection of eggs is also well documented (Bellagio Sea Turtle 
Conservation Initiative 2008; NMFS and USFWS 2013). In Vanuatu, MacKay 
et al. (2014) reported the harvest of five nesting females between 1999 
and 2008. However there is a general understanding that nesting females 
were typically harvested (Petro et al. 2007). Of the 315 nests 
documented on Isabel Island, Solomon Islands during a January 2011 site 
visit at Sasokolo and Litogahira beaches, the majority of nests had 
been poached (Tiwari 2011 unpublished data). Historically, nearly all 
nesting females and eggs were poached on Redova for consumption (Tiwari 
2011 unpublished data). In response, financial incentive programs have 
been established to protect nests and females whereby villagers are 
paid a financial reward for each nest that hatches successfully (TDA 
2013). On Vangunu Island, 10 to 20 nesting females were poached 
annually, in addition to near-total egg collection (Jino et al. 2018). 
In response to declining population trends, the community declared a 
moratorium on the harvest of leatherback turtles in 1999 (Jino et al. 
2018), and a community incentive program providing financial awards has 
helped to reduce harvest pressure (TDA 2013). Despite these efforts and 
protective legislation, the poaching of females and eggs likely 
persists throughout the Solomon Islands (TDA 2013: Tiwari 2011 
unpublished; MacKay et al. 2014).
    Within the West Pacific DPS, many nesting females, foraging 
turtles, and eggs are exposed to both illegal poaching and legal 
harvest. The taking of turtles reduces abundance. The taking of nesting 
females reduces both abundance and productivity. Such impacts are high 
because they directly remove the most productive individuals from the 
DPS, reducing current and/or future reproductive potential. Egg harvest 
reduces productivity; the persistent, and near-total (at some 
locations) collection of eggs guarantees that future population 
recruitment (i.e., nesting female abundance) will be reduced or 
eliminated. Given the declining nesting trend and current nesting 
female abundance of this DPS, the continued and unregulated poaching

[[Page 48394]]

or harvest of leatherback turtles and eggs is unsustainable. Further, 
the harvest of approximately 100 foraging leatherback turtles annually 
at the Kei Islands, Indonesia is likely an unsustainable practice given 
the current low abundance of the population. We conclude that 
overutilization is a major, and the primary, threat to the West Pacific 
DPS, accelerating its risk of extinction.

Disease or Predation

    While we could not find any information on disease for this DPS, 
predation of eggs is a major and well-documented threat to the West 
Pacific DPS, likely second to poaching (i.e., nests not taken by humans 
are typically predated; Bellagio Sea Turtle Conservation Initiative 
2008).
    In Indonesia, predation of eggs by feral pigs, feral dogs, and 
monitor lizards has been documented, with feral pig predation being the 
most detrimental (Hitipeuw and Maturbongs 2002; Tapilatu and Tiwari 
2007; Bellagio Sea Turtle Conservation Initiative 2008). Nest predation 
by domestic and/or feral dogs has been recorded in both Jamursba-Medi 
and Wermon. Predation of nesting females by crocodiles has also been 
documented at Wermon beach (Bellagio Sea Turtle Conservation Initiative 
2008; UNIPA, pers. comm., 2018). At Jamursba-Medi, between June and 
July of 2005, 29.3 percent of nests were destroyed by pigs (Tapilatu 
and Tiwari 2007). Intensive management effort at Jamursba-Medi reduced 
feral pig predation of nests to five percent during the 2016 and 2017 
nesting seasons (Tiwari et al. in prep). Feral pigs and dogs depredated 
17.5 percent of all nests at Wermon during the 2003 and 2004 winter 
nesting season (Hitipeuw et al. 2007). At Wermon, 21 percent of nests 
were lost to predation during the 2004/2005 nesting season (Wurlianty 
and Hitipeuw 2005). At Buru Island in 2017, 16 nests were lost to 
predation by dogs, wild boar, lizards, or saltwater crocodiles (WWF 
2018).
    In Papua New Guinea, predators of eggs include feral dogs, monitor 
lizards, and ghost crabs (Kinch 2009). Depredation of nests by village 
dogs was determined to be an intense threat to nests, with dogs 
consuming all nests laid during the 2003/2004 and 2004/2005 nesting 
seasons at Kamiali beach (Pilcher 2006; I. Kelly, NMFS, pers. comm., 
2018). Predation of nesting females by crocodiles has also been 
documented in a number of locations in Papua New Guinea (Bellagio Sea 
Turtle Conservation Initiative 2008; Kinch 2009). To protect nests, 
Huon Coast communities developed and placed bamboo grids over nests to 
prevent dogs from preying on the eggs (Pilcher 2006; 2009). This, along 
with efforts to reduce egg harvest by humans, resulted in increased 
hatching production from zero to approximately 60 percent between 2006 
and 2013, with over 2,300 nests saved producing approximately 100,000 
hatchlings (Pilcher 2009; 2011; 2013; WRFMC 2015). However, this 
project ended in 2013, and it is unknown if egg protection continues, 
or if nest predation has resumed.
    In this DPS, a large proportion of eggs are exposed to predation, 
especially by dogs and pigs. Predation primarily results in the loss of 
eggs, and the impact of this threat is a reduction of productivity. 
Though leatherback turtles generally produce a large number of eggs and 
hatchlings, predation is widespread throughout the range of the DPS, 
and in some areas, predation rates are as high as 100 percent. We 
conclude that predation poses a threat to the West Pacific DPS.

Inadequacy of Existing Regulatory Mechanisms

    The West Pacific DPS is protected by several regulatory mechanisms. 
For each, we review the objectives of the regulation and to what extent 
it adequately addresses the targeted threat.
    Leatherback turtles are protected by legislation in all four of the 
nations where the West Pacific DPS nests (Indonesia, Papua New Guinea, 
Solomon Islands, and Vanuatu). It is generally illegal to harvest 
leatherback turtles and their eggs. However, laws are not typically 
enforced or followed given customary marine tenure systems that dictate 
near-shore rights. Lack of enforcement or implementation of protective 
laws may be due to: Overall lack of in-country institutional capacity 
and funding for enforcement; the extreme remoteness of beaches; 
customary marine tenure or traditional community-based ownership of 
natural resources in these nations (which includes sea turtles; Kinch 
2006; McDonald 2006) and regulatory government-led legislation, which 
may be incompatible with traditional practices (von Essen et al. 2014). 
There are also nuances related to indigenous harvest (and the 
definition thereof), which is not prohibited in these nations. As a 
result, most leatherback nesting beaches with the exception of 
Jamursba-Medi and Wermon (i.e., beaches with established long-term 
monitoring programs) are not currently protected (or only minimally 
protected) from harvest or poaching of eggs, nesting females, or other 
anthropogenic threats.
    In Indonesia, all sea turtles are protected by law, but there are 
allowances for indigenous peoples (although indigenous provisions are 
not clearly defined). The 1990 Government Regulation Act number 5 
concerning the Conservation of the Natural Resources and the Ecosystem, 
makes the trade of protected wildlife illegal, and those found liable 
can be punished to a maximum of 5-year prison term and fined 100 
million Indonesia Rupiah (approximately 6,500 USD). The protection of 
all sea turtle species (Government Regulation No. 7 on Preserving Flora 
and Fauna Species) came into effect in 1999 (Zainudin et al. 2007). The 
use of protected wildlife is allowed for the purposes of research, 
science, and rescue of the wildlife itself. While the trade and 
exploitation of turtles is illegal in Indonesia, there still exists a 
documented harvest of green turtles in Bali, which contributes to 
public confusion regarding sea turtle protections (Westerlaken 2016).
    In Papua New Guinea, the leatherback turtle is the only species 
protected under the 1976 Fauna (Protection and Control) Act, which 
makes killing of leatherback turtles or taking of leatherback turtle 
eggs illegal, with fines of 500 to 1000 kina (approximately 100 to 300 
USD). Any person who buys or sells or offers for sale, or has in 
possession leatherback turtle eggs or meat can also be fined. The Act 
makes provisions for persons with customary rights to take turtles, but 
states that sea turtles cannot be taken, killed, or sold from May 
through July (Kinch 2006). This is typically the nesting season for 
hard-shelled sea turtle species, but leatherback turtles nest primarily 
during the winter months (November to February). As with most 
Melanesian countries, lands are locally-owned and managed, and the 
national government has little influence outside major cities (Kinch 
2006).
    The Solomon Islands Fisheries Act (1993) regulations protect 
nesting turtles and eggs during the breeding season (June to August and 
November to January); prohibit the sale, purchase, or export of sea 
turtle species or their parts; and contain specific protections for 
leatherback turtles. In the Solomon Islands, more than 85 percent of 
the land is held under customary (locally-managed) marine tenure, and 
the vast majority of the population still lives in rural areas making a 
living from the natural resources on those lands. For centuries, 
communities have practiced traditional models of resource stewardship, 
making implementation and enforcement of national regulations nearly 
impossible. Instead, natural

[[Page 48395]]

resource governance must originate from chiefs and village leaders, 
which requires extensive educational outreach to encourage traditional 
approaches that may be supported by legal or `modern' enforcement 
measures (McDonald 2006).
    Fisheries Regulations under the Vanuatu Fisheries Act (2009) 
prohibit the take, harm, capture, disturbance, possession, sale, 
purchase of or interference with any turtle nest (or any turtle in the 
process of nesting) and the import, or export of green, hawksbill, and 
leatherback turtles or their products (shell, eggs, or hatchlings). The 
Act also prohibits the possession of turtles in captivity. A person may 
apply in writing to the Director of Fisheries for an exemption from all 
or any of these provisions for the purposes of carrying out customary 
practices, education, and/or research. Similar to Papua New Guinea and 
the Solomon Islands, natural resource governance in Vanuatu is best 
directed, realized, and implemented at the community level and not via 
national legislation. Fortunately, traditional practices are 
experiencing a renaissance in Vanuatu and may complement current 
regulatory marine resource management efforts (Hickey et al. 2006).
    Throughout the foraging range of the DPS, there are numerous 
regulatory mechanisms that protect turtles within the DPS. These 
include: RFMOs such as the Western and Central Pacific Fisheries 
Commission (WCPFC) and the IATTC and fisheries management regulations 
in 32 nations where this DPS may occur (Harrison et al. 2018). The 
WCPFC adopted a Sea Turtle Conservation and Management Measure (CMM 
2008-03) to mitigate the impacts on turtles from commercial shallow-set 
fisheries operating in the Western and Central Pacific Ocean. The 
measure included the adoption of FAO (2009) guidelines to reduce sea 
turtle mortality through safe handling practices and to reduce bycatch 
by implementing one of three methods by January 2010. The three methods 
to choose from are: (1) Use only large circle hooks with offsets of 
<=10[deg]; (2) use whole finfish bait; or (3) use any other mitigation 
plan or activity that has been approved by the Commission. This sea 
turtle conservation measure is specific to self-identified shallow-
setting, swordfish-targeting fleets. It does not apply to the 
international Pacific longline deep-set tuna-targeting fisheries, which 
comprise the majority of the longline fisheries and are also known to 
interact with leatherback turtles (Lewison et al. 2004; Beverly and 
Chapman 2007; Roe et al. 2014; Wallace et al. 2013). Technical analysis 
of the sea turtle conservation measure found a very small percentage of 
shallow-set fisheries to be in compliance, with less than one percent 
of Western and Central Pacific Ocean longline effort implementing 
mitigation measures, even though approximately 20 percent of longline 
effort consists of shallow sets (Clarke 2017). Further, many RFMO 
members are not meeting the five percent observer coverage requirement 
resulting in limited bycatch reporting (Clarke 2017).
    In summary, regulatory mechanisms exist to protect leatherback 
turtles and their eggs throughout the range of this DPS. However, most 
are inadequate to reduce the threat that they were designed to address 
due to a lack of implementation or enforcement or inclusion of 
provisions for indigenous harvest. Regulations are also misaligned with 
established traditional practice and management systems. As a result, 
poaching and bycatch remain major threats to the DPS. In summary, we 
consider the inadequacy of the regulatory mechanisms to be a threat to 
the DPS.

Fisheries Bycatch

    Fishery bycatch in coastal and pelagic fisheries is a major threat 
to the West Pacific DPS, which is exposed to domestic and international 
fisheries throughout its extensive foraging range. At-sea bycatch of 
leatherback turtles has been documented for a variety of gillnet and 
longline fisheries in the Pacific Ocean, but little is known about the 
total magnitude or full geographic extent of mortality. Satellite 
telemetry studies have identified movements and revealed fidelity to 
foraging regions of the DPS, specifically in habitats of the North 
Pacific Ocean, southwestern Pacific Ocean, and Indo-Pacific tropical 
seas (Bailey et al. 2012; Benson et al. 2011, Seminoff et al. 2012; Roe 
et al. 2014). The summer nesting component of the population exhibits 
strong site fidelity to the central California foraging area (Benson et 
al. 2011) which puts them at risk during migrations of interacting with 
U.S. and international pelagic longline fleets operating throughout the 
Central and North Pacific Oceans. For example, several of the turtles 
tagged in Papua Barat, Indonesia were known or suspected to have been 
killed in fisheries operating off Japan, Philippines, and Malaysia 
(Benson et al. 2011).
    Historically, significant leatherback bycatch was documented in the 
North Pacific high seas driftnet fishery, which expanded rapidly during 
the late 1970s but was banned in 1992 by a UN resolution (summarized in 
Benson et al. 2015). Wetherall et al. (1993) estimated that over 750 
leatherback turtles were killed in Japanese, Korean, and Taiwanese 
driftnet fisheries during the 1990 to 1991 season, with potentially 
5,000 to 10,000 leatherback turtles bycaught between the late 1970s and 
1992. Based on current knowledge of movement patterns (Benson et al. 
2011), the majority of these bycaught turtles would have originated 
from western Pacific nesting beaches after their boreal summer nesting 
period. Thus, high seas driftnet fishery bycatch was likely a 
significant contributor to the population declines observed at nesting 
beaches during the 1980s and 1990s (Benson et al. 2015).
    Many nations are involved in longline fishing in the Pacific Ocean, 
where two types of vessels are used: (1) Large distant-water freezer 
vessels that undertake long (months) voyages and operate over large 
areas of the region; and (2) smaller offshore vessels with ice or chill 
capacity that typically undertake trips of about one month. Target 
species are yellowfin, bigeye, albacore tuna, and swordfish. The total 
annual number of longline vessels in the western and central Pacific 
region has fluctuated between 3,000 and 6,000 for the last 30 years, 
including the 100 to 140 vessels in the Hawaii longline fisheries (NMFS 
2018).
Pelagic Fisheries
    International longline fisheries are characterized by inconsistent 
reporting and traditional gear configurations, including J-style hooks 
with squid bait, which result in higher interaction and mortality rates 
than for modified gear (Beverly and Chapman 2007; Lewison et al. 2004; 
Swimmer et al. 2017). For example, the Taiwan and China tuna longline 
fisheries are estimated to have bycatch rates several times higher than 
Hawaii longline fisheries (Bartram and Kaneko 2008; Chan and Pan 2012). 
Analyzing multi-national turtle bycatch data from 1990 to 2004, Molony 
(2005) found that the purse seine fishery and the deep, shallow, and 
albacore longline fisheries (operating between 15[deg] N and 31[deg] S) 
take an average of about 100 leatherback turtles annually. Lewison et 
al. (2004) collected fish catch data from 40 nations and turtle bycatch 
data from 13 international observer programs to estimate global 
longline bycatch of loggerhead and leatherback turtles in 2000. In the 
Pacific Ocean, they estimated 1,000 to 3,200 leatherback turtle 
(juvenile and adult) mortalities from pelagic longlining in 2000 
(Lewison et al. 2004). Using effort data from Lewison et al. (2004) and 
bycatch data from Molony (2005), Beverly and

[[Page 48396]]

Chapman (2007) estimated sea turtle longline bycatch to be 
approximately 20 percent of that estimated by Lewison et al. (2004), 
approximately 200 to 640 leatherback turtles annually. These estimates 
include turtles from the East and West Pacific DPS. While the results 
of each of these studies may be feasible, the Lewison et al. 2004 
estimates were based on available data at that time (i.e., less than 30 
percent of longline fishing effort) that was skewed toward fishing 
fleets with relatively better management and data reporting systems, 
and hence extrapolations may have overestimated interaction rates 
(Clarke et. al. 2014). However, Beverly and Chapman (2007) applied 
different catch per unit effort (CPUE) estimates in calculations 
differentiated between deep-set and shallow-set fisheries which have 
different interaction rates and, hence, their estimates may be more 
realistic.
    Despite scientific evidence showing that use of circle hooks and 
finfish bait significantly reduces leatherback turtle bycatch rates in 
longline fisheries (Gilman et al. 2007; Swimmer et al. 2017), nations 
are not required to use this hook/bait combination. The WCPFC Sea 
Turtle Conservation and Management Measure (CMM 2008-03) only applies 
to fleets using shallow-set gear targeting swordfish. Additionally, 
observer program coverage levels in WCPFC longline fisheries have not 
reached the required five percent coverage rate, resulting in limited 
bycatch reporting and likely underreporting (Clarke 2017). Further, 
existing sea turtle mitigation measures are currently only being 
applied to approximately one percent of shallow-set longline fisheries 
in the Convention Area, even though approximately 20 percent of the 
longline effort consists of shallow-sets (Clarke 2017).
    A workshop convened to assess the effectiveness of WCPFC's Sea 
Turtle Conservation and Management Measure found limited reductions in 
interactions and mortalities (Clarke 2017). Fishery observer data 
collected between 1989 and 2015 of 34 purse seine and longline fleets 
across the Pacific documented a total of 2,323 sea turtle interactions, 
of which 331 were leatherback turtles (Clarke 2017). Two bycatch 
hotspot areas were identified: One in central North Pacific (which 
likely reflects the 100 percent observer coverage in the Hawaii 
shallow-set longline fishery) and a second hotspot in eastern Australia 
(Clarke 2017). However, analysis of the data also found that overall 
conservation benefits would have been greater had mitigation measures 
also been applied to deep-set gear and not only to shallow-set 
swordfish fisheries (Clarke 2017).
    While bycatch in pelagic shallow-set swordfish-targeting longline 
fisheries has received the most attention to date, comparable studies 
for deep-set tuna-targeting fisheries are not available due to the more 
complex nature of these fisheries. There may be fewer interactions 
because deep-set fisheries (operating at depths more than 60 m) 
generally have lower bycatch rates, but they also have higher mortality 
rates than shallow-set gear (Lewison et al. 2004; Kaplan 2005; Gilman 
et al. 2007). Pelagic deep-set tuna-targeting fisheries cannot be 
ignored because they also have the potential to interact with 
leatherback turtles and constitute four times greater effort than 
shallow-set fisheries yet do not have RFMO gear mitigation requirements 
(Clarke 2017).
    Wallace et al. (2013), and a global review based on that study (FAO 
2014), categorized longline and gillnet fisheries interactions with 
West Pacific leatherback turtles as high risk but low impact for 
longline and gillnet gear, likely due to insufficient data from this 
data-poor region. Bycatch in small-scale coastal fisheries has been a 
significant contributor to population declines in many regions (Kaplan 
2005; Peckham et al. 2007; Alfaro-Shigueto et al. 2011), yet there is a 
significant lack of information from coastal and small-scale fisheries, 
especially from the Indian Ocean and Southeast Asian region (Lewison et 
al. 2014).
Southeast Asian Fisheries
    Waters of Southeast Asia are heavily fished by a variety of 
gillnets, trawls, fish traps, and a range of different hook and line 
gears, involving hundreds of thousands of fishers (FAO 2011). The West 
Pacific DPS nests, migrates, and forages throughout this densely 
populated and heavily exploited coastal region (Bellagio Sea Turtle 
Conservation Initiative 2008; Benson et al. 2011; Lewison et al. 2014; 
Roe et al. 2014; Harrison et al. 2018).
    There are few quantitative estimates of fisheries interactions near 
nesting beaches of this DPS, and existing reports provide only brief 
snapshots of impacts or are outdated. In Indonesia, between 1980 and 
1993, shark gillnets off the nesting beaches of Jamursba-Medi killed 
two to three nesting females weekly (Tapilatu et al. 2013). As a member 
of the WCPFC and the IOTC, Indonesia must comply with reporting 
requirements and conservation measures as required by these RFMOs. In 
2006, of the 85 sea turtle interactions observed in 539 sets on 10 tuna 
longline vessels, 3 were adult leatherback turtles (Zainudin et al. 
2007). Leatherback turtles are known to migrate through and forage 
within Philippine waters (Benson et al. 2011), and in 2014, aerial 
surveys observed leatherback turtles foraging in high density fishing 
areas (130 to 381 boats; MRF 2010, 2014). Leatherback turtles have also 
stranded dead or injured on Philippine beaches as a result of fishery 
interactions, typically with gillnet gear (Bagarinao 2011; Cruz 2006; 
MRF 2010; MWWP 2018 unpublished). In Malaysia, bycatch studies using an 
interview-based approach revealed that four leatherback turtles were 
caught in gillnets the prior year (Pilcher et al. 2008).
    Fisheries operating out of Australia and New Zealand may result in 
high bycatch and mortality rates for the winter nesting component of 
the DPS that migrates into the Southern Hemisphere (MacKay et al. 2014; 
Harrison et al. 2018). In Australia, some bycatch records exist for 
pelagic longline fisheries (Robins et al. 2002; Stobutzki et al. 2006), 
prawn trawls off Queensland and Northern Territory, gillnet fisheries 
off Queensland and Tasmania, and pot gear off Tasmania (Limpus 2009). 
Gillnet sea turtle bycatch is reported as widespread and includes 
anecdotal reports of leatherback turtles taken in Tasmanian tuna 
gillnet fisheries (Limpus 2009).
    Between 2004 and 2014, the Australian shallow-set fishery had an 
estimated 29 to 178 leatherback interactions, based on two to 10 
observations (average = 4.6 interactions) and four to 10 percent 
observer coverage (MacKay et al. 2014). These data are similar to 
bycatch information extrapolated from interviews with Australian 
fishers (Robins et al. 2002) which identified 162 leatherback turtles 
interactions in 2001 (MacKay et al. 2014). Australia has a sea turtle 
mitigation plan for its Eastern Tuna and Billfish Fishery which sets 
``trigger level'' interaction rates of <=0.0048 turtles per 1,000 hooks 
for each turtle species or 0.0172 turtles per 1,000 hooks overall (DAFF 
2009 in Clarke et al. 2014). In 2013, Australia reported that the 
trigger levels had been exceeded for the third year in a row and as a 
consequence the Australian Fisheries Management Authority required that 
shallow-set vessels in these fisheries use large circle hooks 
consistent with the WCPFC sea turtle measure (CMM 2008-03; Clarke et 
al. 2014).
    In New Zealand, records document 288 instances of stranding or 
commercial and recreational bycatch of leatherback turtles from 1892 to 
2015 (Godoy et al. 2016). New Zealand's surface longline fishery 
captured 90

[[Page 48397]]

leatherback turtles between 2008 and 2015 (Godoy et al. 2016). This is 
likely an underestimate because data were based on low observer 
coverage (5.8 percent overall), with limited observer overage during 
the peak time of leatherback abundance in New Zealand waters (January 
to March). Strandings can also provide opportunities for researchers to 
identify fisheries interactions. MacKay et al. (2014) identified 19 
mortalities in New Zealand and 29 mortalities in Australia. Although 
the cause of most strandings was often unknown, leatherback turtles 
have been found entangled in crab pot gear and monofilament fishing 
nets and ropes. Longline fishing is concentrated off southern 
Queensland and New South Wales, Australia and is the suspected cause of 
41 percent of strandings (n = 12). In Victoria, Tasmania and South 
Australia, 61 percent of strandings (n = 17) involved suspected 
entanglement in inshore fishing gear and crab pots (MacKay et al. 
2014).
U.S. Pacific Pelagic Fisheries
    Detailed bycatch data are available for U.S.-managed pelagic 
fisheries operating in the central and eastern Pacific Ocean due to 
regulatory mandates and high levels of observer coverage. Longline 
fisheries, based out of Hawaii and American Samoa, may interact with 
foraging turtles of the West Pacific DPS. However, only two 
interactions involved individuals of the East Pacific DPS in 1995 and 
2011 (P. Dutton, NMFS, pers. comm., 2018). Prior to 2001, the Hawaii 
longline fishery was estimated to capture about 110 leatherback turtles 
annually, resulting in approximately 9 annual mortalities (McCracken 
2000). Since 2005, the fishery has reduced its estimated mortality to 
seven leatherback turtles annually, and data confidence increased 
significantly due to increased observer coverage (NMFS 2018). The 
fishery was closed in 2001 under court order and re-opened in 2004 as 
two separate fisheries: A shallow-set swordfish-targeting fishery and a 
deep-set tuna-targeting fishery. Management requirements include: Gear 
modification (e.g., circle hooks and fin-fish bait) and handling 
measures designed to reduce sea turtle bycatch rates and post-hooking 
mortality in both fisheries; an annual hard-cap limit on the number of 
allowable interactions in the shallow-set fishery; 100 percent observer 
coverage in the shallow-set fishery; and 20 percent observer coverage 
in the deep-set fishery (50 CFR 665 (Subparts A-C); NMFS 2012, 2014, 
2015). The shallow-set fishery has been closed three additional times 
since reopening in 2004: In 2006, after reaching the hard cap for 
loggerhead turtle interactions (n = 17); in 2011, after reaching the 
hard cap for leatherback turtle interactions (n = 16); and in 2018 
under a stipulated settlement after the Ninth Circuit Court of Appeals 
held that NMFS' no jeopardy determination for loggerheads in the 2012 
biological opinion (9th Circuit 2017) was arbitrary and capricious. See 
Turtle Island Restoration Network v. U.S. Dep't. of Commerce, 878 F.3d 
725 (9th Cir. 2017). Since 2004, leatherback turtle interactions in the 
shallow-set component of the fishery have been reduced by 84 percent 
from 0.03 to 0.01 BPUE as a result fisheries regulations (Swimmer et 
al. 2017). Between 2004 and 2017, there have been 99 total leatherback 
turtle interactions in the shallow-set fishery (or approximately 8 
turtles annually), based on 100 percent observer coverage (WPRFMC 
2018). Between 2002 and 2016, an estimated 168 interactions may have 
occurred in the Hawaii deep-set fishery (or approximately 12 annually), 
based on an extrapolation of data collected at a level of 20 percent 
observer coverage (WPRFMC 2018). Observer coverage of the American 
Samoa longline fishery has varied over time from 5 to 40 percent and 
has had an estimated 59 interactions between 2010 and 2017 (WPRFMC 
2018).
    The U.S. tuna purse seine fishery operating in the Western and 
Central Pacific Ocean anticipates up to 11 leatherback turtle 
interactions annually (NMFS 2006). However, the fishery had fewer 
interactions, with approximately 16 leatherback turtle interactions 
between 2008 and 2015 based on observer coverage ranging from 20 to 100 
percent (NMFS unpublished data).
    From 1990 to 2009, there were 24 observed leatherback turtle 
interactions in the California drift gillnet fishery based on 15.6 
percent per year observer coverage (Martin et al. 2015). Genetic 
analyses indicated that almost all originated from the West Pacific DPS 
(Dutton et al. 1999; NMFS SWFSC unpublished). In 2001, NMFS implemented 
regulations (i.e., a large time/area closure in Central California) 
that reduced interactions by approximately 80 to 90 percent, with only 
two leatherback turtle interactions (both alive) observed based on 20 
to 30 percent observer coverage since regulations were implemented 
(NMFS West Coast Region unpublished). Drift gillnet fishing is 
prohibited annually from August 15 to November 15 within the California 
leatherback turtle conservation area. Currently, NMFS anticipates up to 
10 interactions (or 7 mortalities) over a 5-year period (NMFS 2013).
    In addition, nine fixed gear fisheries operate off the U.S. West 
Coast, including the Federally-managed sablefish pot fishery and the 
state-managed California Dungeness crab fishery. Since 2008, only one 
leatherback interaction has been documented in the sablefish fishery 
(NMFS 2013). The state-managed Dungeness crab fishery may be a newly 
emerging threat: Two documented leatherback entanglements in pot gear 
(mainline or surface buoy) occurred in 2015 and 2016. Fishing effort 
was high, and the fishery had shifted into the Central California 
region, which overlaps somewhat with leatherback foraging habitat (S. 
Benson, NMFS, pers. comm., 2018). In 2019, the State of California 
settled with a non-profit organization in response to a complaint that 
the commercial Dungeness crab fishery was taking leatherback sea 
turtles (and other large whales) without authorization under section 10 
of the ESA. The California Dungeness crab fishery closed in mid-April 
2019 as part of the settlement agreement and again on May 15, 2020 
(just the Central Management Area), due to significant risk of marine 
life entanglement. The northern part of California remains open until 
mid-July unless CDFW decides to take further management action (i.e., 
if risks to large whales and/or leatherbacks is elevated in that area).
East Pacific Pelagic Fisheries
    The West Pacific DPS has a vast trans-Pacific range. Some 
individuals forage in the East Pacific Ocean, where leatherback turtles 
are caught in fisheries of Peru and Chile (Donoso and Dutton 2010; 
Alfaro-Shigueto et al. 2007, 2011, 2018). Of 59 leatherback turtles 
caught in East Pacific fisheries, an estimated 15 percent of 
individuals sampled originated from the West Pacific DPS (Dutton et al. 
2000; Donoso and Dutton 2010). Information compiled by IATTC on sea 
turtle interactions with pelagic longline fisheries operating in the 
East Pacific is limited, given that requirements for longline observer 
coverage of five percent was only implemented in January 2013 (Clarke 
et al. 2014). Additional information on East Pacific fisheries are 
presented in the bycatch section for the East Pacific DPS.
Summary of Fisheries Bycatch
    We conclude that individuals of this DPS are exposed to high 
fishing effort throughout their foraging range, in coastal waters near 
nesting beaches, and when migrating to and from nesting

[[Page 48398]]

beaches, though very little fisheries data are available for coastal 
areas. Bycatch rates in international pelagic and coastal fisheries are 
high, and these fisheries have limited management regulations despite 
hotspots of high interactions in Southeast Asia (Lewison et al. 2004, 
2014; Alfaro-Shigueto et al. 2011; Wallace et al. 2013; Clarke 2017). 
Annual interaction and mortality estimates are only available for U.S.-
managed pelagic fisheries, which operate under extensive fisheries 
regulations that are designed to minimize the capture and mortality of 
endangered and threatened sea turtles (NMFS 2013; Swimmer et al. 2017; 
NMFS 2018). Mortality reduces abundance, by removing individuals from 
the population; it also reduces productivity, when nesting females are 
killed. We conclude that fisheries bycatch is a major threat to the 
West Pacific DPS.

Vessel Strikes

    Vessel strikes are a threat to the West Pacific DPS. Between 1981 
and 2016, there were 11 documented vessel strikes in central California 
(NMFS West Coast Region, unpublished data 2018). Many vessel strikes 
are not reported, and turtles are not recovered.
    The range of the DPS overlaps with many high-density vessel traffic 
areas. Though the potential for exposure is high, we are only aware of 
11 vessel strikes in recent decades. Vessel strikes resulting in 
mortality would lower the abundance of the DPS. However, available data 
does not support characterizing this as a high or moderate impact. We 
conclude that vessel strikes pose a threat to the DPS, albeit of less 
concern than other impacts such as overutilization and fisheries 
interactions.

Pollution

    Pollution includes contaminants, marine debris, and ghost fishing 
gear. Leatherback turtles can ingest small debris, causing internal 
damage and blockage. Larger debris can entangle animals, leading to 
reduced mobility, starvation, and death. Given the amount of floating 
debris in the Pacific Ocean (Lebreton et al. 2018), marine debris has 
the potential to be a significant threat to the DPS. Presently 
available data do not allow for quantifying the precise extent of the 
threat.
    Leatherback turtles feed exclusively on jellyfish and other 
gelatinous organisms and as a result may be prone to ingesting plastics 
resembling their food source (Schuyler et al. 2013). Lebreton et al. 
(2018) estimated plastic debris accumulation to be at least 79,000 
(45,000 to 129,000) tonnes in the Great Pacific Garbage Patch, a 1.6 
million km\2\ of subtropical waters between California and Hawaii. This 
figure is four to 16 times greater than previously reported. 
Entanglement in ghost fishing gear is also a concern (Gilman et al. 
2016), and derelict nets made up approximately 46 percent by piece, and 
86 percent by weight, of debris floating in this area (Lebreton et al. 
2018). The highest risk areas within the range of the West Pacific DPS 
where animals may encounter significant amounts of debris includes the 
north Pacific gyre, the South China Sea, and off of the east coast of 
Australia (Schuler et al. 2015). However, Wedemeyer-Strombel et al. 
(2015) found no plastics in the gastrointestinal tracts of two 
leatherback carcasses from American Samoan and Hawaiian longline 
fisheries from 1993 to 2011. Clukey et al. (2017) found no plastics in 
the gastrointestinal tracts of three leatherback carcasses from Pacific 
longline fisheries captured between 2012 and 2016. However, it is very 
difficult to obtain dead leatherback turtles to study these effects, 
and given the great amount of plastics within environment, such results 
may underestimate ingestion impacts.
    Few studies of pollutants and their effect on leatherback turtles 
were available within the range of this DPS. Harris et al. (2011) found 
the heavy metal exposure in leatherback turtles foraging off the coast 
of California to be nine times higher than the St. Croix nesting 
population, although levels were not expected to be lethal. We do not 
know if there were sub-lethal effects. Stewart et al. (2011) found that 
PCBs are more likely to be transferred from females to their eggs than 
from the environment to eggs.
    Given the large amount of marine debris within the range of the 
DPS, we expect exposure to be high for all life stages despite low 
sample sizes of leatherback turtles with ingested marine debris. 
Potential impacts include death and injury. However, quantitative 
estimates of such impacts are not available. We conclude that pollution 
may be a threat to the DPS.

Natural Disasters

    The best available scientific and commercial data indicate that 
natural disasters are a threat to the DPS but do not allow the impact 
to be quantified. Natural disasters within the range of this DPS 
include: Tsunamis, typhoons, earthquakes, and flash floods. Such 
environmental events are periodic, with localized impacts that do not 
persist over time. These events may reduce nest incubation and hatching 
success in one season or at few locations. While leatherback turtles 
have undoubtedly evolved to sustain such natural impacts, the 
increasing frequency of environmental events as a result of a changing 
climate, which can affect the frequency and intensity of high tides and 
large storms, may hamper productivity and conservation activities (Goby 
et al. 2010; S. Benson, NMFS, pers. comm., 2018). Such events may pose 
additional threats by depositing marine debris on nesting beaches and 
in occupied waters. The 2011 Japan tsunami and the 2006 Indonesian 
earthquake and resulting tsunami likely deposited large amounts of 
debris (i.e., millions of tons) into the foraging and migrating 
habitats of the DPS (Hafner et al. 2014; NOAA 2015). We conclude that 
natural disasters pose a potential threat to the West Pacific DPS.

Climate Change

    Climate change is a threat to the West Pacific DPS. A warming 
climate and rising sea levels can impact leatherback turtles through 
changes in beach morphology, increased sand temperatures leading to a 
greater incidence of lethal incubation temperatures, changes in 
hatchling sex ratios, and the loss of nests or nesting habitat due to 
beach erosion (Benson et al. 2015).
    Elevated egg incubation temperatures can lead to mortality. During 
the 2009/2010 nesting season at the Huon Coast (Papua New Guinea), 
Pilcher (2010) found higher incubation temperatures (32 to 33 [deg]C) 
in exposed nests compared to shaded nests (29 to 30 [deg]C). Sea 
turtles exhibit temperature-dependent sex determination. The incubation 
temperature determines sex ratios and the duration of incubation (i.e., 
thermosensitive period). Along the Huon Coast, incubation duration 
decreased during the nesting season as beach temperatures warmed. 
During the 2006/2007 nesting season, nests laid in November hatched in 
61.8  4.2 days, and nests laid in February hatched in 55.8 
 3.4 days (n = 171 nests; Steckenreuter et al. 2010). 
Assuming that hatchlings were male at temperatures less than 29.2 
[deg]C and female at temperatures greater than 30.5 [deg]C, 
Steckenreuter et al. (2010) estimated that only 7.7 percent of the 
hatchlings were female, indicating a highly male-skewed sex ratio. 
However, given the Pilcher (2010) results, sex ratios are likely 
variable over time and space.
    Climatic change may also alter rainfall levels, which may cool 
beaches and offset increases in sand temperature. At Wermon, the sand 
is black, yet beach temperatures are lower, perhaps because

[[Page 48399]]

peak nesting coincides with the monsoon season (Tapilatu and Tiwari 
2007). Sand temperatures fluctuate between 28.6 and 34.9 [deg]C at 
Jamursba-Medi and between 27.0 and 32.7 [deg]C at Wermon (Tapilatu and 
Tiwari 2007). Hatching success of nests undisturbed by feral pig 
predation was significantly lower in Jamursba-Medi (25.5 percent) than 
Wermon (47.1 percent). Although there was significant variation between 
beaches, Tapilatu and Tiwari (2007) concluded that high sand 
temperatures may exceed the thermal tolerance of leatherback embryos, 
resulting in high embryo mortality and low hatching success at 
Jamursba-Medi. Further, Tapilatu and Tiwari (2007) concluded that high 
average sand temperatures may suggest a female-biased population at 
Jamursba-Medi. However, the mean incubation period of 61.5  
4.7 days (Tapilatu and Tiwari 2007) was similar to the length of 
incubation recorded in Papua New Guinea during the cooler November 
period, which Steckenreuter et al. (2010) suggested produced a male-
biased sex ratio.
    Tapilatu et al. (2013b) found that the daily average sand 
temperatures during the boreal summer (from 2005 to 2012) ranged from 
26.5 to 34.9 [deg]C, suggesting the production of female-biased sex 
ratios and potentially lower hatching success. Further, histological 
examination of dead hatchlings from both summer and winter nesting 
seasons from 2009 to 2019 produced a female-biased sex ratio, which is 
consistent with the relatively warm thermal profiles of the nesting 
beaches (Tapilatu et al. 2013b). Additional impacts of climate change 
include increased sea level rise and storm frequency, resulting in 
greater nest inundation and beach erosion. As sea level rises, King 
Tides are likely to have a greater effect on nests. Climate change may 
also affect prey availability. Saba et al. (2007, 2012) identified a 
correlation between the reproductive frequency of the East Pacific DPS 
and ENSO events. Because the West DPS also forages in the East Pacific 
Ocean, it too may be exposed to variability in productivity.
    The threat of climate change is likely to modify the nesting and 
foraging conditions for turtles of the DPS. Impacts are likely to 
affect productivity. Negative impacts and low hatching success due to 
high beach temperatures and coastal erosion have already been 
documented and are likely to become worse, and thus we conclude that 
climate change is a threat to the West Pacific DPS.

Conservation Efforts

    There are numerous efforts to conserve the leatherback turtle. The 
following conservation efforts apply to turtles of the West Pacific DPS 
(for a description of each effort, please see the section on 
conservation efforts for the overall species): Convention on the 
Conservation of Migratory Species of Wild Animals, Convention on 
Biological Diversity, Convention on International Trade in Endangered 
Species of Wild Fauna and Flora, Convention for the Protection of the 
Marine Environment and Coastal Area of the South-East Pacific (Lima 
Convention), Convention for the Conservation and Management of Highly 
Migratory Fish Stocks in the Western and Central Pacific Ocean (WCPF 
Convention), Convention for the Protection of the Natural Resources and 
Environment of the South Pacific Region, Convention Concerning the 
Protection of the World Cultural and Natural Heritage (World Heritage 
Convention), Eastern Pacific Leatherback Network, Eastern Tropical 
Pacific Marine Corridor Initiative, FAO Technical Consultation on Sea 
Turtle-Fishery Interactions, IAC, MARPOL, IUCN, The Memorandum of 
Understanding of a Tri-National Partnership between the Government of 
the Republic of Indonesia, the Independent State of Papua New Guinea 
and the Government of Solomon Islands, Ramsar Convention on Wetlands, 
RFMOs, Secretariat of the Pacific Regional Environment Programme, 
UNCLOS, and UN Resolution 44/225 on Large-Scale Pelagic Driftnet 
Fishing. Although numerous conservation efforts apply to the turtles of 
this DPS, they do not adequately reduce its risk of this DPS, they do 
not adequately reduce its risk of extinction.

Extinction Risk Analysis

    After reviewing the best available information, the Team concluded 
that the West Pacific DPS is at high risk of extinction. The DPS 
exhibits a total index of nesting female abundance of 1,277 females at 
two currently monitored beaches over the most recent remigration 
interval. These beaches may represent 75 percent of total DPS nesting 
activity. This abundance makes the DPS vulnerable to stochastic or 
catastrophic events that increase its extinction risk. This DPS 
exhibits low hatching success and decreasing nest and population trends 
due to past and current threats, which are likely to further lower 
abundance and increase the risk of extinction. The DPS exhibits genetic 
diversity and metapopulation structure, with nesting aggregations 
distributed throughout four nations. Nesting occurs during two seasons 
(winter and summer), with year-round nesting at some locations and uses 
multiple foraging areas, throughout the Pacific Ocean. Thus, the DPS 
has some resilience to stochastic events and environmental 
perturbations at nesting beaches and foraging areas. However, its 
abundance and declining trends place the DPS at risk of extinction as a 
result of past threats.
    Current threats also contribute to the risk of extinction of this 
DPS. The overutilization of turtles and eggs, as a result of legal and 
illegal harvest, is the primary threat to this DPS, reducing abundance 
and productivity. Abundance and productivity are further reduced by 
fisheries bycatch. Juvenile and adult turtles are taken by numerous, 
international, coastal, and pelagic fisheries throughout the extensive, 
pan-Pacific foraging range of the DPS. Predation (especially by dogs 
and pigs) reduces productivity at high rates. Erosion and inundation 
result in habitat loss and modification that reduces productivity and 
contributes to low hatching success. Additional threats include: 
Pollution, vessel strikes, and natural disasters. Climate change is an 
increasing threat that results in reduced productivity. Though many 
regulatory mechanisms exist, they do not adequately reduce threats.
    We conclude, consistent with the team's findings, that the West 
Pacific DPS is at risk of extinction. Its nesting female abundance 
makes the DPS highly vulnerable to threats. The declining nesting trend 
further contributes to its risk of extinction. While the DPS has 
spatial structure and diversity, the resilience provided by those 
factors is likely to be eroded by the reduced and declining abundance. 
Past egg and turtle harvest reduced the abundance and productivity of 
this DPS and remains a primary threat. Fisheries bycatch is also a 
primary threat that reduces abundance by removing mature and immature 
individuals from the population. Predation is also a major threat to 
productivity. Though numerous conservation efforts apply to this DPS, 
they do not adequately reduce the risk of extinction. We conclude that 
the West Pacific DPS is in danger of extinction throughout its range 
and therefore meets the definition of an endangered species. The 
threatened species definition does not apply because the DPS is 
currently in danger of extinction (i.e., at present), rather than on a 
trajectory to become so within the foreseeable future.

[[Page 48400]]

East Pacific DPS

    The Team defined the East Pacific DPS as leatherback turtles 
originating from the East Pacific Ocean, north of 47[deg] S, south of 
32.531[deg] N, east of 117.124[deg] W, and west of the Americas. In the 
south, the cold waters of the Antarctic Circumpolar Current likely 
restrict the nesting range of this DPS. We placed the northern and 
western boundaries at the border between the United States and Mexico 
because this DPS forages primarily in the East Pacific Ocean, off the 
coasts of Central and South America.
    The range of the DPS (i.e., all documented areas of occurrence) is 
centered in the eastern Pacific Ocean but may include distant waters 
for foraging, as demonstrated by a turtle satellite-tracked to waters 
off the Tonga Trench and a turtle captured by the Hawaii longline 
fishery, genetically assigned to the population we refer to in this 
finding as the East Pacific DPS (P. Dutton, NMFS, pers. comm., 2018). 
Records indicate that the DPS occurs in the waters of the following 
nations: Chile; Colombia; Costa Rica; Ecuador; El Salvador; France 
(Clipperton Island); Guatemala, Honduras; Mexico; Nicaragua; Panama; 
Peru; and the United States (Hawaiian Islands) (Wallace et al. 2013).
    Leatherback turtles of the East Pacific DPS nest primarily on 
beaches in Mexico, Costa Rica, and Nicaragua. In Mexico, where the 
largest nesting aggregations occur, nesting beaches are found in 11 
states, over 7,828 kilometers as far north as Baja California Sur 
(Sarti 2002). The following beaches in Mexico host approximately 40 to 
50 percent of total nesting for the nation: Mexiquillo 
(Michoac[aacute]n), Tierra Colorada (Guerrero), and Cahuit[aacute]n, 
Chacahua, and Barra de la Cruz (Oaxaca; Gaona Pineda and 
Barrag[aacute]n Rocha 2016). In Costa Rica, approximately 75 percent of 
nesting occurs within the Parque Nacional Marino Las Baulas (Guanacaste 
Province) at three nesting beaches: Playa Ventanas; Playa Grande; and 
Playa Langosta (based on recent abundance estimates from 2011-2015; 
Santidri[aacute]n Tomillo et al. 2017). In Nicaragua, small numbers of 
leatherback turtles nest on Playa Salamina-Costa Grande and Veracruz de 
Acayo (Chacocente Wildlife Refuge) (FFI 2018). Rare nesting events have 
been documented in Guatemala (n = 6), El Salvador (n = 4), and Panama 
(n = 4), with none in Honduras (Sarti et al. 1999).
    Generally, the nesting season starts in October and ends in March 
(Santidri[aacute]n Tomillo et al. 2007; Eckert et al. 2012). Nesting is 
generally bound between 10[deg] N and 20[deg] N, falling within the 
northeast corner of the Intertropical Convergence Zone. The nesting 
beaches share similarly warm temperatures, moderate annual rainfall, 
and seasonal dynamics (Saba et al. 2012). In general, nesting beach 
habitat for leatherback turtles is associated with deep water and 
strong waves and oceanic currents, but shallow water with mud banks are 
also used by leatherback turtles. Beaches with coarse-grained sand and 
free of rocks, coral, or other abrasive substrates also appear to be 
selected by leatherback turtles (reviewed by Eckert et al. 2012).
    Foraging areas of the East Pacific DPS include coastal and pelagic 
waters of the southeastern Pacific Ocean. Leatherback turtles are 
widely dispersed on the high seas throughout the eastern Pacific Ocean 
(Shillinger et al. 2008). They also forage in coastal areas off the 
coast of Peru and Chile (Alfaro-Shigueto et al. 2007; Eckert 1997; 
Donoso and Dutton 2010). Using satellite telemetry, Morreale et al. 
(1996) tracked the movements of eight post-nesting females and 
identified a persistent southbound migration corridor from Las Baulas 
National Park toward the Galapagos Islands. Eckert (1997) found a 
similar pattern, tracking seven post-nesting females from Mexiquillo in 
a similar direction; while three continued to the same foraging habitat 
as the Costa Rican nesting females, four shifted their movements away 
from the South American coast, when a strong El Ni[ntilde]o caused a 
warm water anomaly. Additional tracking of 46 post-nesting females from 
Las Baulas National Park over a 3-year period (2004/2005 to 2006/2007) 
confirmed the persistent migratory corridor (Shillinger et al. 2008). 
The turtles navigated the equatorial current system, south to around 
5[deg] S latitude and negotiated the strong alternating eastward-
westward flows of the equatorial current, swimming predominantly in a 
southward direction and moving rapidly through the productive 
equatorial region. They then dispersed throughout the South Pacific 
Gyre ecosystem, which is characterized by low phytoplanktonic biomass. 
The South Pacific Gyre contains ample mesoplankton forage base, as 
demonstrated by tuna longline fisheries effort in the eastern tropical 
Pacific Ocean (Shillinger et al. 2008). Of the 46 turtles, only one 
leatherback moved into coastal foraging areas, which had been 
documented earlier by Eckert (1997). During the course of the tracking 
duration, this female occupied nearshore foraging habitats along the 
coast of Central America, which represents highly productive areas when 
compared with oceanic areas. Researchers have hypothesized that high 
bycatch along the coastal areas of Central and South America could have 
extirpated a coastal migratory phenotype in this population (Saba et 
al. 2007). Recently, Harrison et al. (2018) determined that post-
nesting females from Las Baulas National Park spent 78.2 percent of 
their time on the high seas, 17.8 percent of their time in Costa Rica's 
EEZ, and 3.7 percent of their time around the Galapagos Islands.
    In summary, preferred foraging areas for the East Pacific DPS are 
characterized by low sea surface temperatures and high mesoscale 
variability. Post-nesting females migrate relatively quickly through 
areas that contain the strong equatorial currents as well as high 
chlorophyll-a concentrations, likely because of the strong currents. 
While swimming speed was significantly higher in areas of high 
chlorophyll levels, the association between these two variables was 
weak (Shillinger et al. 2008). Once past this area, they appear to 
forage in the southern part of their range in the South Pacific 
Subtropical Convergence, where there is a sharp gradient in primary 
production. In this area, Ekman upwelling may accelerate the transport 
of nutrients and consequently increase prey availability. Seasonally, 
leatherback turtles from the East Pacific DPS foraged at higher 
southerly latitudes during the austral summer (November to February), 
which may reflect seasonal patterns in prey abundance during higher 
latitudes (Bailey et al. 2012).

Abundance

    The total index of nesting female abundance for the East Pacific 
DPS is 755 females. We based this total index on 13 nesting 
aggregations in: Mexico (Mexican Commission for Natural Protected 
Areas; L. Sarti, CONANP, pers. comm. 2018); Costa Rica 
(Santidri[aacute]n Tomillo et al. 2017; Leatherback Trust 2018); and 
Nicaragua (FFI 2018). Our total index does not include several 
unquantified nesting aggregations in Mexico, Costa Rica, and Nicaragua. 
To calculate the index of nesting female abundance for nesting beaches 
in Mexico (i.e., 572 females), we added the total number of nesting 
females between the 2013/2014 and 2016/2017 nesting seasons (i.e., a 4-
year remigration interval; L. Sarti, CONANP, pers. comm., 2018) at each 
beach. We performed a similar calculation for Costa Rica (n = 165 
females). To

[[Page 48401]]

calculate the index of nesting female abundance in Nicaragua (i.e., 20 
females), we divided the total number of nests between the 2014/2015 
and 2017/2018 nesting seasons (i.e., a 4-year remigration interval; 
Santradi[aacute]n Tomillo et al. 2007) by the clutch frequency (7.2 
clutches/season; Santradi[aacute]n Tomillo et al. 2007).
    This number represents an index of nesting females for this DPS 
because it only includes available data from recently and consistently 
monitored nesting beaches. While rare or sporadic nesting may occur on 
other beaches, consistent and standardized monitoring only occurs at 
these beaches, which are for the most part protected.
    Our total index of nesting female abundance is similar to published 
abundance estimates for this DPS. The IUCN Red List assessment 
estimated the total number of mature individuals (males and females) at 
633 turtles, based first on dividing the average annual number of nests 
(n = 926) by the estimated clutch frequency (n = 7.2, Reina et al. 
2002) to obtain an average annual number of nesting females. This value 
was then multiplied by the average remigration interval (n = 3.7 years, 
Reina et al. 2002; Santidri[aacute]n Tomillo et al. 2007) to obtain a 
total number of adult females that included nesting as well as non-
nesting turtles. In order to account for adult males, the authors 
assumed that the sex ratio of hatchlings produced on nesting beaches in 
the East Pacific (approximately 75 percent female, or 3:1 female:male 
ratio) reflected the natural adult sex ratio (Wallace et al. 2013). A 
more recent analysis of primary sex ratios that included multiple years 
of data and considered hatching success (i.e., lower in hot nests) 
estimated primary sex ratios at Playa Grande, Costa Rica as 
approximately 85 percent female (Santidri[aacute]n Tomillo et al. 
2014). In Mexico, the female to male ratio is closer to 1.1:1 (A. 
Barragan, Kutzari, pers. comm., 2019).
    In Mexico, the beaches included in our total index represent 
approximately 70 to 75 percent of total nesting in that nation (Gaona 
Pineda and Barragan Rocha 2016). However, our total index does not 
include nesting females from Agua Blanca (40 km in Baja California); 
Playa Ventura (6 km), Playa San Valent[iacute]n (21 km), Piedra de 
Tlacoyunque (44 km in Guerrero), and La Tuza (16 km in Oaxaca) (Sarti 
et al. 2007). These beaches are not regularly monitored for nesting, 
which is thought to be rare or of low abundance (L. Sarti, CONANP, 
pers. comm., 2018).
    In Costa Rica, 75 percent of nesting occurred at Las Baulas 
National Park (summarized in Santidri[aacute]n Tomillo et al. 2017), 
although the recent nesting at other beaches may lower this percentage. 
These beaches include: Naranjo, Cabuyal, Nombre de Jes[uacute]s, 
Ostional, and Caletas. The longest data set was provided for Naranjo, 
which has been intermittently covered from 1971 to 2015. Limited 
nesting has been documented at Playa Coyote and at Playa Caletas, which 
is a high energy eight kilometer beach located on the Nicoya Peninsula 
(Squires 1999). Given the lack of nesting events for Caletas in recent 
years, it may no longer host leatherback nesting, despite the fact that 
the Playa Caletas/Ario National Wildlife Refuge was created in 2004 to 
protect leatherback turtles (Gaos et al. 2008).
    In Nicaragua, leatherback turtles nest at three beaches. Salamina 
Costa Grande and Veracruz de Acayo (in the Rio Escalante Chacocente 
Wildlife Refuge) host the most nesting and have been subject to the 
most consistent monitoring. Small numbers of females also nest at Juan 
Venado National Reserve, which is not consistently monitored (V. Gadea, 
FFI, personal communication, 2018).
    Nesting is rare in other nations (Sarti et al. 1999). Nesting is 
very uncommon in Ecuador with one record of a female attempting to nest 
(according to local reports) in Atacames, a province of Esmeraldas 
(Salas 1981). Sarti et al. (1999) reported six nests at Playa Puntilla, 
El Salvador, but overall nesting is low and/or unknown throughout the 
nation. In Guatemala, nesting is rare, with reports by Sarti et al. 
(1999) recording only eight nests during an entire season, and more 
recently, zero to six nests per year along the Pacific coast of 
Guatemala (Muccio and Flores 2015). Past nesting sites included Hawai 
beach, La Candelaria, Taxico, Santa Rosa, and the zone adjacent to the 
border with El Salvador, as reported by Chac[oacute]n-Chaverri (2004). 
Although nesting has been documented at Barqueta National Refuge, 
little is known about nesting in Panama (Chac[oacute]n-Chaverri 2004).
    Our total index of nesting female abundance (755 females) places 
the DPS at risk for environmental variation, genetic complications, 
demographic stochasticity, negative ecological feedback, and 
catastrophes (McElhany et al. 2000; NMFS 2017). These processes, 
working alone or in concert, place small populations at a greater 
extinction risk than large populations, which are better able to absorb 
losses in individuals. Due to its small size, the DPS has relatively 
little capacity to buffer such losses. Historical abundance estimates 
were much greater (e.g., 75,000 leatherback nesting females estimated 
in Pacific Mexico from a 1980 aerial survey ((Pritchard 1982). However, 
this estimate was derived from a brief aerial survey and may have been 
an overestimate (Pritchard 1996)), indicating that this population at 
one time had the capacity for a much larger nesting population. 
Therefore, the current nesting female abundance is likely an indicator 
of past and current threats, and given the intrinsic problems of small 
population size, elevates the extinction risk of this DPS.

Productivity

    The East Pacific DPS exhibits a decreasing nest trend since 
monitoring began, with a 97.4 percent decline since the 1980s or 1990s, 
depending on the nesting beach (Wallace et al. 2013). Despite intense 
conservation efforts, the decline in nesting had not been reversed as 
of 2011 (Benson et al. 2015). We found a declining nest trend at some 
of the remaining, small nesting aggregations. Abundance at Las Baulas, 
Costa Rica (previously the single largest nesting aggregation) at its 
peak was seven times the current abundance at Playa Barra de la Cruz/
Playa Grande, Mexico (currently the largest nesting aggregation). From 
1988/1989 to 2015/2016, the number of nesting females at Las Baulas 
declined -15.5 percent annually (sd = 3.8 percent; 95 percent CI = -
23.1 to -7.8 percent; f = 0.998; mean annual nests = 315).
    In recent decades (after a historical decline), nest counts have 
increased at some beaches in Mexico. The Playa Tierra Colorada nest 
trend has increased by 0.6 percent annually (sd = 8.9 percent; 95 
percent CI = -17.1 to 18.9 percent; f = 0.536; mean annual nests = 153) 
between the 1996/1997 and 2016/2017 nesting seasons. Over the same time 
period, nesting at Playa Barra de la Cruz/Playa Grande increased by 9.5 
percent annually (sd = 8.0 percent; 95 percent CI = -6.5 to 25.8 
percent; f = 0.918; mean annual nests = 122). In contrast, nest counts 
at Cahuit[aacute]n decreased from 1997/1998 through 2016/2017, with a 
median trend of -4.3 percent annually (sd = 9.7 percent; 95 percent CI 
= -22.1 to 17.6 percent; f = 0.716; mean annual nests = 123).
    We lack adequate data on nesting in Nicaragua to estimate trends.
    Our trend analysis yields similar results to other published 
findings. The IUCN Red List assessment concluded that this 
subpopulation is decreasing and has declined by -97.4 percent over the 
past three generations (Wallace et al. 2013). The number of nests at 
Mexico nesting beaches has declined precipitously in recent decades 
(Benson et al. 2013). Historically, Mexico hosted

[[Page 48402]]

the largest leatherback turtle nesting aggregation in the world, with 
75,000 nesting females estimated during an aerial survey in 1980 
((Pritchard 1982). However, this estimate was derived from a brief 
aerial survey and may have been an overestimate (Pritchard 1996)). 
Prior to that aerial survey, Marquez et al. (1981) reported that the 
nesting beach of San Juan Chacahua (Oaxaca) was the most important 
nesting site in Mexico, with approximately 2,000 females nesting each 
season. Researchers also identified Tierra Colorada and Mexiquillo as 
important nesting sites, with approximately 3,000 to 5,000 nests per 
season. Monitoring of the nesting assemblage at Mexiquillo has been 
continuous since 1982. During the mid-1980s, more than 5,000 nests per 
season were documented along 4 km of this nesting beach. By 1993, less 
than 100 nests were counted along the entire 18 km beach (Sarti 2002). 
According to Sarti et al. (1996), nesting declined at this location at 
an annual rate of over 22 percent from 1984 to 1995. Researchers from 
the National University of Mexico recorded 3,000 to 5,000 nests 
annually from 1982 to 1989 at primary nesting beaches, with sharp 
declines observed in 1993 to 1994 at the nesting sites at Mexiquillo, 
Tierra Colorada, Chacahua and Barra de la Cruz. These early reports 
were generally snapshots (e.g., local unpublished data) of leatherback 
nesting activity in Mexico, until 1995, when a more coordinated 
conservation effort took shape in the form of complete nesting surveys 
for the entire Pacific coast of Mexico (Eckert 1997). In 1995, 
``Proyecto Laud'' (Leatherback Project) was formed to estimate the 
population size using comprehensive surveys. In 1995 and 1996, Proyecto 
Laud estimated approximately 1,100 females nesting throughout Mexico; 
the next two seasons, they estimated between 236 and 250 nesting 
females, and declines continued. Currently, based on data from 2014 
through 2018 (preliminary) between 100 and 250 females nest at all the 
protected beaches in Mexico.
    In Costa Rica, the number of nesting females per season declined 
from 1,367 females in 1988 to 117 females in 1998 (Spotila 2000). While 
there were increases in the number of nesting females during the 1999/
2000 season (224 females) and 2000/2001 season (397 females), the 
population has shown a steady decline, with less than 30 nesting 
females in recent years (i.e., through 2016; The Leatherback Trust 
2018).
    In Nicaragua, 108 leatherback turtles nested on Playa Chacocente 
from October to December, 1980; in January 1981, 100 turtles nested in 
a single night on Playa El Mogote (Arauz 2002). An aerial survey of 
Playa El Mogote during the 1998/1999 nesting season revealed a nesting 
density of 0.72 turtles per kilometer (Sarti et al. 1999 in Arauz 
2002). During the 2000/2001 nesting season, community members near 
Playa El Mogote reported that 210 leatherback nests had been deposited. 
That number decreased to 29 nests during the 2001/2002 nesting season 
(Arauz 2002). At Playa Veracruz 48 nesting females were identified 
between 2002 and 2010 (Urteaga et al. 2012). Between 2002 and 2014, 
Salazar et al. (2019) recorded 340 nests, indicating a downward trend. 
Considering the best available data, nesting has declined in Nicaragua.
    Nesting females of the East Pacific DPS are generally smaller and 
produce fewer eggs per clutch than turtles from other leatherback 
populations (Sarti et al. 2007; Piedra et al. 2007; Santidri[aacute]n 
Tomillo et al. 2007). For example in Mexico, nesting females have a 
mean size of 144 cm CCL and 62 eggs per clutch; the average total 
fecundity per females was estimated to be 341 eggs per season, with a 
maximum of 744 eggs deposited in a season (Sarti et al. 2007). The low 
productivity parameters, drastic reductions in overall nesting female 
abundance, and current declines in nesting place the DPS at risk of 
extinction, especially given the limited nesting female abundance.

Spatial Distribution

    The DPS is characterized by somewhat continuous and low density 
nesting across long stretches of beaches along the coast of Mexico and 
Central America. Santidri[aacute]n Tomillo et al. (2017) found a 
contraction of the Costa Rica's overall nesting distribution since the 
1990s.
    The best available genetic data indicate a high degree of 
connectivity among nesting aggregations. Dutton et al. (1999) did not 
find any genetic differentiation between nesting populations in Mexico 
(Playa Mexiquillo) and Costa Rica (Playa Grande) based on analysis of 
mtDNA control region sequences. Additional analyses of mtDNA sequences 
and nuclear DNA (microsatellites) from three index nesting beaches in 
Mexico also failed to find genetic differentiation (Barragan and Dutton 
2000; Dutton et al. unpublished).
    Based on monitoring of tagged nesting females, researchers 
documented female interchange between nesting beaches within Mexico and 
within Costa Rica. However, only one interchange has been documented 
between Mexico and Costa Rica (Sarti et al. 2007). Interchange between 
nesting beaches may occur during or between nesting seasons and may 
depend on the distance between nesting sites, which can be fairly 
large, especially in Mexico. For example, the distance between Tierra 
Colorada and Cahuit[aacute]n is 25 kilometers, and up to 18.7 percent 
of nesting females visit both beaches within a season (average of nine 
percent). Mexiquillo is located approximately 475 kilometers from the 
closest other nesting beach (Tierra Colorada), and researchers found no 
interchange of females within seasons. However, a few females were 
found to nest in either Mexiquillo and/or Tierra Colorado between 
seasons (Sarti et al. 2007).
    In Costa Rica, nesting females move among the three nesting beaches 
of Las Baulas National Park, within and between seasons, particularly 
between Playa Grande and Playa Langosta, although researchers study 
both Playa Grande and Playa Ventanas in combination. According to data 
gathered over 10 years of research (mid 1990s through the mid-2000s), 
an average of 71 percent of females nested only on Playa Grande, 10 
percent nested only on Playa Langosta, and 18 percent nested on both 
beaches in a given season. In other seasons, females have been shown to 
shift and nest primarily on a different beach. Within two seasons, 82 
percent of nesting females at Playa Langosta also nested at Playa 
Grande and 100 percent of nesting females at Playa Langosta within 
three seasons occasionally also nested at Playa Grande 
(Santidri[aacute]n Tomillo et al. 2007). At the less abundant nesting 
beaches in Costa Rica, the exchange rate between females ranged between 
7 and 28 percent. For example, at Ostional, 12 out of the 43 identified 
females were observed at least once at other sites (28 percent), while 
at Naranjo, 4 out of 21 identified females were also observed at other 
beaches (19 percent). At Cabuyal, 2 out of 15 turtles were observed at 
other beaches (13 percent), while 1 out of 15 females at Caletas were 
observed elsewhere (7 percent) (Santidri[aacute]n Tomillo et al. 2017).
    The foraging range of the DPS extends into coastal and pelagic 
waters of the southeastern Pacific Ocean. Individuals forage in the 
Pacific Gyre ecosystem and along the coasts of Peru and Chile, with 
variation resulting from the location of upwelling and ENSO effects. 
Researchers have hypothesized that high bycatch along the coastal 
foraging phenotype in this population (Saba et al. 2007). Recently, 
Harrison et al. (2018) determined that post-nesting females from Las 
Baulas National Park spent 78.2 percent of their time on the

[[Page 48403]]

high seas, 17.8 percent of their time in Costa Rica's EEZ, and 3.7 
percent of their time around the Galapagos Islands.
    Multiple nesting and foraging distributions likely help to buffer 
the DPS against local catastrophes or environmental changes that would 
otherwise modify nesting habitat or limit prey availability. Nesting 
aggregations are largely connected. However, there is less exchange 
among distant nesting beaches. Foraging turtles are vulnerable to 
perturbations in ocean conditions due to climate change, ENSO, and the 
Pacific Decadal Oscillation.

Diversity

    The East Pacific DPS exhibits genetic diversity, as demonstrated by 
moderate to high mtDNA haplotypic diversity (h = 0.66-0.71; Dutton et 
al. 1999). Such diversity likely provides the DPS with some capacity 
for adapting to long-term environmental changes, such as cyclic or 
directional changes in ocean environments due to natural and human 
causes (McElhany et al. 2000; NMFS 2017). Nesting habitat is mainly 
restricted to mainland beaches along the same coast. The DPS does not 
exhibit temporal or seasonal nesting diversity, with most nesting 
occurring between October and March. This limits resilience. For 
example, short-term spatial and temporal changes in the environment are 
likely to affect all nesting females in a particular year. The foraging 
strategies are somewhat diverse, with turtles foraging in coastal and 
oceanic waters. However, most turtles forage in the East Pacific Ocean, 
where they are similarly exposed to the effects of climate change, 
ENSO, or the Pacific Decadal Oscillation. Thus, the DPS has limited 
resilience.

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    The destruction or modification of habitat is a threat at many 
nesting beaches used by turtles of the East Pacific DPS. Foraging 
habitat has also been characterized as marginal, particularly in the 
eastern tropical Pacific Ocean (pelagic environment) due to relatively 
low productivity. Coastal habitat, which is normally associated with 
high productivity, may have been marginalized due to high levels of 
interactions with coastal artisanal fisheries.
    Development threatens the DPS by modifying the preferred beach 
habitat for nesting. Sustained and substantial development along the 
northern and southern ends of the nesting beach at Playa Grande in Las 
Baulas National Park, and in adjacent areas, has resulted in the loss 
of nesting beach habitat in addition to the removal of much of the 
natural beach vegetation. As a result, erosion has increased and led to 
other environmental damages to sand that are associated with human 
development, including significant changes to elevation, water content, 
particle size, pH, salinity, organic content and calcium carbonate 
content (Clune and Paladino 2008). Within the past two decades, 
beachfront development in the town of Tamarindo (across Tamarindo Bay 
from Playa Grande) has resulted in the degradation of nesting beach 
habitat, including: Pollution from artificial light, solid and chemical 
wastes, beach erosion, unsustainable water consumption, and 
deforestation. Hotels in this area have replaced a significant 
leatherback nesting area at Playa Tamarindo, which hosted significant 
nesting in the 1970s and 1980s (Wallace and Piedra 2012). Playa 
Langosta, which is just across from Tamarindo, is inundated with lights 
and noise from the town (Wallace and Piedra 2012). Currently, 
development has been curtailed due mainly to water issues (i.e., 
drought). Any additional development would damage the current 
hydrology. The Leatherback Trust, a local nonprofit working at Las 
Baulas National Park, has acquired some properties to prevent 
development, but property costs have increased over time. At Las Baulas 
National Park, 10 percent of nests were being inundated by tidal flows. 
To mitigate this threat, nests at risk of tidal inundation were 
relocated to another site on the same beach or into a hatchery. 
Hatchling production slightly increased due to the establishment of the 
hatchery, where approximately two percent of hatchlings were produced 
from 1998 to 2004 (Santidri[aacute]n Tomillo et al. 2007). We conclude 
that coastal development in Costa Rica is a threat to this DPS.
    In Mexico, the extent of development near nesting beaches is 
generally low, given the remoteness of the beaches in Baja California 
and on the mainland. Reviewing the location of these nesting beaches, 
we found very few roads or development nearby. The main nesting beaches 
remain somewhat isolated, with very few roads or development adjacent 
to the nesting beaches. Thus, there is limited threat due to artificial 
lighting and generally little to no beach driving except perhaps that 
associated with monitoring efforts (L. Sarti, CONANP, pers. comm., 
2018). In 2002, the Commission for Natural Protected Areas designated 
two of the index beaches (Mexiquillo and Tierra Colorada) as natural 
protected areas (turtle sanctuaries), which helped protect nesting 
habitat. Subsequently, in 2003, three of the index beaches (Mexiquillo, 
Tierra Colorada, and Cahuit[aacute]n) were listed as Ramsar Sites, 
which are wetland sites designated to be of international importance 
under the Ramsar Convention.
    At Veracruz de Acayo beach in Nicaragua, Salazar et al. (2019) note 
that while conservation efforts has reduced the threat of poaching, the 
establishment of tourism-focused coastal development that do not comply 
with the existence of management plans could threaten the nesting 
habitat.
    While nesting beaches within this DPS are generally remote and/or 
protected due to monitoring and existence of national parks and 
wildlife refuges, nesting females, hatchlings, and eggs at Las Baulas 
National Park (Costa Rica) nesting beaches are exposed to the 
modification of nesting habitat, as a result of development. This 
threat impacts the DPS by reducing nesting and hatching success, thus 
lowering the productivity of the DPS. We conclude that habitat loss and 
modification is a threat to the East Pacific DPS.

Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    The harvest of nesting females and eggs was the primary cause of 
the historical decline in abundance of the East Pacific DPS. Since 
then, laws have been passed to protect eggs and turtles. However, 
poaching still occurs.
    In Mexico, Sarti et al. (2007) attributed the decline of nesting 
females to the killing of adult females and intensive egg harvest. 
Adult females were historically killed at nesting beaches and in open 
waters (Sarti et al. 1994; Sarti et al. 1998). Since 1990, the harvest 
of turtles and eggs has been prohibited by national legislation. 
However, poaching pressure remains high wherever beach patrols do not 
occur (Santidri[aacute]n Tomillo et al. 2017). For example, Mexiquillo 
produced hatchlings every season in the 1980s. However, even with 
efforts to protect the nests in place, 60 to 70 percent of the total 
number of clutches were poached. Nichols (2003) notes that leatherback 
turtles were once harvested off Baja California, but their meat is now 
considered inferior for human consumption. At present, leatherback 
turtles are not generally captured for their meat or skin, but the 
poaching of nesting females has been known to occur on beaches such as 
Piedra de Tlacoyunque, Guerrero (Sarti et al. 2000).

[[Page 48404]]

    Although poaching of turtles and eggs has been consistently reduced 
over the years, it still occurs at high levels. Effective conservation 
and protection depends on human presence at the nesting beaches 
(Santidri[aacute]n Tomillo et al. 2017). Without such protection, 
poaching is likely to escalate. This may have occurred at one of the 
primary nesting beaches (Mexiquillo), where monitoring and conservation 
has not taken place in recent years due to safety concerns (L. Sarti, 
CONANP, pers. comm., 2018). Since the mid-1990s, Proyecto La[uacute]d 
has been relocating clutches (usually within 1-2 hours of deposition) 
to protected fenced areas and releasing hatchlings in different areas 
of the beach. These efforts are intended to protect the eggs from 
poachers/predators and the hatchlings from predators (Sarti et al. 
2007).
    In Costa Rica, the population decline was predominantly caused by 
egg harvest. Ninety percent of eggs were collected on one of the major 
nesting beaches, Playa Grande, a decade or more prior to the reduction 
of nesting females (Santidri[aacute]n Tomillo et al. 2007). In the 
1950s, there were few nesting females at Playa Grande (Wallace and 
Piedra 2012). In the late 1960s and early 1970s, the number of nesting 
turtles increased to more than 100 nesting females nightly (Wallace and 
Piedra 2012). In the early 1970s, newly constructed roads provided 
access to people from distant villages and cities, and egg harvest 
increased to more than 90 percent by the late 1970s (Wallace and Piedra 
2012). Such high levels of egg harvest persisted for nearly two decades 
(Wallace and Saba 2009). Despite protection of nesting beaches at Las 
Baulas National Park, illegal poaching of eggs still occurs, though 
rarely. The black market for eggs remains strong; local bars throughout 
Guanacaste and elsewhere continue to offer shots of raw sea turtle egg 
yolks accompanying beer or liquor (Wallace and Piedra 2012).
    In 1991, the Parque Nacional Marino Las Baulas was created and 
subsequently ratified by law in 1995. The Park consists of three 
leatherback nesting beaches: Playa Grande, Playa Ventanas, and Playa 
Langosta. The establishment of the park ensured increased protection at 
all three nesting beaches, greatly reducing egg poaching in the area. 
Poaching of eggs was reduced from 90 percent prior to 1990/1991, to 50 
percent in 1990/1991, 25 percent in 1991 through 1993, and near 0 
percent in 1993/1994 (Santridi[aacute]n Tomillo et al. 2007). To 
mitigate poaching, nests are often relocated. However, relocation may 
reduce hatching success (reviewed in Hern[aacute]ndez et al. 2007; 
Eckert et al. 2012). In Playa Grande, Costa Rica, fewer females were 
produced in translocated nests; cooler nests due to a lower number of 
metabolizing embryos may have reduced hatchling success (Sieg et al. 
2011).
    In Nicaragua, prior to protection in the early 2000s, poachers took 
nearly 100 percent of the nests at the three nesting beaches. Nesting 
beach protection has occurred at Veracruz since 2002, Juan Venado since 
2004, and Salamina since 2008. An average of ten community team members 
(mostly ex-poachers) monitor beaches seasonally. From 2002 to 2010, up 
to 420 nests were recorded and an estimated 94 were protected (Urteaga 
et al. 2012). While Veracruz de Acayo and Salamina are protected at 100 
percent, Isla Juan Venado is not permanently monitored. Therefore, 
poaching is likely to occur. Poaching occurs at high levels at other 
beaches, such as Playa El Mogote. During the 2001/2002 nesting season, 
23 of 29 nests were poached (79 percent), and the remaining six nests 
were protected in a hatchery (Arauz 2002). Due to the high level of 
poaching in this area, when possible, researchers from Flora & Fauna 
International relocated 98 nests between 2002 and 2004. However, these 
nests had a low emergence rate (22 percent; Urteaga and Chac[oacute]n 
2008).
    Extensive and prolonged effects of comprehensive egg harvest have 
depleted the leatherback population in Costa Rica and Mexico, with egg 
harvest levels of nearly 90 percent for about two decades (Sarti et al. 
2007; Santidri[aacute]n Tomillo et al. 2008; Wallace and Saba 2009). 
Currently, nesting females and eggs of the East Pacific DPS are exposed 
to poaching. Though efforts have reduced the levels of poaching of both 
eggs and nesting turtles, egg poaching remains high and affects a large 
proportion of the DPS. Poaching of nesting females reduces both 
abundance (through loss of nesting females) and productivity (through 
loss of reproductive potential). Such impacts are high because they 
directly remove the most productive individuals from DPS, reducing 
current and/or future reproductive potential. Egg harvest reduces 
productivity only, but over a long period of time, this also reduces 
recruitment and thus abundance. Given the high exposure and impacts, we 
conclude that overutilization, as a result of poaching, poses a major 
threat to the DPS.

Disease or Predation

    Little is known about diseases and parasites in leatherback 
turtles, although fibropapillomatosis has been described as a major 
epizootic disease in hard shelled turtles. A fibropapilloma tumor (in 
regression) was found on one nesting female at Mexiquillo, Mexico in 
1997 (Huerta et al. 2002). Various bacteria have also been documented 
in leatherback eggs. Soslau et al. (2011) sampled eggs laid on a Costa 
Rican beach to determine if bacteria were contributing to the low 
hatching rate (50 percent). The bacteria identified (i.e., species of 
the Bacillus, Pseudomonas, and Aeromonas genera) are known pathogens to 
humans and may account for developmental arrest of the turtle embryo 
(Soslau et al. 2011).
    Numerous predators prey on East Pacific leatherback turtles 
throughout their life stages. Eggs and hatchlings are eaten by crabs, 
ants, birds, reptiles, mammals, and fish (Eckert et al. 2012). In Costa 
Rica, during the 1993/1994 nesting season, several nests were lost to 
predation and infestation by maggots (Schwandt et al. 1996). In the 
Nicoya Peninsula, on the Pacific coast of Costa Rica, Squires (1999) 
documented evidence of potential nest predation by dogs, coyote, and 
raccoon. Predation of hatchlings by dogs and raccoons has increased in 
Playa Grande due to an increase in development in the area (P. 
Santridi[aacute]n Tomillo, The Leatherback Trust, pers. comm., 2019).
    For adult turtles, principal predators at sea include killer 
whales, crocodiles (Pritchard 1981), and sharks, while nesting females 
are taken by crocodiles (Bedding and Lockhart 1989), tigers, and 
jaguars (Pritchard 1971). Sarti et al. (1994) observed a lone male 
killer whale feeding on a single gravid female near Michoac[aacute]n, 
Mexico, apparently consuming only certain parts of the turtle and 
discarding others (e.g., female reproductive organs). In summary, eggs, 
hatchlings, and some adults are exposed to predation. For this DPS, the 
primary impact is to productivity (i.e., reduced egg and hatching 
success). Predation on nesting females, while rare, reduces abundance 
and productivity. Nest predation is mitigated through screening of 
nests, relocation of nests to hatcheries and releasing hatchlings in 
safer areas of the beach, and protecting nesting females from large 
predators such as dogs and jaguars (Sarti et al. 2007); some of these 
efforts are funded through the MTCA. We conclude that predation is a 
threat to the East Pacific DPS.

Inadequacy of Existing Regulatory Mechanisms

    Several international regulatory mechanisms apply to turtles in 
this DPS. The IAC, in particular, prohibits the harvest of turtles and 
eggs. CITES

[[Page 48405]]

limits all international trade of the species. There are also 
international efforts to reduce fisheries bycatch.
    In 2015, at the 7th Conference of the Parties, the IAC resolved to 
prioritize conservation actions in their work programs that would help 
``reverse the critical situation of the leatherback sea turtle in the 
Eastern Pacific.'' Specifically, parties were urged to: (1) Submit 
leatherback bycatch information annually to the IAC Secretariat; (2) 
improve leatherback turtle fishery monitoring efforts through the use 
of on-board observers; (3) report annually on the measures they have 
taken to reduce leatherback bycatch in their fisheries; (4) enhance 
leatherback nest monitoring and protection to increase hatchling 
survival and protect nesting beach habitat; (5) foster safe handling 
and release of incidentally bycaught leatherback turtles in fisheries; 
and (6) agree to a five-year strategic plan containing key activities 
related to the resolution (CIT-COP7-2015-R2). The strategic plan was 
patterned after the Regional Action Plan for Reversing the Decline of 
the Eastern Pacific Leatherback (http://savepacificleatherbackturtles.org) and included measures to reduce 
fisheries bycatch of adult and subadult leatherback turtles, the 
identification of high risk areas with fisheries and leatherback 
turtles, the identification and protection of important areas for 
leatherback turtle survival in different life stages, the elimination 
of any consumption and illegal use of leatherback turtles, and nesting 
site protection.
    As mandated by the 1994 North American Agreement for Environmental 
Cooperation, the Commission for Environmental Cooperation (CEC) 
encourages Canada, the United States, and Mexico to adopt a continental 
approach to the conservation of flora and fauna. In 2003, this mandate 
was strengthened as the three North American nations launched the 
Strategic Plan for North American Cooperation in the Conservation of 
Biodiversity. The North American Conservation Action Plan (NACAP) 
initiative began as an effort promoted by the three nations, through 
the CEC, to facilitate the conservation of marine and terrestrial 
species of common concern. In 2005, the CEC supported the development 
of a NACAP for Pacific leatherback turtles by Canada, the United 
States, and Mexico. Identified actions in the plan addressed three main 
objectives: (1) Protection and management of nesting beaches and 
females; (2) reducing mortalities from bycatch throughout the Pacific 
Basin; and (3) waste management, control of pollution, and disposal of 
debris at sea.
    In 2015, the Eastern Pacific Leatherback Network (also known as La 
Red de la Tortuga La[uacute]d del Oc[eacute]ano Pacifico (Red 
La[uacute]d OPO) (www.savepacificleatherbacks.org)) was formed to 
address the critical need for regional coordination of East Pacific 
leatherback conservation actions to track conservation priorities and 
progress at the population level. This network has brought together 
conservationists, researchers, practitioners and government 
representatives from 22 institutions across nine East Pacific nations 
with varying priorities, capacities and historical experiences in 
leatherback research and conservation to contribute to shared 
activities, projects, and goals. Through these efforts, Red La[uacute]d 
OPO now has mutually-agreed upon mechanisms for sharing information and 
data, as well as standardized protocols for nesting beach monitoring 
and bycatch assessments/fishing practices.
    The Convention for the Protection of Natural Resources and 
Environment of the South Pacific, also known as the Noumea Convention, 
has been in force since 1990 and includes 26 Parties (as of 2013). The 
purpose of the Convention is to protect the marine environment and 
coastal zones of the South-East Pacific, and beyond that area, the high 
seas up to a distance within which pollution of the high seas may 
affect that area.
    In 2015, the IATTC passed a resolution that requires large longline 
vessels fishing in the eastern tropical Pacific Ocean to carry 
observers. Cooperating parties that have documented interactions with 
sea turtles in their longline fleet are required to maintain at least 
five percent observer coverage and provide an annual report to the 
IATTC. Unfortunately, the forms used by observers to report incidents 
are not standardized, so in some cases, the reports did not include 
species identification, condition of the released turtles, and location 
of the interactions, and the five percent minimum coverage is often not 
met. Nations without reported bycatch of sea turtles simply provided a 
statement to that effect. In the few reports we reviewed, leatherback 
turtles comprised some of the bycatch in the eastern tropical Pacific 
Ocean, but there were few details on the events (C. Fahy, NMFS, pers. 
comm., 2018). In 2007, the IATTC passed a resolution requiring nations 
to conduct research on sea turtle bycatch reduction measures in their 
longline fleets (e.g., use of circle hooks and fish bait). Despite 
results in both the Atlantic and Pacific longline fleets showing that 
use of circle hooks/fish bait significantly reduced leatherback bycatch 
rates (Swimmer et al. 2017), nations are not required to use this hook/
bait combination. In 2017, at an IATTC sea turtle bycatch reduction 
workshop, the United States presented findings on longline bycatch 
reduction and proposed a stronger resolution that would require use of 
this methodology. However, some nations resisted, and the resolution 
did not move forward for consideration at the annual IATTC meeting.
    Throughout the world, illegal, unreported, and unregulated (IUU) 
fishing leads to underestimates of bycatch. In Mexico, there is a lack 
of effective fisheries governance, resulting in highly uncertain 
fishery statistics. For example, from 1950 to 2010, total fisheries 
catch, including estimated IUU catch and discarded bycatch, was nearly 
twice as high as the official statistics (Cisneros-Montemayor et al. 
2013). Thus, the bycatch threat of commercial fisheries in Mexico may 
be higher than currently estimated.
    In addition, several international treaties and/or regulatory 
mechanisms protect East Pacific leatherback turtles. While no single 
law or treaty can be 100 percent effective at minimizing anthropogenic 
impacts to sea turtles in these areas, there are several international 
conservation agreements and laws in the region that, when taken 
together, provide a framework within which sea turtle conservation 
advances can be made (Frazier 2012). In addition to protection provided 
by local marine reserves throughout the region, sea turtles may benefit 
from the following broader regional effort: (1) The Eastern Tropical 
Pacific (ETP) Marine Corridor (CMAR) Initiative supported by the 
governments of Costa Rica, Panama, Colombia, and Ecuador, which is a 
voluntary agreement to work towards sustainable use and conservation of 
marine resources in these nations' waters; (2) the ETP Seascape Program 
managed by Conservation International that supports cooperative marine 
management in the ETP, including implementation of the CMAR; (3) the 
IATTC and its bycatch reduction efforts through resolutions on sea 
turtles, observer coverage, etc.; (4) the IAC, which is designed to 
lessen impacts on sea turtles from fisheries and other human impacts; 
and (5) the Permanent Commission of the South Pacific (Lima 
Convention), which has developed an Action Plan for Sea Turtles in the 
Southeast Pacific.

[[Page 48406]]

    Most nations within the range of the East Pacific DPS have laws 
prohibiting the harvest of turtles and eggs. This applies to nesting 
turtles and those captured at sea. National laws in Mexico (1990 
Presidential Decree), Costa Rica (2002 Presidential Decree N[deg]8325: 
The Law of Protection, Conservation, and Recuperation of Marine 
Turtles), and Nicaragua (Law No. 651 and Ministrial Resolution No. 043-
2005) protect nesting females and eggs and nesting beaches. However, 
poaching remains a major threat. Although laws prohibit the harvest of 
turtles in Peru, fishermen consume leatherback turtles bycaught in 
small-scale fisheries (Alfaro-Shigueto et al. 2011), indicating 
inadequate enforcement of existing laws. In other nations where 
leatherback turtles of this DPS are bycaught, the turtles are released 
and not retained (e.g., Chile; Donoso and Dutton 2010).
    Several protected areas have been established throughout the range 
of the DPS. Most of the nesting beaches in Mexico and Costa Rica are 
protected from egg and turtle poaching, with effective monitoring to 
ensure low levels of poaching. Poaching likely continues at unprotected 
and remote beaches, and at those that contain an extensive coastline 
that is difficult to monitor and protect. Protected nesting beaches in 
Mexico include: Mexiquillo (until 2013); Playa de Tierra Colorada, 
Playa Cahuit[aacute]n, Playa San Juan, Bahia de Chacahua, and Playa 
Barra de la Cruz. Protected nesting beaches in Costa Rica include: Las 
Baulas National Park (Playa Grande, Playa Langosta, and Playa 
Ventanas), Naranjo (National Park), Cabuyal (under no official 
management category), Nombre De Jes[uacute]s (under no official 
management category), Ostional (wildlife refuge), and Caletas (wildlife 
refuge). Protected nesting beaches in Nicaragua include: Salamina-Costa 
Grande, Veracruz de Acayo (Chacocente Wildlife Refuge).
    Marine protected areas also exist. The waters of the Las Baulas 
National Park, which represents a hotspot for inter-nesting females and 
breeding males, are protected out to 22.2 km as a no-take zone for all 
fishing activity. However, satellite telemetry data for nesting females 
at these beaches over three seasons revealed that the turtles move well 
outside these boundaries during their inter-nesting period, which makes 
them vulnerable to fisheries outside the park (Shillinger et al. 2010). 
Data from 44 females that were tagged off Las Baulas National Park 
revealed a high use habitat within 6 nm from the nesting beaches, but 
overall revealed a generally large range, covering over 33,000 km\2\, 
from the Nicoya Peninsula, east into the Gulf of Nicoya in Costa Rica, 
and north to coastal habitats within 30 kilometers offshore from 
southern Nicaragua. The marine areas adjacent to this protected 
boundary are not managed under any type of status (Shillinger et al. 
2010). Fisheries within Costa Rica and Nicaragua's EEZ include trawl, 
gillnet and longline that continue to operate.
    In summary, numerous regulatory mechanisms exist to protect 
leatherback turtles, eggs, and nesting habitat throughout the range of 
this DPS. Although the regulatory mechanisms provide some protection to 
the species, many do not adequately reduce the threat that they were 
designed to address, generally as a result of limited implementation or 
enforcement. As a result, bycatch, incomplete nesting habitat 
protection, and poaching remain threats to the DPS. We conclude that 
the inadequacy of existing regulatory mechanisms is a threat to the 
East Pacific DPS.

Fisheries Bycatch

    Bycatch in commercial and recreational fisheries, both on the high 
seas and off the coasts, is the primary threat to the East Pacific DPS. 
This threat affects the DPS by reducing the abundance of all life 
stages of the DPS (with the likely exception of hatchlings).
    Integrating catch data from over 40 nations and bycatch data from 
13 international observer programs, Lewison et al. (2004) estimated the 
numbers of leatherback turtles taken globally by pelagic longliners to 
be more than 50,000 leatherback turtles in just one year (2000). With 
over half of the total fishing effort (targeting tuna and swordfish) 
occurring in the Pacific Ocean, an estimated 20,000 to 40,000 
leatherback turtles interacted with longline fishing during the year 
studied. Fishing effort was highest in the central South Pacific Ocean 
(south of Hawaii), which overlaps with the foraging range of this DPS. 
Because observers are in place on only a fraction of longline vessels 
in the eastern tropical Pacific Ocean, and a requirement came into 
effect only recently through an IATTC resolution, these estimates are 
likely a minimum. More recently, Molony (2005) and Beverly and Chapman 
(2007) estimated sea turtle longline bycatch to be approximately 20 
percent of that estimated by Lewison et al. (2004), or approximately 
200 to 640 leatherback turtles annually. Where tuna species are 
targeted, bycatch of turtles in the deep-set longline gear often 
results in mortality due to drowning. Additional studies indicate the 
high impact of industrial longline fleets on leatherback turtles (e.g., 
Spotila et al. 1996, 2000).
    In their global study of sea turtle bycatch, where available, 
Wallace et al. (2013) found that longline bycatch had a low impact, but 
that net bycatch had a high impact on the East Pacific RMU. The impact 
of local artisanal fleets (using gillnets and longlines) that fish 
closer to shore is less documented.
    In Mexico, leatherback turtles wash to shore entangled in longlines 
and driftnet, indicating interaction and mortality (Sarti et al. 2007). 
Ortiz-Alvarez et al. (2019) conducted a bycatch survey across 48 
different ports (933 fishers) in Mexico, Nicaragua and Costa Rica 
between October 2016 and July 2017 in an effort to improve the 
understanding of leatherback bycatch in artisanal fisheries, 
particularly where data are lacking. The surveys represented on average 
over 30 percent of the fishing fleet per port for both Nicaragua and 
Costa Rica and 6 percent per port for Mexico. In Mexico, where gillnets 
were the most frequently reported gear, fishers (n = 709) reported an 
estimated bycatch of 300 leatherback turtles in the previous year, with 
65 percent in ``good condition;'' 76 percent of fishers released 
turtles alive (three percent consumed or sold the turtles). Estimated 
average bycatch rates per vessel were 1.0 for Costa Rica and Nicaragua 
and 2.3 for Mexico. In Costa Rica, leatherback turtles were primarily 
caught in longlines and released alive; 75 percent of the Costa Rican 
fishermen reported that bycaught leatherback turtles were in ``good 
condition.'' In Nicaragua, where gillnets were the most frequently 
reported gear, 18 percent of fishers reported that leatherback turtles 
were in ``good condition;'' 76 percent of fishers released turtles 
alive (six percent consumed or sold the turtles; Ortiz-Alvarez et al. 
(2019).
    Recent surveys of 765 Ecuadorian, Peruvian, and Chilean fishermen 
(at 43 ports, representing 28 to 63 percent of ports) reported the 
following leatherback interaction rates (as a percentage of total 
interactions with sea turtles): 2.81 percent of 40,480 interactions 
(32.5 percent mortality) in Ecuador, 14.87 of 5,828 interactions (50.8 
percent mortality) in Peru, and 27.83 percent of 170 interactions (3.2 
percent mortality) in Chile (Alfaro-Shigueto et al. 2018). Mortality 
rates reported for all sea turtles were 3.2 percent in Chile, 32.5 
percent in Ecuador, and 50.8 percent in Peru (Alfaro-Shigueto et al 
2018).
    The swordfish gillnet fisheries in Peru and Chile may have 
contributed to the decline of the DPS. The decline in the nesting 
population at Mexiquillo

[[Page 48407]]

occurred at the same time that effort doubled in the Chilean driftnet 
fishery (Eckert 1997). Using data collected from Frazier and Montero 
(1990) regarding leatherback takes in a swordfish gillnet fishery from 
one port in Chile (San Antonio), and extrapolating to other ports in 
Chile and Peru, with an increased level of effort observed through the 
mid-1990s, Eckert (2007) estimated that a minimum of 2,000 leatherback 
turtles were killed annually by the combined swordfish fishing 
operations (only gillnet) off Peru and Chile. After some fleets 
switched from large mesh gillnet to longline to target swordfish, this 
estimate has declined by at least an order in magnitude. Research 
conducted in the Chilean large-mesh gillnet fishery to reduce bycatch 
of marine mammals and sea turtles indicates that less than five 
leatherback turtles have interacted with the fishery (on observed 
vessels) since 2014, and all were released alive (C. Fahy, NMFS, pers. 
comm., 2018).
    In Peru, the capture of leatherback turtles has been prohibited 
since 1976, although retention of bycaught leatherback turtles 
continues (FAO 2004). From 1985 to 1999, based on field books, diaries, 
specimen data sheets, fishery statistics files and unpublished reports, 
30 leatherback turtles were captured in fisheries (in Alfaro-Shigueto 
et al. 2007). From July 2000 to November 2003, observers at 8 ports, 
from Mancora in northern Peru to Morro Sama in the south, reported 133 
leatherback turtles caught by artisanal fishing gear, with 76 percent 
caught in gillnets and 24 percent caught in longlines targeting fish, 
sharks, and rays (Alfaro-Shigueto et al. 2007). Of the total caught, 
41.4 percent (n = 55) were released alive and 58.6 percent (n = 78) 
were retained for human consumption. Of the leatherback turtles 
retained and measured (n = 6), the size ranged from 98 to 123 cm curved 
carapace length (CCL), indicating that both subadults and adults are 
encountered by artisanal fisheries off Peru. Researchers recently 
assessed and quantified sea turtle mortality levels in one fishing 
village in central-southern Peru (San Andr[eacute]s) through sampling 
dump sites (97.3 percent) and strandings (2.7 percent) over a 5-year 
period (2009 to 2014). Of 953 carapaces recorded, leatherbacks 
comprised only 1.4 percent of sea turtles (n = 13). However, this study 
still confirmed that they were consumed or sold for human consumption. 
With a mean CCL of 113.0 cm (range: 80 to 135, n = 10), 70 percent of 
the leatherbacks were juveniles and 30 percent were sub-adults. There 
were no adults. Researchers noted that the meat was used to support 
separate demands: Fishermen families' consumption, local trade, and 
``special'' orders from Lima (Quispe et al. 2019). Using data from 
shore-based and on-board observers, Alfaro-Shigueto et al. (2011) 
estimated the mean annual leatherback bycatch as follows: 40 turtles 
(with a range of 37 to 44) in the driftnet fishery, with 80 percent 
released alive; six turtles (with a range of 3 to 9) in the dolphinfish 
longline fishery, all released alive; and 26 turtles (with a range of 
24 to 27) in the shark longline fishery, all released alive. Alfaro-
Shigueto et al. (2015) assessed the bycatch of leatherback turtles in 
driftnet vessels in northern Peru (through at-sea monitoring) and 
central Peru (shore-based monitoring). From December 2013 to November 
2014, 31 leatherback turtles were captured, of which 13 died. 
Interactions occurred primarily with juveniles and subadults (mean CCL 
was 125.1  14.8). Nearshore driftnets from San Jose 
(northern Peru) captured 20 leatherback turtles (five dead). At least 
one animal was butchered, indicating that even animals caught alive may 
be killed, despite Peruvian laws restricting such practices. 
Approximately 3,000 net vessels fish along the coast of Peru, but only 
a fraction were included in this study (Alfaro-Shigueto et al. 2015). 
Efforts are being made to patrol nets to reduce bycatch, conduct 
extensive education and outreach, and increase regulation and 
enforcement (Alfaro-Shigueto et al. 2015). A review of information 
collected from official statistics, literature, and surveys of beaches 
and dumpsites revealed that the size of captured leatherback turtles 
declined over the years. In 1987, the mean CCL of captured leatherback 
turtles was 117  10.65 cm, while in 2005, the mean CCL was 
109.27  14.4, possibly indicating overexploitation due to 
systematic and sustained harvests, particularly during El Ni[ntilde]o 
years (Campos et al. 2009). Greater captures of all sea turtles, 
including leatherback turtles, occurred during periods of El 
Ni[ntilde]o, when turtles are more likely to be found in more coastal 
waters (where there is increased artisanal fishery activity) due to 
environmental variability and availability of jellyfish in those areas 
(Campos et al. 2009).
    In Chile, a commercial fishery was established in 2001 that 
permitted longlining for swordfish (shallow-set) with the condition 
that all vessels were required to take an observer on board to collect 
information on bycatch. Between 2001 and 2005, over 10 million hooks 
were observed, and leatherback turtles were the most common species 
caught (n = 284), with the majority (n = 282) released alive. 
Leatherback turtles were caught primarily between 24[deg] S and 38[deg] 
S (furthest south was 38[deg]39' S and 84[deg]15' W) in less than 4 
percent of the sets with an overall mean of 0.0268 turtles per one 
thousand hooks. Size estimates revealed both juveniles and adults. 
Fishermen were trained to use the best practices for de-hooking, 
disentangling, and releasing sea turtles, which likely increased the 
survival rate of leatherback turtles (Donoso and Dutton 2010). 
Researchers recently presented information on the incidental capture of 
sea turtles in industrial and artisanal longlines, gillnets and 
artisanal espinel (i.e., small-scale handline or longline) fisheries 
all targeting swordfish off Chile (Z[aacute]rate et al. 2019). Over an 
8-year period (2006-2014), 182 leatherbacks were documented as bycatch 
(mortality of bycaught turtles was not reported). Over this study 
period, 44 percent of turtles were caught in industrial longline, 28 
percent in artisanal espinel, 17 percent in gillnets and 11 percent in 
artisanal longline (with sea turtle species undefined). Researchers 
noted that while observer coverage in the industrial longline fleet has 
been generally high (>70 percent of total fishing trips), the 
monitoring coverage of artisanal espinel and gillnets is very low (<3 
percent). Thus, these estimates of bycatch can be considered minimal. 
While the number of industrial and artisanal vessels has declined (from 
12 vessels in 2001 to 3 vessels in 2014, the number of artisanal 
espinel and gillnet vessels has not declined, remaining around 90 
vessels (Z[aacute]rate et al. 2019).
    We conclude that juvenile and adult life stages of the East Pacific 
DPS are exposed to high fishing effort throughout their foraging range 
and in coastal waters near nesting beaches. Mortality is also high in 
some fisheries, with reported mortality rates of up to 58 percent due 
in part to the use of gillnets and as well as consumption of bycaught 
turtles in Peru. As noted above, there have been efforts by individual 
nations and regional fishery management organizations to mitigate and 
reduce the threat of bycatch, but those efforts have not been 
successful at ameliorating the risks. We conclude that fisheries 
bycatch remains a major threat to the East Pacific DPS.

Pollution

    Pollution is a threat to the East Pacific DPS. Pollution includes 
contaminants, marine debris, and ghost fishing gear. The South Pacific 
Garbage Patch, discovered in 2011 and confirmed in

[[Page 48408]]

2017, contains an area of elevated levels of marine debris and plastic 
particle pollution, most of which is concentrated within the ocean's 
pelagic zone and in area where leatherback turtles forage for many 
years of their life. The area is located within the South Pacific Gyre, 
which spans from waters east of Australia to the South American 
continent and as far north as the Equator.
    Given the amount of floating debris in the Pacific Ocean (Lebreton 
et al. 2018), marine debris has the potential to be a significant 
threat to the East Pacific leatherback population. The precise impact 
cannot be quantified using the best available data. Leatherback turtles 
subsist primarily on jellyfish and other gelatinous zooplankton and may 
be prone to ingesting plastics resembling their food source (Mrosovsky 
1981; Schuler et al. 2013, 2015). Dead leatherback turtles have been 
found choked on plastic bags, and phthalates derived from plastics have 
been found in leatherback egg yolk (Lebreton et al. 2018).
    Prior to the early 1990s, high seas driftnet fisheries freely 
operated in the Pacific Ocean and interacted with thousands of sea 
turtles. Researchers estimated that over 1,000 leatherback turtles were 
taken by the combined fleets of Japan, Korea, and Taiwan during a one-
year period (Wetherall 1997). However, because genetic analyses of 
Pacific leatherback turtles were relatively new at that time, the data 
does not indicate the nesting beach origin of those bycaught 
leatherback turtles. In 1992, a UN moratorium banned high seas driftnet 
fisheries, so that active large scale driftnets no longer pose a threat 
to leatherback turtles. However, numerous discarded driftnets continue 
to entangle and drown leatherback turtles in a phenomenon known as 
``ghost fishing'' (Gilman et al. 2016),
    In 2007, the IATTC passed a resolution pertaining to sea turtle 
bycatch in purse seine and longline fisheries which primarily target 
tuna. In order to address the marine debris and potential interactions 
with sea turtles in the eastern tropical Pacific Ocean, fishermen are 
required to disentangle sea turtles entangled in fish aggregating 
devices, even if the device does not belong to the vessel.
    Only a few studies of levels or effects of toxins on leatherback 
turtles have examined effects to their health and fitness, as well as 
any effects to eggs and hatchlings. Sill et al. (2008) sampled non-
viable leatherback eggs and hatchlings that died in the egg chamber at 
Las Baulas National Park. Researchers analyzed the samples for metals 
and other toxicants to explore the relationship between pollution and 
hatching success for 30 females. Metal levels were highly variable, but 
there were no significant differences within and between groups of 
females, and none of the pesticides tested were present in the samples 
(Sill et al. 2008). Overall, the study found no relationship between 
metal concentrations and hatching success. The researchers postulated 
that eggs may take up some metals from the nest environment and deposit 
other metals in the egg shell, as unhatched eggs contained more nickel, 
copper, and cadmium and contained significantly less iron, manganese 
and zinc than dead hatchlings (Sill and Paladino 2008).
    As with all leatherback turtles, entanglement in and ingestion of 
marine debris and plastics is a threat that likely kills several 
individuals a year. However, data are not available because most 
affected turtles are not observed. Given the amount of pollution 
turtles are exposed to throughout their lifetime, this has the 
potential to be a significant threat to the East Pacific leatherback 
population, although the impact cannot be quantified using the best 
available data. We conclude that pollution is a threat to this DPS.

Oceanographic Regime Shifts

    The East Pacific DPS is affected by oceanographic regime shifts. In 
the eastern equatorial Pacific Ocean, reductions in productivity 
parameters are primarily associated with ENSO, during which sex ratios 
become biased up to 100 percent female (Santidri[aacute]n Tomillo et 
al. 2014). There is also an effect on hatching and emergence success in 
North Pacific Costa Rica (Santidri[aacute]n Tomillo et al. 2012): 
During El Ni[ntilde]o years, hatching success is very low due to dry 
and hot conditions on the nesting beaches and is high during La 
Ni[ntilde]a events due to increased precipitation in this area. La 
Ni[ntilde]a events are characterized by high phytoplankton 
productivity, cooler sea surface temperatures, enhanced precipitation 
in northwestern Costa Rica, and cooler air temperatures. These factors 
lead to increases in the biomass and distribution of gelatinous 
zooplankton, the primary food of leatherback turtles. Foraging success 
and the frequency of reproduction are enhanced following such periods 
of high primary productivity (Saba et al. 2007). Nesting seasons that 
follow the La Ni[ntilde]a events, result in peaks in the number of 
nesting females, higher than average hatching success and emergence 
rates, and a larger proportion of male hatchlings (Saba et al. 2012). 
Saba et al. (2008) found that a shift from 1 [deg]C to -1 [deg]C in the 
El Ni[ntilde]o sea surface temperature anomaly resulted in a five-fold 
increase in leatherback remigration probabilities at Playa Grande. Such 
large-scale regime shifts are likely to affect the entire DPS. 
Productivity is positively (La Ni[ntilde]a) or negatively (El 
Ni[ntilde]o) impacted. Wallace et al (2006) hypothesize that prey 
availability related to ENSO exacerbates the effects of fisheries 
bycatch mortality, resulting in declining trends. Because of the small 
abundance of the DPS, extended El Ni[ntilde]o events are likely to pose 
a threat to the East Pacific DPS.

Climate Change

    Climate change is a threat to the East Pacific DPS. The impacts of 
climate change include: Increases in temperatures (air, sand, and sea 
surface); sea level rise; increased coastal erosion; more frequent and 
intense storm events; and changes in oceanographic regimes and 
currents.
    Climate projections assessed by the IPCC indicate that Central 
America is very likely (defined as 90 to 99 percent probability; IPCC 
2007) to become warmer and likely (defined as 66 to 90 percent 
probability; IPCC 2007) to become drier by 2100 (Saba et al. 2012). In 
addition, climate variability is likely to change the strength and 
frequency of El Ni[ntilde]o events, although there is less scientific 
consensus on the frequency and magnitude of changes to these events. A 
climate-forced population dynamics model developed by Saba et al. 
(2012) showed sea surface temperatures to be highly correlated with 
large phytoplankton productivity throughout a 100-year projection to 
the year 2100. Relative to a stable nesting population given mean 
surface air temperatures and precipitation from 1975 to 1999, Saba et 
al. (2012) estimated that the nesting population at Playa Grande would 
decline at a rate of 7 (1) percent per decade over the next 
century of climate change under a scenario which considered increasing 
emissions from 2000 to 2100 (A2 scenario). Similar declines occurred 
for other scenarios (Special Report on Emissions Scenarios 2007). The 
nesting population was projected to remain stable up until around 2030 
but reduced 75 percent by the year 2100. Hatching success and emergence 
rates, which would decrease associated with 2.5 [deg]C warming of the 
nesting beaches, served as a primary driver of the decline. 
Santidri[aacute]n Tomillo et al. (2012) developed a similar climate 
forcing model, which considered projected changes associated with El 
Ni[ntilde]o events

[[Page 48409]]

and demonstrated that hatching success would decline from approximately 
42 to 18 percent by 2100, while emergence rates would decline between 
approximately 76 to 29 percent. The authors concluded that even with 
protection at the primary nesting beaches in Costa Rica, with the 
general warming of Central America in the near future, the chances of a 
new nesting area emerging with more ideal conditions (i.e., cooler and 
wetter) is unlikely (Santidri[aacute]n Tomillo et al. 2012).
    Increasing sand temperature is an existing threat to the DPS. The 
long-term data set on leatherback turtles nesting at Playa Grande, 
Costa Rica indicates reduced emergence success, skewed sex ratios, and 
increased hatchling mortality as a result of increased sand temperature 
(Santidri[aacute]n Tomillo et al. 2015). From 2004 to 2013, primary sex 
ratios fluctuated between a minimum sex ratio of 41 percent females 
(and the only year with a male-biased hatchling production) to 100 
percent females produced during two seasons (Santidri[aacute]n Tomillo 
et al. 2014). Low emergence success and low hatchling output (i.e., 
higher mortality as a result of high sand temperatures) were associated 
with a strongly biased female ratio, because these resulted from 
female-producing high temperatures. Variability in these results occur 
during and between nesting seasons, largely due to highly variable 
climatic conditions in northwestern Costa Rica, resulting in ``boom-
bust'' cycles in leatherback hatchling production and primary sex 
ratios (in Santidri[aacute]n Tomillo et al. 2014). Sand temperatures 
are projected to continue to increase, which will likely result in a 
further decline in the number of hatchlings produced (Santidri[aacute]n 
Tomillo et al. 2014). An increase in the percentage of females could 
potentially benefit the productivity of the DPS in the short-term. 
However, any such benefits would be tempered by the associated lower 
emergence and hatchling success rates. Relocation of sea turtle 
clutches that may be ``doomed'' due to high sand temperatures and 
inundation is a common conservation practice, particularly at areas 
with warming beaches. However, relocation is not always possible and is 
also associated with lower emergence and hatchling success rates.
    In addition to climate change influencing the nesting beach habitat 
of eastern Pacific leatherback turtles, the impacts of a warming ocean 
may also affect the environmental variables of their pelagic migratory 
and foraging habitat, which may further increase population declines. 
As mentioned previously, the preferred foraging habitat of eastern 
Pacific is characterized by relatively low sea surface temperatures and 
low levels of chlorophyll-a. Using information derived from satellite 
tracked leatherback turtles, which established migratory pathways and 
core foraging habitat (as summarized in Shillinger et al. 2008), in 
combination with generalized additive mixed models, researchers were 
able to project that between 2001 and 2100, there would be a net loss 
of the core foraging habitat of the DPS. The loss was predicted to be a 
15 percent decline over the next century (Willis-Norton et al. 2014). 
Depending on whether this population is able to shift their preferred 
migratory routes and foraging habitat over time (which is unclear), 
remigration intervals may shorten or lengthen, which could influence 
reproductive productivity.
    Climate change is a threat to the East Pacific DPS that affects 
nesting females (e.g., remigration interval and fitness), their progeny 
(e.g., hatching success, embryonic development, and feminization of 
hatchlings), and foraging subadult and adult leatherback turtles. 
Detrimental impacts of increased sand temperatures have already 
occurred and are likely to continue or worsen. Foraging areas will also 
be impacted via changes in ocean productivity, sea surface 
temperatures, and availability of prey.

Conservation Efforts

    There are numerous efforts to conserve the leatherback turtle. The 
following conservation efforts apply to turtles of the East Pacific DPS 
(for a description of each effort, please see the section on 
conservation efforts for the overall species): Convention on the 
Conservation of Migratory Species of Wild Animals, Convention on 
Biological Diversity, Convention on International Trade in Endangered 
Species of Wild Fauna and Flora, Convention for the Protection of the 
Marine Environment and Coastal Area of the South-East Pacific (Lima 
Convention), Convention for the Conservation and Management of Highly 
Migratory Fish Stocks in the Western and Central Pacific Ocean (WCPF 
Convention), Convention Concerning the Protection of the World Cultural 
and Natural Heritage (World Heritage Convention), Eastern Pacific 
Leatherback Network, Eastern Tropical Pacific Marine Corridor 
Initiative, FAO Technical Consultation on Sea Turtle-Fishery 
Interactions, IAC, MARPOL, IUCN, Ramsar Convention on Wetlands, RFMOs, 
Secretariat of the Pacific Regional Environment Programme, UNCLOS, and 
UN Resolution 44/225 on Large-Scale Pelagic Driftnet Fishing. Although 
numerous conservation efforts apply to the turtles of this DPS, they do 
not adequately reduce its risk of extinction.

Extinction Risk Analysis

    After reviewing the best available information, the Team concluded 
that the East Pacific DPS is at high risk of extinction. The DPS 
exhibits a total index of nesting female abundance of 755 females at 
monitored beaches. Such a limited nesting population size makes this 
DPS vulnerable to stochastic or catastrophic events that increase its 
extinction risk. This DPS exhibits a decreasing nest trend, which along 
with lower than-average productivity metrics, has the potential to 
further reduce abundance and increase the risk of extinction. The 
nesting range is somewhat limited to the Pacific Central American 
coast, with little diversity among sites. Thus, stochastic events could 
have catastrophic effects on nesting for the entire DPS, with no 
distant subpopulations to buffer losses or provide additional 
diversity. Most foraging occurs in the eastern Pacific Ocean, which is 
subject to oceanographic regimes shifts that expose the DPS to low-
productivity events. Based on these demographic factors, we find the 
DPS to be at risk of extinction as a result of past threats.
    Current threats also contribute to the risk of extinction of this 
DPS. Fisheries bycatch is the major threat, capturing, and often 
killing, turtles throughout their foraging areas, thus reducing 
abundance. There are few mechanisms in place, including internationally 
through the IATTC or other bilateral or international instruments and 
through monitoring and enforcement of coastal fisheries laws, to 
mitigate or reduce bycatch. Overutilization is also a major threat. 
Historically, harvest of turtles and eggs reduced the once high 
abundance of turtles to current low levels. The poaching of eggs 
continues, reducing productivity, especially at unprotected beaches, 
where egg collection may reach 100 percent and nesting females may also 
be at risk of poaching. The effects of climate change, including the 
observed and predicted increase in frequency and strength of ENSO 
events (i.e., oceanographic regime shifts), are threats to this DPS, 
given its restricted foraging range and the vulnerability of nesting 
beaches to high sand temperatures and low levels of rainfall, which 
affect sex ratios and emergence and hatching success (i.e., 
productivity). Additional threats include: Habitat loss and 
modification;

[[Page 48410]]

predation; and pollution. Development modifies nesting habitat. 
However, most beaches are protected throughout the nesting range. 
Though many regulatory mechanisms are in place, they do not adequately 
reduce the impact of these threats. Further, it is important to note 
that efforts (e.g., relocation) to protect and mitigate threats from 
the harvest of turtles and eggs, predation, and environmental impacts 
related to erosion and lethal temperatures are dependent upon the 
presence of monitoring or management programs. Some of these are 
dependent on funding from the MTCA. Even when undertaken, these efforts 
may not be successful.
    We determine, consistent with the Team's findings, that the East 
Pacific DPS is currently in danger of extinction. Its nesting female 
abundance and declining trend make the DPS highly vulnerable to 
threats. Though numerous conservation efforts apply to this DPS, they 
do not adequately reduce the risk of extinction. We conclude that the 
East Pacific DPS is currently in danger of extinction throughout its 
range and therefore meets the definition of an endangered species. The 
threatened species definition does not apply because the DPS is 
currently at risk of extinction (i.e., at present), rather than on a 
trajectory to become so within the foreseeable future.

Leatherback Turtle, Overall Species

    The petition under review sought specifically to identify the NW 
Atlantic population of leatherback sea turtles as a separate DPS and 
assign it a different status from the global listing. As explained 
throughout this finding, we have determined that seven leatherback 
populations would satisfy the tests for recognition under our DPS 
Policy (i.e., that they are discrete from one another and significant 
to the overall species), and we have referred to these hypothetically, 
for purposes of our analysis only, as DPSs. This includes the NW 
Atlantic DPS. However, we have also determined that, even if these 
populations were formally recognized as DPSs through a listing process 
under the Act, each of the DPSs would have the same status as the 
overall species, which is currently listed throughout its range 
(globally) as endangered. Nothing in the petition or in the best 
available information we have reviewed has led us to conclude that 
there is any basis to disturb the long-standing global listing, which 
remains in effect and is unaffected by this finding. For completeness, 
here we present an overview of current information pertaining to the 
status of the overall species, including a summary of some of the key 
information from the DPS-specific sections as well as an evaluation of 
the demographic factors affecting the overall species.
    As explained in the Background section, the leatherback turtle was 
originally listed as endangered in 1970 under the precursor to the ESA 
and was carried forward as an ``endangered species'' when the ESA 
became effective. The Services designated the nesting beaches at Sandy 
Point, St. Croix (43 FR 43688; September 26, 1978) and surrounding 
marine waters (44 FR 17710; March 23, 1979) as critical habitat. NMFS 
designated additional marine habitat along 41,914 square miles (108,558 
square km) of the U.S. West Coast as critical habitat (77 FR 4170; 
January 26, 2012). The Services issued the recovery plans for 
leatherback turtles in the U.S. Caribbean, Atlantic, and Gulf of Mexico 
(1991) and U.S. Pacific (1998; https://www.fisheries.noaa.gov/action/recovery-plans-leatherback-sea-turtle).
    The species has the widest distribution of any reptile, with a 
global range extending from 71[deg] N, based on an at-sea capture off 
Norway (Carriol and Vader 2002) to 47[deg] S, based on an at-sea 
sighting off New Zealand (Eggleston 1971; Eckert et al. 2012). The 
species has several thermoregulatory adaptations to allow such a large 
latitudinal range, maintain its core temperature while foraging, and 
avoid overheating during nesting. These include its large size, low 
metabolic rates, countercurrent heat exchange at the base of its limbs, 
and peripheral insulation (Frair et al. 1972; Greer et al. 1973; 
Paladino et al. 1990; Fossette et al. 2009; Bostrom et al. 2010; Eckert 
et al. 2012; Casey et al. 2014; reviewed in Wallace and Jones 2015).
    Nesting is restricted to mainly tropical or subtropical beaches. 
However, nesting also occurs on temperate beaches of the SW Indian 
Ocean (Pritchard and Mortimer 1999). Nesting usually occurs on high-
energy beaches (Pritchard 1976), resulting in high rates of natural 
erosion. The primary factors influencing shoreline suitability for 
nesting appear to be a lack of abrasive substrate material, a deep-
water approach to minimize energy expenditure needed to reach nesting 
sites, and proximity to oceanic currents that can facilitate hatchling 
dispersal (Eckert et al. 2012). Leatherback turtles appear to prefer 
wide, long beaches with a steep slope, deep rock-free sand, and an 
unobstructed deep water or soft-bottom approach (Pritchard and Mortimer 
1999; Eckert et al. 2015). As a result, it has been proposed that the 
choice of nesting location is based on site characteristics within a 
geographic location (MacKay et al. 2014).
    Foraging areas are generally characterized by zones of upwelling, 
including off the edges of continents, where major currents converge, 
and in deep-water eddies (Saba 2013). Important foraging areas include 
but are not limited to: upwelling off the west coasts of North and 
South America (Benson et al. 2011; Roe et al. 2014); Benguela Current 
Marine Ecosystem (Honig et al. 2007); and Canadian waters on the 
Scotian Shelf (James et al. 2005a, 2006b, 2007b).

Abundance

    Adding together the total indices of nesting female abundance for 
all DPSs, the total index of nesting female abundance for the species 
is 32,174 females. This number, however, should be considered as a 
compilation of seven populations ranging in size from 27 to 20,659 
nesting females because nesting female exchange does not occur between 
DPSs.
    Comparisons with historical accounts of nesting female abundance 
are complicated by the discovery of new nesting beaches over time, 
changes in remigration intervals and/or clutch frequency, and modified 
observational effort. Abundance estimates for even large nesting 
beaches were not available prior to 1950 (Rivalan et al. 2006), several 
large nesting beaches were not discovered until the 1960s or later 
(NMFS and USFWS 2013), and monitoring efforts were variable over time. 
Pritchard's 1971 global estimate of 29,000 to 40,000 nesting females 
included a maximum estimate (i.e., 40,000 nesting females) based on the 
assumption that large nesting aggregations had yet to be discovered 
(Pritchard 1971); this estimate did not include large nesting female 
abundances from the East Pacific and SE Atlantic Oceans. At that time, 
the nesting aggregation at Terengganu, Malaysia nesting population was 
thought to be one of the largest; however it has since been extirpated 
(Chan and Liew 1996). In 1982, Pritchard revised his initial global 
estimate to 115,000 nesting females, based largely on the nesting 
beaches in Pacific Mexico (n = 75,000; Pritchard 1982). However, the 
1982 estimate was extrapolated from a brief aerial survey and may have 
been an overestimate (Pritchard 1996). When the Mexico nesting 
population collapsed, Spotila (1996) estimated the total global 
estimate to be 34,500 nesting females, with a range of 26,200 to 42,900 
nesting females. However, this estimate did not include the nesting 
aggregation in Gabon, which in 2002 was identified as

[[Page 48411]]

the largest in the world at that time, with tens of thousands of 
nesting females (Witt et al. 2009). Recent data indicate less than 
9,000 nesting females in Gabon (Formia in progress). Thus, we find that 
leatherback nesting female abundance has declined rapidly in several 
populations. Our total index of nesting female abundance for the 
species, which does include the largest nesting aggregations from all 
DPSs, is lower than previous estimates by at least 10,000 females.
    Species go extinct through the loss of populations. Therefore, the 
loss of any of these populations (which we refer to in this finding 
hypothetically as DPSs) would increase the extinction risk of the 
species. Most of the DPSs exhibit total indices of nesting female 
abundances that place them at risk for environmental variation, genetic 
complications, demographic stochasticity, negative ecological feedback, 
and catastrophes (McElhany et al. 2000; NMFS 2017). The current total 
index of nesting female abundance for the species reflects the impact 
of threats that have affected the species to this point. This reduced 
abundance renders it particularly vulnerable to threats and contributes 
to its extinction risk.

Productivity

    Nest trends are decreasing across the species, except at the least 
abundant nesting aggregation in Brazil (i.e., the SE Atlantic DPS), 
with a total index of 27 nesting females, which is increasing by 4.8 
percent annually. Current nest trends are declining at rates ranging 
from -0.3 percent (within the SW Indian DPS) to -9.3 percent (the 
overall decline for the NW Atlantic DPS). Historical declines are even 
larger. Aerial surveys of nesting beaches in Mexico detected declines 
from over 70,000 nesting females in 1982 to fewer than 250 in 1998, 
with an annual mortality rate of 22.7 percent (Spotila 2000) and an 
overall decline of 97.4 percent in three generations (Wallace et al. 
2013). The Terengganu, Malaysia nesting aggregation has declined by 
17.9 percent annually from 1967 to 2010. It was been reduced to less 
than one percent of its original size between the 1950s and 1995 (Chan 
and Liew 1996) and is now considered functionally extirpated. 
Significant declines in nesting have been documented for other 
populations (Benson et al. 2015). Declining nesting trends reflect the 
impact of threats that have been operating on the species, and these 
trends increase the extinction risk of the species.

Spatial Distribution

    The species occurs over a broad spatial range, in tropical and 
temperate waters worldwide, from 71[deg] N to 47[deg] S (Goff and Lien 
1988; Carriol and Vader 2002; McMahon and Hayes 2006; Shillinger et al. 
2008; Wallace et al. 2010; Benson et al. 2011; Eckert et al. 2012). It 
nests and forages across a wide spatial range, which provides some 
degree of resilience against local impacts to nesting and foraging 
areas. The DPSs are reproductively isolated with little to no gene flow 
connecting them. However, within some DPSs there is fine-scale 
population structure (Dutton et al. 1999; Dutton et al. 2003; Dutton et 
al. 2013; Molfetti et al. 2013). These subpopulations exhibit 
metapopulation dynamics, which make a DPS more resilient to stochastic 
and environmental changes. It is likely that all DPSs once exhibited 
such dynamics, given the ephemeral, high-energy beaches where they nest 
and their regional, but not necessarily beach-specific, philopatry 
(Dutton et al. 1999; Dutton et al. 2013). However, the reduction of 
nesting aggregations within a DPS has likely reduced or removed this 
structure, and the associated resilience, in some DPSs and in the 
overall species.

Diversity

    Relative to other sea turtle species, the leatherback turtle has 
low genetic diversity and shallow mtDNA coalescence (Dutton et al. 
1999), reflecting its recent global radiation, i.e., Post-Pleistocene 
expansion from a refugium in the Indian Ocean (Dutton et al. 1999). As 
a species, it uses diverse and widely distributed nesting and forage 
areas. Differences in size at maturity, remigration rate, clutch 
frequency, and clutch size likely reflect environmental variability 
among DPSs (Saba et al. 2008; Saba et al. 2015). The age of the species 
and its flexible use of multiple foraging and nesting areas indicate 
that the species has some resilience to stochastic and environmental 
changes.

Present or Threatened Destruction, Modification, or Curtailment of 
Habitat or Range

    The destruction or modification of nesting habitat is a threat to 
most leatherback turtles, and in some areas, this threat is major, as a 
result of development, erosion, or obstruction from logs. By the year 
2025, the UN Educational, Scientific and Cultural Organization (2001) 
forecasts that human population growth and migration will result in 75 
percent of people living within 60 km of the sea. This will place 
significant additional pressure on coastal habitats.
    Coastal development and associated activities cause accelerated 
erosion rates and interruption of natural shoreline migration (National 
Research Council 1990). Numerous beaches are eroding due to both 
natural (e.g., storms, sea level changes, waves, shoreline geology) and 
anthropogenic (e.g., development and expansion, construction of 
armoring structures, groins, jetties, marinas, coastal development, 
inlet dredging) factors. Such shoreline erosion has led and will 
continue to lead to a loss of nesting habitat for leatherback turtles 
and potential loss of nests from inundation. Erosion or inundation and 
accretion of sand above incubating nests appear to be the principal 
abiotic factors that negatively affect incubating egg clutches in some 
areas (Dow et al. 2007; USFWS 1999; NMFS and USFWS 2013). Shoreline 
structuring can also physically prevent females from reaching suitable 
nesting habitat or prevent them from returning to sea (Witherington et 
al. 2011).
    Low hatching success, relative to other sea turtle species, is 
characteristic of many leatherback populations despite high fertility 
rates (reviewed by Bell et al. 2003; Eckert et al. 2012). Nest 
relocation is undertaken as a conservation measure in some locations 
when erosion (or poaching and predation) threaten the viability of a 
nest. However, studies have found that hatching success of nests in 
hatcheries or nests relocated to another area of a beach is lower than 
in situ nests (reviewed in Hern[aacute]ndez et al. 2007; Eckert et al. 
2012). In addition, nest relocation results in altered sand 
temperatures, which influences the sex ratio of hatchlings produced 
(Sieg et al. 2011).
    Coastal development and expansion also contributes to habitat 
degradation via artificial lighting (i.e., light pollution). The 
presence of artificial lighting on or adjacent to nesting beaches 
alters the behavior of nesting females (often deterring nesting) and is 
often fatal to post-nesting females and emerging hatchlings, when they 
are attracted to terrestrial light sources and drawn away from the 
water (Witherington 1992; Sella et al. 2006; Witherington et al. 2014). 
As hatchlings head toward lights or meander along the beach, their 
exposure to predators and likelihood of desiccation are greatly 
increased. Artificial lighting may also affect hatchlings that 
successfully find the water, causing them to be misoriented after 
entering the surf zone or while in nearshore waters.

[[Page 48412]]

    The modification of nesting habitat generally results in loss of 
productivity for the species, as a result of reductions in nest and 
hatching success. In addition, several DPSs experience nesting beach 
habitat modifications (e.g., artificial lighting, logs, and other 
obstructions) that result in the death of nesting females and 
hatchlings. Therefore, abundance is also reduced, posing an even 
greater threat to the continued existence of the turtles of the DPS. 
The loss and modification of nesting habitat poses a major threat to 
the species.

Overutilization for Commercial, Recreational, Scientific, or 
Educational Purposes

    Historically, the harvest of turtles and eggs was the primary 
threat to the species, leading to the loss of severe depletion of many 
nesting aggregations worldwide (Spotila et al. 1996). At one point in 
time, egg harvest was ubiquitous with all nests taken at many beaches 
(Chan and Liew 1996; Sarti et al. 2007; reviewed by Eckert et al. 
2012). For the NW Atlantic, NE Indian, and West Pacific DPSs, legal 
harvest of turtle and/or eggs continues. Despite laws in many 
countries, the poaching of eggs continues at most nesting beaches, 
ranging in severity from minor at monitored or protected beaches to 
near 100 percent harvest at unmonitored beaches. Nesting females, and 
turtles caught at sea, continue to be poached for their meat, eggs, and 
fat in many locations (Eckert et al. 2012). As described in detail in 
the prior sections evaluating the status of each individual DPS, the 
harvest of eggs and turtles is a threat to each and to the species 
overall, and for the NE Indian and West Pacific DPSs, it is a primary 
threat. The legal and illegal harvest of turtles and eggs poses a 
threat to the species.

Disease or Predation

    We do not have adequate information on disease to assess its impact 
on the species. However, we have enough information to conclude that 
predation is clearly a threat. Numerous species prey on leatherback 
eggs and hatchlings. Eckert et al. (2012) provide an exhaustive list of 
the documented predators for each life stage and area. For eggs, common 
predators include ants, ghost crabs, monitor lizards, crows, mongoose, 
domestic and feral dogs, and feral pigs (Eckert et al. 2012). For 
hatchlings, common predators include the terrestrial predators listed 
above as well as numerous species of carnivorous fish, including 
sharks. Sharks and killer whales, and in some areas jaguars and 
crocodiles, prey on subadult and adult turtles. Predation on eggs and 
hatchlings is common and reduces productivity of the species; predation 
on subadults and adults is less prevalent but reduces abundance when it 
occurs. Predation is a threat to the species, and for some DPSs, it is 
a major threat.

Inadequacy of Existing Regulatory Mechanisms

    Numerous regulatory mechanisms provide certain protections to sea 
turtles at the international, regional, national, and local levels. For 
example, the harvest of sea turtles and their eggs is prohibited by 
regional conventions and national laws. Fisheries bycatch is also 
addressed, although not comprehensively, by several international and 
national instruments and/or governing bodies. However, as we detail 
below and has been discussed in prior sections reviewing each 
individual DPS, these measures are often poorly implemented or 
enforced, resulting in inadequate protections against the threats they 
are designed to ameliorate.
    In some nations (e.g., South Africa) sea turtles were among the 
first species to receive legal protections and have been the focus of 
concentrated conservation efforts. However, current regulatory 
mechanisms often fall short of preventing further population declines 
and ensuring persistence (Eckert et al. 2012). For many nations the 
regulations in place are inadequate (usually due to lack of enforcement 
and implementation) to address the impacts of a wide range of 
anthropogenic activities that directly injure and kill turtles, disturb 
eggs, disrupt necessary behaviors, and alter terrestrial and marine 
habitats used by the species. In many areas, regulations for the 
harvest of turtles and eggs are inadequate due to a lack of 
enforcement. In some areas, the regulation of fisheries bycatch do not 
adequately reduce associated mortality. Fishery observer coverage is 
often inadequate to accurately estimate leatherback bycatch.
    Due in part to their worldwide distribution and highly migratory 
nature, combined with nesting site fidelity, leatherback turtles 
require international, national, regional, and local protection. Hykle 
(2002) and Tiwari (2002) reviewed the value of some international 
instruments and concluded that they vary in their effectiveness. Often, 
international treaties do not realize their full potential because: 
They do not include all key nations; do not specifically address sea 
turtle conservation; are handicapped by the lack of a sovereign 
authority to promote enforcement; and/or lack of legally-binding 
requirements. Lack of implementation or enforcement by some nations may 
make them less effective than if they were implemented in a more 
consistent manner across the target region. A thorough discussion of 
this topic is available in the 2002 special issue of the Journal of 
International Wildlife Law and Policy: International Instruments and 
Marine Turtle Conservation (Hykle 2002). Additional information on 
national, regional, and local protection is provided in the prior 
sections of this finding relating to each individual DPS.
    In summary, numerous regulatory mechanisms protect leatherback 
turtles, eggs, and nesting habitat throughout the range of the species. 
Although the regulatory mechanisms provide some protection, many do not 
adequately reduce the threat that they were designed to address, 
generally as a result of limited implementation or enforcement. As a 
result, bycatch, incomplete nesting habitat protection, and poaching 
remain threats to the species. We conclude that the inadequacy of the 
regulatory mechanisms is a threat to the leatherback turtle.

Fisheries Bycatch

    Fisheries bycatch is the primary threat to leatherback turtles 
(Crowder 2000; Spotila et al. 2000; Lewison et al. 2004; Wallace et al. 
2011; Wallace et al. 2013; Angel et al. 2014). It is a primary threat 
to all DPSs. Leatherback turtles are susceptible to bycatch in a wide 
range of fisheries, from large scale commercial to artisanal. Gear 
types that affect leatherbacks include: longlines, purse seines, 
driftnets, gillnets, trawls, pots/traps, and pound nets (Gray and Diaz 
2017). Turtles often drown after becoming entangled in nets and other 
gear or become injured and possibly die as a result of hooking or 
interactions with the gear. While bycatch in pelagic shallow-set 
swordfish longline fisheries has received the most attention to date, 
small-scale coastal fisheries occur worldwide, employing over 99 
percent of the world's 51 million fishers (FAO 2011).
    Bycatch data are most commonly collected by trained observers on 
fishing vessels or via surveys or interviews (Lewison et al. 2015). 
Though often the best available data on bycatch, observer data 
generally cover less than five percent of fisheries' total effort 
(Finkbeiner et al. 2011) and are rarely available for small-scale 
fisheries (Wallace et al. 2013; Lewison et al. 2015). The use of 
different metrics also makes the data difficult to compare

[[Page 48413]]

among fisheries, gear types, and regions (Lewison et al. 2015). 
Therefore, estimates of bycatch and resulting mortality often 
underestimate the magnitude of this threat.
    Furthermore, IUU fishing is a significant yet unquantified threat 
to sea turtles worldwide. In addition to killing and injuring turtles, 
it undermines national and regional efforts to estimate fisheries 
bycatch. IUU fishing represents up to 26 million tonnes of fish caught 
annually (http://www.fao.org/iuu-fishing/en/). We have no estimates of 
the impacts to leatherback turtles from IUU fishing, though interaction 
and mortality rates are likely high because of the magnitude of this 
additional fishing pressure and because it is unregulated.
    Generally, leatherback turtles do not attempt to consume the bait 
associated with fishing gear, as other sea turtles do, but become 
entangled in fishing gear (Lewison et al. 2015). Longline fisheries 
involve the deployment of a horizontal main line and vertical 
branchlines with baited hooks, which may entangle leatherback turtles. 
Bycatch reduction measures include using circle hooks, finfish bait, 
minimizing soak times, and limiting mainline length (Angel et al. 2014; 
https://www.fisheries.noaa.gov/national/bycatch/fishing-gear-pelagic-longlines#risks-to-sea-turtles). Purse seines capture schools of fish 
in a vertical wall of netting that can be closed at the bottom (https://www.fisheries.noaa.gov/national/bycatch/fishing-gear-purse-seines); 
bycatch rates are generally much lower than longline bycatch rates 
(Angel et al. 2014). Leatherback turtles also become entangled and 
drowned in drift or set gillnets (https://www.fisheries.noaa.gov/national/bycatch/fishing-gear-gillnets). Gillnets can be devastating to 
leatherback populations when set near nesting beaches and represent the 
primary threat to leatherback turtles in some areas (e.g., Trinidad; 
Eckert and Eckert 2005). Trawl fisheries drag nets along the substrate 
or through the water column and can capture and drown sea turtles. 
Although TEDs may mitigate this threat, they are not always required or 
used in all areas. Vertical lines extending and/or connecting pot and 
trap gear with surface buoys commonly entangle and can kill leatherback 
turtles.
    Longline and net fisheries are often the greatest threats to 
leatherback turtles. In a global study of sea turtle bycatch, Wallace 
et al. (2013) compiled data (n = 239 records) published between 1990 
and 2011 to compare gear types (longline, net, and trawl) and their 
impacts to leatherback RMUs, which are similar to the DPSs discussed in 
this rule, though their exact boundaries differ. Wallace et al. (2013) 
defined high bycatch impact as follows: A weighted median bycatch per 
unit effort (BPUE) greater than or equal to one; median mortality rate 
greater than or equal to 0.5; and affecting adult or subadult turtles. 
They found that longline bycatch had a high impact on SW Atlantic, SE 
Atlantic, and SW Indian RMUs and that net bycatch had a high impact on 
the NW Atlantic and East Pacific RMUs (Wallace et al. 2013).
    Integrating catch data from over 40 nations and bycatch data from 
13 international observer programs, Lewison et al. (2004) estimated the 
numbers of leatherback turtles taken by pelagic longliners to be more 
than 50,000 leatherback turtles in just one year (2000). With over half 
of the total fishing effort (targeting tuna and swordfish) occurring in 
the Pacific Ocean, an estimated 20,000 leatherback turtles interacted 
with longline fishing gear, with 1,000 to 3,200 mortalities in 2000 
(Lewison et al. 2004). However, Beverly and Chapman (2007) estimated 
sea turtle longline bycatch mortality to be approximately 20 percent of 
that estimated by Lewison et al. (2004), or approximately 200 to 640 
leatherback turtle mortalities annually. We consider the estimate of 
Beverly and Chapman (2007) to be more realistic, considering the low 
nesting females abundance of Pacific leatherback turtles, and because 
Beverly and Chapman (2007) combined the effort data from Lewison et al. 
(2004) with bycatch data from Molony (2005) that differentiated between 
deep-set and shallow-set fisheries (which have different interaction 
rates).
    In the Pacific Ocean, Roe et al. (2014) predicted leatherback 
turtle bycatch hotspots by comparing the satellite tracks of 135 adult 
turtles with longline fishing effort. The greatest bycatch risk 
occurred adjacent to primary nesting beaches of the West Pacific DPS. 
Bycatch risk was also high in the South Pacific Gyre, where the East 
Pacific DPS forages. Expanding on this study, a study of observer data 
from 34 swordfish-targeting shallow-set longline fleets found there 
were 331 leatherback turtle interactions between 1989 and 2015 (Clarke 
2017). Clarke (2017) identified two bycatch hotspot areas: Central 
North Pacific Ocean and eastern Australia (Clarke 2017).
    In the Atlantic Ocean, Fossette et al. (2014) compared leatherback 
telemetry data to longline fishing effort data from ICCAT to identify 
nine areas in which leatherback turtles are exposed to bycatch 
associated with high longline fishery pressure. The high pressure 
fishing areas include foraging areas in the North and South Atlantic 
Ocean and in waters off Brazil and western Africa. These high pressure 
fishing areas are not comparable to those identified by Roe et al. 
(2014), who used a different methodology, but both studies identify 
high risk areas within each ocean basin.
    Additional bycatch information that we have set out in prior 
sections specific to each DPS applies to our consideration of the risk 
to the overall species. In summary, fisheries bycatch is a threat that 
is encountered by numerous juvenile and adult leatherback turtles. 
Mortality rates are often high, and individuals that are released may 
experience injuries or sublethal effects associated with entanglement, 
submergence, or handling. Fisheries bycatch reduces abundance, and when 
it prevents nesting females from returning to nesting beaches, reduces 
productivity as well. Fisheries bycatch is the primary threat to the 
leatherback species.

Vessel Strikes

    Vessel strikes pose a threat to the species throughout its range. 
As mature individuals move from oceanic foraging areas into coastal 
waters to reproduce, they are exposed to a greater concentration of 
vessels. Vessel strikes off nesting beaches may injure or kill these 
individuals, reducing the abundance and productivity of the DPS. Most 
vessel strikes likely go unnoticed or unreported, making this threat 
potentially much more significant that documented occurrences would 
suggest. Vessel strikes are a threat to the leatherback species.

Pollution

    We define pollution as including contaminants, marine debris, and 
ghost or derelict fishing gear. Such interactions are likely to go 
unnoticed and unreported and thus likely present a more significant 
impact than documented occurrences would suggest. Leatherback turtles 
of all life stages are vulnerable to oil spills, on land and at sea, 
where exposure to oil and dispersants occurs via contact (i.e., 
physical fouling), inhalation, or ingestion (reviewed by Stacy et al. 
in press).
    Marine debris is ubiquitous throughout the range of the species. 
Marine debris includes plastics (including plastic bags), 
microplastics, derelict fishing gear (e.g., ghost nets and other 
discarded or lost gear), and other man-made materials. Leatherback 
turtles may directly consume floating plastics, mistaking it for their 
gelatinous prey or accidentally ingest plastics while foraging. In 
particular, plastic bags appear similar to jellyfish in the marine

[[Page 48414]]

environment, inappropriately triggering the sensory cue to feed 
(Schuyler et al. 2014; Nelms et al. 2016). Plastic bags have been found 
during necropsy of stranded leatherback turtles, and phthalates derived 
from plastics have been found in leatherback egg yolk (Lebreton et al. 
2018). Mrosovsky et al. (2009) reviewed 408 necropsy records from 1885 
to 2007 and found evidence of plastic in the gastrointestinal tract of 
34 percent of leatherback turtles, including some cases in which the 
plastic obstructed the passage of food through the gut. The most 
commonly identified items were plastic bags, fishing lines, twine, and 
fragments of mylar balloons. Ghost or derelict fishing gear include 
discarded or lost nets, line, and other gear. Ghost fishing gear can 
drift in the ocean and fish unattended for decades and kill numerous 
individuals (Wilcox et al. 2013). The main sources of ghost fishing 
gear are gillnet, purse seine, and trawl fisheries (Stelfox et al. 
2016). Marine debris affects leatherback turtles via ingestion or 
entanglement and can reduce food intake and digestive capacity, cause 
distress and/or drowning, expose turtles to contaminants, and in some 
cases cause direct mortality (Mrosovsky et al. 2009; NMFS and USFWS 
2013). In terms of microplastics, all samples analyzed from all species 
(including leatherbacks) had microplastics evident in their gastro-
intestinal tracts (Duncan et al. 2018). Given the increase of pollution 
entering the marine environment over the past 30 years or approximately 
5.2 to 19.3 million tonnes per year (Lebreton et al. 2018), we conclude 
that pollution is a threat to the species.

Natural Disasters and Oceanographic Regime Shifts

    Leatherback turtles are susceptible to the impacts of natural 
disasters and oceanographic regime shifts as a result of their nesting 
and foraging preferences. Nesting usually occurs on high-energy beaches 
that are inherently unstable (Pritchard 1976) and which are susceptible 
to natural erosion. The primary factors influencing shoreline 
suitability for nesting appear to be a lack of abrasive substrate 
material, a deep-water approach to minimize energy expenditure needed 
to reach nesting sites, and proximity to oceanic currents that can 
facilitate hatchling dispersal (Eckert et al. 2012). Leatherback 
turtles nest lower on the beach than other species, exposing their 
nests to erosion and inundation. Storm events, King Tides, tsunamis, 
and hurricanes can destroy or modify preferred nesting beaches of some 
DPSs.
    Gelatinous prey have relatively low energy content, requiring 
leatherback turtles to consume large quantities to meet metabolic 
demands (Heaslip et al. 2012; Jones et al. 2012). Leatherback turtles 
likely maximize their caloric intake by aligning their foraging 
behavior to prey distribution abundance. Foraging areas are generally 
characterized by zones of upwelling, including off the edges of 
continents, where major currents converge, and in deep-water eddies 
(Saba 2013). Some of these areas experience oceanographic regime shifts 
that alter water temperature, downwelling, Ekman upwelling, sea surface 
height, chlorophyll-a concentration, and mesoscale eddies (Bailey et 
al. 2013; Benson et al. 2011). These shifts alter prey availability, 
and thus productivity parameters (e.g., remigration rates, clutch size, 
and clutch frequency), for leatherback turtles. Some DPSs are not 
affected by such shifts because they have access to diverse foraging 
areas, such as: coastal and pelagic waters; subtropical, temperate, and 
boreal waters; and ephemeral eddies (Neeman et al. 2015). Such 
flexibility allows the leatherback turtle to consume large amounts of 
prey at various locations throughout the year.
    We conclude that natural disasters and oceanographic regime shifts 
are threats to the species, affecting some but not all populations, 
depending on the location of nesting and foraging areas. These threats 
reduce productivity by reducing nesting, nesting habitat, and nest and 
hatching success.

Climate Change

    Climate change is a threat that affects leatherback turtles of all 
life stages and within all DPSs. A warming climate and rising sea 
levels can impact leatherback turtles through changes in beach 
morphology, increased sand temperatures leading to a greater incidence 
of lethal incubation temperatures, changes in hatchling sex ratios, and 
the loss of nests or nesting habitat due to beach erosion (Benson et 
al. 2013).
    Impacts from climate change, especially due to global warming, are 
already being observed and are likely to become more apparent in future 
years (IPCC 2007a). In its Fifth Assessment Report, the IPCC (2014) 
stated that the globally averaged combined land and ocean surface 
temperature data has shown a warming of 0.85 [deg]C from 1880 to 2012. 
The mean rate of globally averaged sea level rise was 1.7 millimeters 
annually between 1901 and 2010, 2.0 millimeters annually between 1971 
and 2010, and 3.2 millimeters annually between 1993 and 2010. Climate 
model projections exhibit a wide range of plausible scenarios for both 
temperature and precipitation over the next several decades. The global 
mean surface temperature change for the period 2016 to 2035 relative to 
1986 to 2005 will likely be in the range of 0.3 [deg] to 0.7 [deg]C 
(medium confidence; IPCC 2014). The global ocean temperature will 
continue to warm, and increases in seasonal and annual mean surface 
temperatures are expected to be larger in the tropics and Northern 
Hemisphere subtropics (i.e., where leatherback turtles nest; IPCC 
2014). Under Representative Concentration Pathway 8.5, the change in 
global mean sea level rise for the mid- and late 21st century relative 
to the reference period of 1986 to 2005 is projected to be 0.30 meters 
higher from 2046 to 2065 and 0.63 meters higher from 2081 to 2100, with 
a rate of sea level rise during 2081 to 2100 of 8 to 16 millimeters 
annually (medium confidence; IPCC 2014).
    For all sea turtles, including leatherback turtles, a warming 
climate and rising sea levels are likely to result in changes in beach 
morphology, increased sand temperatures leading to a greater incidence 
of lethal incubation temperatures, changes in hatchling sex ratios, and 
the loss of nests and nesting habitat due to beach erosion (Benson et 
al. 2015; Hamann et al. 2013). Leatherback turtles are most likely to 
be affected by climate change at nesting beaches due to warming 
temperatures, sea level rise, and storm events and due to oceanic 
changes that are likely to alter foraging and migration. Warming 
temperatures and increased precipitation at nesting beaches affect 
reproductive output including hatching success, hatchling emergence 
rate, and hatchling sex ratios (e.g., Hawkes et al. 2009). Sea level 
rise results in a reduction or shift in available nesting beach 
habitat, an increased risk of erosion and nest inundation (e.g., Boyes 
et al. 2010), and reduced nest success (Fish et al. 2005; Fuentes et 
al. 2010; Fonseca et al. 2013). Increased frequency and severity of 
storm events impact nests and nesting habitat, thus reducing nesting 
and hatching success (e.g., Van Houtan and Bass 2007; Fuentes and Abbs 
2010). Changes in productivity affect the abundance and distribution of 
forage species, resulting in changes in the foraging behavior and 
distribution of leatherback turtles (e.g., Saba et al. 2008, 2012) as 
well as changes in leatherback fitness and growth. Changes in water 
temperature lead to a shift in range and changes in phenology (timing 
of nesting seasons,

[[Page 48415]]

timing of migrations) and different threat exposure (e.g., Saba et al. 
2008, 2012).
    Increasing sand temperatures will alter the thermal regime of 
incubating nests, resulting in altered sex ratios and reduced hatching 
output (Hawkes et al. 2009). Leatherback turtles exhibit temperature-
dependent sex determination (reviewed by Binckley and Spotila 2015), 
whereby phenotypic sex is determined by temperatures experienced during 
the thermosensitive period of egg incubation. A 1:1 sex ratio is 
produced when this pivotal temperature lies between 29.2 and 30.4 
[deg]C for leatherback turtles in Malaysia, 29.2 and 29.8 [deg]C in 
French Guiana/Suriname, and 29.2 and 29.5 [deg]C in Pacific Costa Rica 
(Binckley and Spotila 2015). Warmer temperatures produce more female 
embryos (Mrosovsky et al. 1984; Hawkes et al. 2007), but temperatures 
over 32 [deg]C are likely to result in death. As temperatures continue 
to increase, emergence rates decrease (Santidri[aacute]n Tomillo et al. 
2015), removing any advantage of increased female production. 
Santidri[aacute]n Tomillo et al. (2015) conclude that leatherback 
turtles may not survive if temperatures rise as projected by current 
climate change models. Increases in precipitation might temporarily 
reduce the temperatures at some nesting beaches thereby mitigating some 
impacts relative to increasing sand temperatures.
    Beach erosion and nest inundation already threaten leatherback 
nesting habitat globally. Sea level rise is likely to increase the 
number of nests lost to erosion and inundation. Such loss of nests is 
especially problematic in areas prone to storm events, which are likely 
to increase in intensity and duration, and in areas where coastal 
development impedes natural shoreline migration.
    Climate change is also likely to alter the productivity in some 
marine environments, which could affect leatherback prey availability. 
With reports on the increasing incidence of jellyfish blooms in some 
locations, there is the perception that jellyfish abundance is 
increasing globally (Condon et al. 2012), which could result in more 
prey for leatherback turtles (Hawkes et al. 2009). However, after 
analyzing all available long-term datasets on jellyfish abundance, 
Condon et al. (2012) found that there is no robust evidence for a 
global increase in jellyfish. Rather, jellyfish populations undergo 
larger, worldwide oscillations with an approximate 20-year periodicity 
(Condon et al. 2012). Additional monitoring is needed to determine 
whether the weak linear trend in jellyfish abundance since 1970 
represents an actual increase or is a phase of an oscillation (Condon 
et al. 2012). Therefore, the effects of climate change on productivity 
are uncertain.
    As described in prior sections with respect to each individual 
population, some impacts from climate change have already been 
observed. At several nesting beaches, increased erosion occurs, and sex 
ratios are severely skewed toward females. Beach erosion reduces 
productivity. Although the skew toward females could increase 
productivity in the short-term, it is often correlated with low 
hatching success. For these reasons, climate change is a threat to the 
species.

Conservation Efforts

    The ESA requires the Services to make their listing determinations 
solely on the basis of the best scientific and commercial data 
available, after conducting a status review, and after taking into 
account those efforts, if any, being made by any State or foreign 
nation to protect the species, whether by predatory control, protection 
of habitat and food supply, or other conservation practices, within any 
area under its jurisdiction, or on the high seas (16 U.S.C. 1533 
(b)(1)(A)). In addition, the Services published a policy for the 
evaluation of domestic conservation efforts which have yet to be 
implemented or to show effectiveness (68 FR 15100; March 28, 2003). We 
did not identify any conservation efforts that required such evaluation 
for leatherbacks (i.e., the conservation efforts reviewed are 
international in nature or have already been implemented to a 
sufficient degree that they have a track record of being effective or 
not being effective). Several conservation efforts have been previously 
discussed in prior sections evaluating regulatory mechanisms with 
respect to each DPS. Therefore, the list below describes only those 
conservation efforts that have not been previously discussed and that 
apply generally to the leatherback species rather than being clearly 
associated with a particular population. We considered these efforts 
prior to making our listing determination. After reviewing these 
efforts, we concluded that they have been somewhat effective, in that 
they have prevented this endangered species from going extinct. 
However, these efforts have not reduced the threats to a level at which 
protections under the ESA are no longer necessary.
    African Convention on the Conservation of Nature and Natural 
Resources (Algiers Convention): Adopted in September 1968, the 
contracted states were ``to undertake to adopt the measures necessary 
to ensure conservation, utilization and development of soil, water, 
floral and faunal resources in accordance with scientific principles 
and with due regard to the best interests of the people.'' The Algiers 
Convention recently has undergone revision, and its objectives are to 
enhance environmental protection, foster conservation and sustainable 
use of natural resources, and harmonize and coordinate policies in 
these fields with a view to achieving ecologically rational, 
economically sound, and socially acceptable development policies and 
programs. Additional information is available at http://www.unep.ch/regionalseas/legal/afr.htm.
    Atlantic Sea Turtle Network (ASO): Created in 2003 to foster 
greater collaboration in southern Brazil, Uruguay, and Argentina for 
the protection of sea turtles and their habitats. ASO represents dozens 
of local and regional NGOs and government agencies as well as hundreds 
of community members. ASO and its partners have significantly advanced 
policies to protect sea turtles from fisheries interactions, which is 
one of the most severe threats in the region. Brazil plays a major role 
in South American (and global) sea turtle conservation and research, 
and it serves as an example to other countries. Projeto TAMAR, a 
partnership of the Centro TAMAR/ICMBio, government agencies, and 
Fundac[atilde]o Pr[oacute] TAMAR, has been active since 1980. Today, 
the group carries out sea turtle research and conservation from 22 
stations on the coast and the offshore islands of Brazil. Another NGO 
based in the southern Brazilian state of Rio Grande do Sul, called NEMA 
has been collecting systematic sea turtle stranding data since 1990. 
Those data have been instrumental to conservation efforts in Brazil and 
have shown that southern Brazil has the highest stranding rates for 
loggerheads in the western Atlantic Ocean.
    Association of Southeast Asian Nations (The ASEAN) Ministers on 
Agriculture and Forestry (AMAF): A Memorandum of Understanding (MoU) on 
ASEAN sea turtle conservation was created in 1999. From this, a Sea 
Turtle Conservation and Protection Program and Work plan has developed; 
research and monitoring activities have also been produced regionally 
(Kadir 2000). The objectives of this Memorandum of Understanding, 
initiated by ASEAN, are to promote the protection, conservation, 
replenishing, and recovery of sea turtles and their habitats based on 
the best

[[Page 48416]]

available scientific evidence, taking into account the environmental, 
socio-economic and cultural characteristics of the Parties. It 
currently has nine signatory states in the South East Asian Region 
(http://document.seafdec.or.th/projects/2012/seaturtles.php).
    Andaman and Nicobar Island Environmental Team (ANET): A division of 
the Centre for Herpetology/Madras Crocodile Bank Trust has been 
conducting surveys and monitoring since 1991. Over the last few years, 
conservation and monitoring of sea turtles in these islands has been 
carried by Dakshin Foundation and Indian Institute of Science in 
collaboration with ANET, centered around a leatherback monitoring 
program on Little Andaman Island. A multi- institution stakeholder 
platform for marine conservation, including government and non- 
governmental agencies, was established by these groups to facilitate 
the conservation of marine turtles and other endangered species 
(Tripathy et al. 2012). The Trust, along with the Wildlife Institute of 
India and Ministry of Environment and Forests, produced a series of 
manuals on sea turtle conservation, management and research to help 
forest officers, conservationists, NGOs and wildlife enthusiasts 
conduct sea turtle conservation and research programs (ANET, 2003 as 
cited in Shanker and Andrews 2004). A consolidated manual has been 
produced to achieve these goals by Dakshin Foundation and the Trust 
(Tripathy et al. 2012).
    Central American Regional Network: This collaborative effort 
created the national sea turtle network in each country of the region, 
as well as the development of first hand tools, such as a regional 
diagnosis, a 10-year strategic plan, a manual of best practices, and 
four regional training and information workshops for people in the 
region (e.g., Chac[oacute]n and Arauz, 2001). This initiative is 
managed by stakeholders in various sectors (private, non-governmental 
and governmental) across the region.
    Convention on the Conservation of Migratory Species of Wild Animals 
(CMS): This Convention, also known as the Bonn Convention or CMS, is an 
international treaty that focuses on the conservation of migratory 
species and their habitats. As of December 2018, the Convention had 127 
Parties, including Parties from Africa, Central and South America, 
Asia, Europe, and Oceania. While the Convention has successfully 
brought together about half the countries of the world with a direct 
interest in sea turtles, it has yet to realize its full potential 
(Hykle 2002). Its membership does not include a number of key 
countries, including Canada, China, Indonesia, Japan, Mexico, Oman, and 
the United States. Under the CMS, two Memoranda of Understanding (MOUs) 
apply to leatherback turtles: The MOU concerning Conservation Measures 
for Marine Turtles of the Atlantic Coast of Africa and the MOU on the 
Conservation and Management of Marine Turtles and their Habitats of the 
Indian Ocean and South-East Asia. Additional information is available 
at http://www.cms.int.
    Convention on Biological Diversity (CBD): The primary objectives of 
this international treaty are: (1) The conservation of biological 
diversity, (2) the sustainable use of its components, and (3) the fair 
and equitable sharing of the benefits arising out of the utilization of 
genetic resources. This Convention has been in force since 1993 and had 
193 Parties as of March 2013. While the Convention provides a framework 
within which are broad conservation objectives, it does not 
specifically address sea turtle conservation (Hykle 2002). Additional 
information is available at http://www.cbd.int.
    Convention on International Trade in Endangered Species of Wild 
Fauna and Flora (CITES): Known as CITES, this Convention was designed 
to regulate international trade in a wide range of wild animals and 
plants. CITES was implemented in 1975 and currently has 183 Parties. 
Although CITES has been effective at minimizing the international trade 
of sea turtle products, it does not limit legal harvest within 
countries, nor does it regulate intra-country commerce of sea turtle 
products (Hykle, 2002). The leatherback turtle is included (since 1977) 
in CITES Appendix I, which bans trade, including individuals and 
products, except as permitted for exceptional circumstances, not to 
include commercial purposes (Lyster 1985). Additional information is 
available at http://www.cites.org.
    Convention on the Conservation of European Wildlife and Natural 
Habitats: Also known as the Bern Convention, the goals of this 
instrument are to conserve wild flora and fauna and their natural 
habitats, especially those species and habitats whose conservation 
requires the cooperation of several States, and to promote such 
cooperation. The Convention was enacted in 1982 and currently includes 
51 European and African States and the European Union. Additional 
information is available at http://www.coe.int/t/dg4/cultureheritage/nature/bern/default_en.asp.
    Convention for the Co-operation in the Protection and Development 
of the Marine and Coastal Environment of the West and Central African 
Region (Abidjan Convention): The Abidjan Convention covers the marine 
environment, coastal zones, and related inland waters from Mauritania 
to Namibia. The Abidjan Convention countries are Angola, Benin, 
Cameroon, Cape Verde, Congo, Cote d'Ivoire, Democratic Republic of 
Congo, Equatorial Guinea, Gabon, Gambia, Ghana, Guinea, Guinea-Bissau, 
Liberia, Mauritania, Namibia, Nigeria, Sao Tome and Principe, Senegal, 
Sierra Leone, and Togo. The Abidjan Convention is an agreement for the 
protection and management of the marine and coastal areas that 
highlights sources of pollution, including pollution from ships, 
dumping, land-based sources, exploration and exploitation of the sea-
bed, and pollution from or through the atmosphere. The Convention also 
identifies where co-operative environmental management efforts are 
needed. These areas of concern include coastal erosion, specially 
protected areas, combating pollution in cases of emergency and 
environmental impact assessment.
    Convention for the Protection Management and Development of the 
Marine and Coastal Environment of the Eastern African Region (Nairobi 
Convention): The Nairobi Convention was signed in 1985 and came into 
force in 1996. This instrument ``provides a mechanism for regional 
cooperation, coordination and collaborative actions, and enables the 
Contracting Parties to harness resources and expertise from a wide 
range of stakeholders and interest groups towards solving interlinked 
problems of the coastal and marine environment.'' Parties are 
responsible for ``the conservation and wise management of the sea 
turtle populations frequenting their waters and shores [and] agree to 
work closely together to improve the conservation status of the sea 
turtles and the habitats upon which they depend.'' The Western Indian 
Ocean-Marine Turtle Task Force, which was created under the Nairobi 
Convention and the IOSEA, plays a role in sea turtle conservation. This 
is a technical, non-political working group comprised of specialists 
from eleven countries: Comoros, France (La R[eacute]union), Kenya, 
Madagascar, Mauritius, Mozambique, Seychelles, Somalia, South Africa, 
United Kingdom and Tanzania, as well as representatives from inter-
governmental organizations, academic, and non-governmental 
organizations within the region. Additional information is available at 
http://www.unep.org/NairobiConvention.

[[Page 48417]]

    Convention for the Protection of the Marine Environment of the 
North-East Atlantic: Also called the OSPAR Convention, this 1992 
instrument combines and updates the 1972 Oslo Convention against 
dumping waste in the marine environment and the 1974 Paris Convention 
addressing marine pollution stemming from land-based sources. The 
convention is managed by the OSPAR Commission, which is comprised of 
representatives from 15 signatory nations (Belgium, Denmark, Finland, 
France, Germany, Iceland, Ireland, Luxembourg, The Netherlands, Norway, 
Portugal, Spain, Sweden, Switzerland, and United Kingdom), as well as 
the European Commission, representing the European Community. The 
mission of the OSPAR Convention ``. . . is to conserve marine 
ecosystems and safeguard human health in the North-East Atlantic by 
preventing and eliminating pollution; by protecting the marine 
environment from the adverse effects of human activities; and by 
contributing to the sustainable use of the seas.'' Leatherback turtles 
are included on the OSPAR List of Threatened and/or Declining Species 
and Habitats, used by the OSPAR Commission for setting priorities for 
work on the conservation and protection of marine biodiversity. 
Additional information is available at http://www.ospar.org.
    Convention for the Protection and Development of the Marine 
Environment of the Wider Caribbean Region: Also called the Cartagena 
Convention, this instrument that benefits turtles of the Northwest 
Atlantic leatherback DPS, has been in place since 1986 and currently 
has 38 member states and territories. Under this Convention, the 
component that relates to leatherback turtles is the Protocol 
Concerning Specially Protected Areas and Wildlife (SPAW) that has been 
in place since 2000. The goals are to encourage Parties ``to take all 
appropriate measures to protect and preserve rare or fragile 
ecosystems, as well as the habitat of depleted, threatened or 
endangered species, in the Convention area.'' The SPAW protocol has 
partnered with WIDECAST to develop a program of work on sea turtle 
conservation, which has helped many of the Caribbean nations to 
identify and prioritize their conservation actions through Sea Turtle 
Recovery Action Plans. Each recovery action plan summarizes the known 
distribution of sea turtles, discusses major causes of mortality, 
evaluates the effectiveness of existing conservation laws, and 
prioritizes implementing measures for stock recovery. The objective of 
the recovery action plan series is not only to assist Caribbean 
governments in the discharge of their obligations under the SPAW 
Protocol, but also to promote a regional capability to implement 
science-based sea turtle management and conservation programs. 
Additional information is available at http://www.cep.unep.org/about-cep/spaw.
    Convention on Nature Protection and Wildlife Preservation in the 
Western Hemisphere (Washington or Western Hemisphere Convention): 
Elements of the Convention include the protection of species from 
human-induced extinction, the establishment of protected areas, the 
regulation of international trade in wildlife, special measures for 
migratory birds and stressing the need for co-operation in scientific 
research and other fields are all elements of wildlife conservation. 
Additional information is available at http://www.oas.org/juridico/english/treaties/c-8.html.
    Convention for the Protection of the Marine Environment and Coastal 
Area of the South-East Pacific (Lima Convention): This Convention's 
signatories include all countries along the Pacific Rim of South 
America from Panama to Chile. Among other resource management 
components, this Convention established protocol for the conservation 
and management of protected marine resources. Stemming from this 
Convention is the Commision Permanente del Pacifico Sur (CPPS) that has 
developed a Marine Turtle Action Plan for the Southeast Pacific that 
outlines a strategy for protecting and recovering marine turtles in 
this region. Convention for the Protection of the Natural Resources and 
Environment of the South Pacific Region (Noumea Convention): This 
Convention has been in force since 1990 and currently includes 26 
Parties. The purpose of the Convention is to protect the marine 
environment and coastal zones of the South-East Pacific within the 200-
mile area of maritime sovereignty and jurisdiction of the Parties and, 
beyond that area, the high seas up to a distance within which pollution 
of the high seas may affect that area. Additional information is 
available at http://www.unep.org/regionalseas/programmes/nonunep/pacific/instruments/default.asp.
    Convention Concerning the Protection of the World Cultural and 
Natural Heritage (World Heritage Convention): The World Heritage 
Convention was signed in 1972 and, as of November 2007, 185 states were 
parties to the Convention. The instrument requires parties to take 
effective and active measures to protect and conserve habitat of 
threatened species of animals and plants of scientific or aesthetic 
value. The World Heritage Convention currently includes 31 marine 
sites. Additional information is available at http://whc.unesco.org/en/conventiontext.
    Convention for the Conservation and Management of Highly Migratory 
Fish Stocks in the Western and Central Pacific Ocean (WCPF Convention): 
The convention entered into force on 19 June 2004. The WCPF Convention 
draws on many of the provisions of the UN Fish Stocks Agreement [UNFSA] 
while, at the same time, reflecting the special political, socio-
economic, geographical and environmental characteristics of the western 
and central Pacific Ocean (WCPO) region. The WCPFC Convention seeks to 
address problems in the management of high seas fisheries resulting 
from unregulated fishing, over-capitalization, excessive fleet 
capacity, vessel re-flagging to escape controls, insufficiently 
selective gear, unreliable databases and insufficient multilateral 
cooperation in respect to conservation and management of highly 
migratory fish stocks.
    Convention for the Prohibition of Fishing with Long Driftnets in 
the South Pacific: This regional convention, also known as the 
Wellington Convention, was adopted in 1989 in Wellington, New Zealand, 
and entered into force in 1991. The objective of the Convention is ``to 
restrict and prohibit the use of drift nets in the South Pacific region 
in order to conserve marine living resources.'' Additional information 
is available at http://www.mfat.govt.nz/Treaties-and-International-Law/01-Treaties-for-which-NZ-is-Depositary/0-Prohibition-of-Fishing.php.
    Eastern Pacific Leatherback Network: Also known as La Red de la 
Tortuga La[uacute]d del Oc[eacute]ano Pacifico (La[uacute]d OPO) 
(www.savepacificleatherbacks.org) was formed to address the critical 
need for regional coordination of East Pacific leatherback conservation 
actions necessary to track conservation priorities and progress at the 
population level. Led by Fauna & Flora International, this network has 
brought together conservationists, researchers, practitioners and 
government representatives from 22 institutions across nine East 
Pacific countries with varying priorities, capacities and historical 
experiences in leatherback research and conservation to contribute to 
shared activities, projects, and goals. Through these efforts, 
La[uacute]d now has mutually-agreed upon mechanisms for sharing 
information and data, as well as

[[Page 48418]]

standardized protocols for nesting beach monitoring and bycatch 
assessments/fishing practices.
    The Eastern Tropical Pacific Marine Corridor (CMAR) is a regional 
and cross-border initiative for the conservation and sustainable use of 
the region's marine and coastal resources. Its objective is to 
sustainably manage biodiversity through ecosystem based management and 
the development of regional intergovernmental strategies with support 
of non-governmental organizations and international cooperation 
agencies.
    United Nations' Food and Agricultural Organization (FAO) Technical 
Consultation on Sea Turtle-Fishery Interactions: While not a true 
international instrument for conservation, the 2004 FAO of the UN's 
technical consultation on sea turtle-fishery interactions was 
groundbreaking in that it solidified the commitment of the lead UN 
agency for fisheries to reduce sea turtle bycatch in marine fisheries 
operations. Recommendations from the technical consultation were 
endorsed by the FAO Committee on Fisheries (COFI) and called for the 
immediate implementation by member nations and Regional Fishery 
Management Organizations (RFMOs) of guidelines to reduce sea turtle 
mortality in fishing operations, developed as part of the technical 
consultation. Currently, all five of the tuna RFMOs call on their 
members and cooperating non-members to adhere to the 2009 FAO 
``Guidelines to Reduce Sea Turtle Mortality in Fishing Operations,'' 
which describes all the gear types sea turtles could interact with and 
the latest mitigation options. The Western and Central Pacific 
Fisheries Commission (http://www.wcpfc.int) has the most protective 
measures (CMM 2008-03), which follow the FAO guidelines and ensure safe 
handling of all captured sea turtles. Fisheries deploying purse seines, 
to the extent practicable, must avoid encircling sea turtles and 
release entangled turtles from fish aggregating devices. Longline 
fishermen must carry line cutters and use dehookers to release sea 
turtles caught on a line. Longliners must either use large circle 
hooks, whole finfish bait, or mitigation measures approved by the 
Scientific Committee and the Technical and Compliance Committee.
    Inter-American Tropical Tuna Convention (IATTC) has enacted a 
resolution to mitigate the impact of tuna fishing vessels on sea 
turtles by reducing bycatch, injury, and mortality of sea turtles. The 
IATTC has also developed a memorandum of understanding with the IAC. 
For more information, see http://www.iattc.org/PDFFiles/Resolutions/IATTC/_English/C-07-03-Active_Sea%20turtles.pdf.
    The International Commission for the Conservation of Atlantic Tunas 
(ICCAT) has adopted a resolution for the reduction of sea turtle 
mortality (Resolution 03-11), encouraging States to submit data on sea 
turtle interactions, release sea turtles alive wherever possible, and 
conduct research on mitigation measures. It calls for implementing the 
FAO Guidelines for sea turtles, avoiding encirclement of sea turtles by 
purse seiners, safely handling and releasing sea turtles, and reporting 
on interactions. The Commission does not have any specific gear 
requirements applicable to longline fisheries. ICCAT is currently 
undertaking an ecological risk assessment to better understand the 
impact of its fisheries on sea turtle populations. For more information 
see http://www.iattc.org/. Other international fisheries organizations 
that may influence leatherback turtle recovery include the Southeast 
Atlantic Fisheries Organization (http://www.seafo.org) and the North 
Atlantic Fisheries Organization (http://nafo.int). These organizations 
regulate trawl fisheries in their respective Convention areas. Given 
that sea turtles can be incidentally captured in these fisheries, both 
organizations have sea turtle resolutions calling on their Parties to 
implement the FAO Guidelines on sea turtles as well as to report data 
on sea turtle interactions.
    The Indian Ocean Tuna Commission (IOTC) is playing an increased 
role in turtle conservation. Resolution 05/08, superseded by Resolution 
09/06 on Sea Turtles, sets out reporting requirements related to 
interactions with sea turtles and accordingly provides an executive 
summary per species for adoption at the Working Party on Ecosystem and 
By-catch and then subsequently at the Scientific Committee. In 2011, 
IOTC developed a ``Sea Turtle Identification Card'' to be distributed 
to all longliners operating in the Indian Ocean (www.iotc.com). In 
2012, the Indian Ocean Tuna Commission (IOTC) began requiring its 31 
contracting Parties to report sea turtle bycatch and to use safe 
handling and release techniques for sea turtles on longline vessels.
    Indian Ocean--South-East Asian Marine Turtle Memorandum of 
Understanding (IOSEA): Under the auspices of the Convention of 
Migratory Species, the IOSEA memorandum of understanding provides a 
mechanism for States of the Indian Ocean and South-East Asian region, 
as well as other concerned States, to work together to conserve and 
replenish depleted marine turtle populations. This collaboration is 
achieved through the collective implementation of an associated 
Conservation and Management Plan. Currently, there are 33 Signatory 
States. The United States became a signatory in 2001. The IOSEA has an 
active sub-regional group for the Western Indian Ocean, which has 
improved collaboration amongst sea turtle conservationists in the 
region. Further, the IOSEA website provides reference materials, 
satellite tracks, on-line reporting of compliance with the Convention, 
and information on all international mechanisms currently in place for 
the conservation of sea turtles. Finally, at the 2012 Sixth Signatory 
of States meeting in Bangkok, Thailand, the Signatory States agreed to 
procedures to establish a network of sites of importance for sea 
turtles in the IOSEA region (http://www.ioseaturtles.org).
    Inter-American Convention for the Protection and Conservation of 
Sea Turtles (IAC): This Convention is the only legally binding 
international treaty dedicated exclusively to sea turtles and sets 
standards for the conservation of these endangered animals and their 
habitats with a large emphasis on bycatch reduction. The Convention 
area is the Pacific and the Atlantic waters of the Americas. Currently, 
there are 15 Parties. The United States became a Party in 1999. The IAC 
has worked to adopt fisheries bycatch resolutions, carried out 
workshops on Caribbean sea turtle conservation, and established 
collaboration with other agreements such as the Convention for the 
Protection and Development of the Marine Environment of the Wider 
Caribbean Region and the International Commission for the Conservation 
of Atlantic Tunas. Additional information is available at http://www.iacseaturtle.org.
    International Convention for the Prevention of Pollution from Ships 
(MARPOL): The MARPOL Convention is a combination of two treaties 
adopted in 1973 and 1978 to prevent pollution of the marine environment 
by ships from operational or accidental causes. The 1973 treaty covered 
pollution by oil, chemicals, and harmful substances in packaged form, 
sewage and garbage. The 1978 MARPOL Protocol was adopted at a 
Conference on Tanker Safety and Pollution Prevention which included 
standards for tanker design and operation. The 1978 Protocol 
incorporated the 1973 Convention as it had not yet been in force and is 
known as the International Convention for the Prevention of Marine 
Pollution from Ships, 1973, as modified by the Protocol

[[Page 48419]]

of 1978 relating thereto (MARPOL 73/78). The 1978 Convention went into 
force in 1983 (Annexes I and II). The Convention includes regulations 
aimed at preventing and minimizing accidental and routine operations 
pollution from ships. Amendments passed since have updated the 
convention.
    International Union for Conservation of Nature (IUCN): The IUCN 
Species Programme assesses the conservation status of species on a 
global scale. This assessment provides objective, scientific 
information on the current status of threatened species. The IUCN Red 
List of Threatened Species provides taxonomic, conservation status and 
distribution information on plants and animals that have been globally 
evaluated using the IUCN Red List Categories and Criteria. This system 
is designed to determine the relative risk of extinction, and the main 
purpose of the IUCN Red List is to catalogue and highlight those plants 
and animals that are facing a higher risk of global extinction (i.e., 
those listed as Critically Endangered, Endangered and Vulnerable). 
Additional information is available at http://www.iucnRed List.org/about.
    Marine Turtle Conservation Act (MTCA): The MTCA is a key element of 
sea turtle protection in the United States and internationally. This 
Act authorizes a dedicated fund to support marine turtle conservation 
projects in foreign countries, with emphasis on protecting nesting 
populations and nesting habitat. Additional information is available at 
https://www.fws.gov/international/wildlife-without-borders/marine-turtle-conservation-fund.html.
    Memorandum of Agreement between the Government of the Republic of 
the Philippines and the Government of Malaysia on the Establishment of 
the Turtle Island Heritage Protected Area: Through a bilateral 
agreement, the Governments of the Philippines and Malaysia established 
The Turtle Island Heritage Protected Area (TIHPA), made up of nine 
islands (6 in the Philippines and 3 in Malaysia). The following 
priority activities were identified: management-oriented research, the 
establishment of a centralized database and information network, 
appropriate information awareness programs, a marine turtle resource 
management and protection program, and an appropriate ecotourism 
program (Bache and Frazier 2006).
    Memorandum of Understanding of a Tri-National Partnership between 
the Government of the Republic of Indonesia, the Independent State of 
Papua New Guinea and the Government of Solomon Islands: This agreement 
promotes the conservation and management of Western Pacific leatherback 
turtles at nesting sites, feeding areas and migratory routes in 
Indonesia, Papua New Guinea and Solomon Islands. This is done through 
the systematic exchange of information and data on research, population 
and migratory routes monitoring, nesting sites and feeding areas 
management activities for Western Pacific leatherback turtles and by 
enhancing public awareness of the importance of conserving these 
turtles and their critical habitats. http://awsassets.wwf.or.id/downloads/mou_trinationalpartneshipagreement_clean.pdf.
    Memorandum of Understanding Concerning Conservation Measures for 
Marine Turtles of the Atlantic Coast of Africa (Abidjan Memorandum): 
This MOU was concluded under the auspices of the Convention on the 
Conservation of Migratory Species of Wild Animals (CMS) and became 
effective in 1999. The MOU area covers 26 Range States along the 
Atlantic coast of Africa extending approximately 14,000 km from Morocco 
to South Africa. The goal of this MOU is to improve the conservation 
status of marine turtles along the Atlantic Coast of Africa. It aims at 
safeguarding six marine turtle species--including the leatherback 
turtle--that are estimated to have rapidly declined in numbers during 
recent years due to excessive exploitation (both direct and incidental) 
and the degradation of essential habitats. This includes the protection 
of the life stages from hatchlings through adults with particular 
attention paid to the impacts of fishery bycatch and the need to 
include local communities in the development and implementation of 
conservation activities. However, despite this agreement, killing of 
adult turtles and harvesting of eggs remains rampant in many areas 
along the Atlantic African coast. Additional information is available 
at http://www.cms.int/species/africa_turtle/AFRICAturtle_bkgd.htm.
    National Sea Turtle Conservation Project in India: Launched in 1998 
with the aim of protecting Lepidochelys olivacea, but it also has 
conservation and protection strategies for all the other turtle species 
nesting in the country. This project was undertaken by the Indian 
government to oversee: Surveys, monitoring programs, fisheries 
interactions, community and NGO participation, awareness raising and 
education, research support, and other support for regional and 
international co-operation and collaboration for sea turtles 
conservation (Choudhury et al., 2001).
    North American Agreement for Environmental Cooperation: As mandated 
by the 1994 North American Agreement for Environmental Cooperation, the 
Commission for Environmental Cooperation (CEC) encourages Canada, the 
United States, and Mexico to adopt a continental approach to the 
conservation of flora and fauna. In 2003, this mandate was strengthened 
as the three North American countries launched the Strategic Plan for 
North American Cooperation in the Conservation of Biodiversity. The 
North American Conservation Action Plan (NACAP) initiative began as an 
effort promoted by the three countries, through the CEC, to facilitate 
the conservation of marine and terrestrial species of common concern. 
In 2005, the CEC supported the development of a NACAP for Pacific 
leatherbacks by Canada, the United States, and Mexico. Identified 
actions in the plan addressed three main objectives: (1) protection and 
management of nesting beaches and females; (2) mortality reduction from 
bycatch throughout the Pacific Basin; and (3) waste management, control 
of pollution, and disposal of debris at sea.
    Ramsar Convention on Wetlands: The Convention on Wetlands, signed 
in Ramsar, Iran, in 1971, is an intergovernmental treaty, which 
provides the framework for national action and international 
cooperation for the conservation and wise use of wetlands and their 
resources. Currently, there are 158 parties to the convention, with 
1,752 wetland sites, including important marine turtle habitat. 
Additional information is available at http://www.ramsar.org.
    Secretariat of the Pacific Regional Environment Programme (SPREP): 
SPREP's turtle conservation program seeks to improve knowledge about 
sea turtles in the Pacific through an active tagging program, as well 
as maintaining a database to collate information about sea turtle tags 
in the Pacific. SPREP supports capacity building throughout the central 
and southwest Pacific. SPREP established an action plan for the Pacific 
Islands (http://www.sprep.org/).
    South-East Atlantic Fisheries Organization (SEAFO): SEAFO manages 
fisheries activities in the Southeast Atlantic high seas area, 
excluding tunas and billfish. SEAFO adopted Resolution 01/06, ``to 
Reduce Sea Turtle Mortality in Fishing Operations,'' in 2006. The 
Resolution requires Members to: (1) Implement the FAO Guidelines; and 
(2) establish on-board observer programs to collect information on sea 
turtle

[[Page 48420]]

interactions in SEAFO-managed fisheries. This Resolution is not legally 
binding. Additional information is available at http://www.seafo.org.
    South Atlantic Association: In the southwest Atlantic, the South 
Atlantic Association is a multinational group that includes 
representatives from Brazil, Uruguay, and Argentina and meets bi-
annually to share information and develop regional action plans to 
address threats including bycatch (http://www.tortugasaso.org/). At the 
national level, Brazil has developed a national plan for sea turtle 
bycatch reduction that was initiated in 2001 (Marcovaldi et al. 2002). 
This national plan includes various activities to mitigate bycatch, 
including time-area restrictions of fisheries, use of bycatch reduction 
devices, and working with fishermen to successfully release live-
captured turtles. In Uruguay, all sea turtles are protected from human 
impacts, including fisheries bycatch, by presidential decree (Decreto 
Presidencial 144/98).
    United Nations Convention on the Law of the Sea (UNCLOS): To date, 
155 countries, including most mainland countries lining the western 
Pacific, and the European Community have joined in the convention. The 
United States has signed the treaty and abides by some provisions, but 
the Senate has not ratified it. Aside from its provisions defining 
ocean boundaries, the convention establishes general obligations for 
safeguarding the marine environment through mandating sustainable 
fishing practices and protecting freedom of scientific research on the 
high seas. Additional information is available at http://www.un.org/Depts/los/index.htm.
    United Nations' Food and Agricultural Organization (FAO): The FAO 
published guidelines for sea turtle protection, entitled Technical 
Consultation on Sea Turtle-Fishery Interactions (FAO 2005). The UN 1995 
Code of Conduct for Responsible Fisheries (FAO 2004) provides 
guidelines for the development and implementation of national fisheries 
policies, including gear modification (e.g., circle hooks, fish bait, 
deeper sets, and reduced soak time), new technologies, and management 
of areas where fishery and sea turtle interactions are more severe. The 
guidelines stress the need for mitigation measures, data on all 
fisheries, fishing industry involvement, and education for fishers, 
observers, managers, and compliance officers (FAO 2004).
    United Nations Resolution 44/225 on Large-Scale Pelagic Driftnet 
Fishing: In 1989, the UN called, in a unanimous resolution, for the 
elimination of all high seas driftnets by 1992. Additional information 
is available at http://www.un.org/documents/ga/res/44/a44r225.htm.
    Although numerous conservation efforts apply to the species, they 
do not adequately reduce its risk of extinction for the reasons 
discussed previously.

Extinction Risk Analysis

    The best available information is consistent with the species' 
current ``endangered'' listing. The species exhibits a global total 
index of nesting female abundance of 32,060 females at monitored 
beaches. This number is lower than historical estimates of nesting 
female abundance (n = 115,000, Pritchard 1982; and n = 34,500, Spotila 
1996), which did not include the large, but then unknown, Gabon nesting 
aggregation. Limited nesting female abundance is a major source of 
concern for most DPSs, whose small population sizes place them in 
danger of stochastic or catastrophic events that increase extinction 
risk. The limited nesting female abundance increases the extinction 
risk of the species.
    The species also exhibits declining nesting trends for all but one 
of the DPSs. With the exception of the DPS with the smallest index of 
nesting female abundance (i.e., SW Atlantic DPS, with 27 nesting 
females), the DPSs are declining at rates of 0.3 to 9.3 percent 
annually. Even low levels of decline are a threat for DPSs with limited 
nesting female abundance, and nesting declines of approximately nine 
percent (i.e., NW and SE Atlantic DPSs) are unsustainable. Total 
declines of 97 and 99 percent have occurred within the East Pacific and 
NE Indian DPSs, respectively, since nesting was first identified and 
quantified for these populations. The declining trends in nesting 
increase the extinction risk of the species.
    The species exhibits broad nesting and foraging ranges. However, 
metapopulation dynamics have likely been reduced, with reductions in 
abundance and the loss of some nesting aggregations. The species also 
demonstrates little genetic diversity, relative to other sea turtle 
species. Although the species demonstrates some resilience to threats, 
overall we find it to be at risk of extinction, due to limited 
abundance and declining nesting trends, which reflect the cumulative 
impacts of threats that have acted on the species in the past (and in 
many cases continue to act on the species).
    Current threats continue to place the species in danger of 
extinction. The primary threat to the species is bycatch in commercial 
and artisanal, pelagic and coastal, fisheries. Fisheries bycatch 
reduces abundance by removing individuals from the population. Because 
several fisheries operate near nesting beaches, productivity is also 
reduced when nesting females are prevented from returning to nesting 
beaches. The harvest of eggs and turtles is also a major threat to the 
species. Illegal poaching occurs throughout the range of the species, 
and harvest is legal but poorly documented in some nations. The loss 
and modification of nesting habitat is another major threat, reducing 
productivity and, in some instances, abundance, when nesting females 
die as a result of artificial lighting or obstructions preventing them 
from returning to sea. Predation results in the loss of eggs and 
hatchlings, reducing productivity of the species. Additional threats 
that occur throughout the range of the species include vessel strikes, 
pollution, marine debris, oil and gas exploration, and climate change. 
Natural disasters and oceanographic regime shifts are threats in some 
areas. Though many regulatory mechanisms are in place, they do not 
adequately reduce the impact of these threats.
    Based on our review of the best available scientific and commercial 
data, we find nothing that is inconsistent with the leatherback 
species' current listing as an endangered species. In sum, the best 
available information is consistent with the current listing status of 
the leatherback sea turtle as an endangered species throughout its 
range. The threatened species definition does not apply because the 
species is currently in danger of extinction (i.e., at present), rather 
than on a trajectory to become so within the foreseeable future.

Final Determination

    The Services determined that the best available scientific and 
commercial information would support recognizing seven populations as 
DPSs (including the NW Atlantic) because they meet the discreteness and 
significance criteria for DPSs. However, we found that--even were they 
to be recognized and listed separately--all DPSs meet the definition of 
an endangered species because they are in danger of extinction 
throughout all of their ranges. The leatherback turtle is currently 
listed throughout its range as an endangered species. Replacing this 
listing with seven endangered DPSs would not be consistent with 
Congressional guidance to use the authority to list DPSs ``sparingly'' 
while encouraging the conservation of genetic diversity (see Senate 
Report 151, 96th

[[Page 48421]]

Congress, 1st Session). Such guidance clearly indicates that the 
Services have some discretion to determine whether or not to recognize 
DPSs that would require disaggregating an existing listing even where 
those populations can be shown to meet the discreteness and 
significance tests of the DPS Policy.
    After determining that all seven populations would have the same 
status as the overall species, we next considered whether there was any 
reason to nevertheless replace the global (range-wide) listing with 
individual listings for the seven DPSs. We conclude that disaggregating 
the global listing is not warranted. It would be inconsistent with 
Congressional guidance and run counter to the conservation purposes of 
the Act to disaggregate the current listing into DPSs, because those 
DPSs would have the same listing status as the whole currently. 
Disaggregating this listing would bring about significant complications 
and possible public confusion without any meaningful corresponding 
conservation benefit. Replacing the range-wide listing with seven DPSs 
having the same status would not provide leatherback turtles with an 
overriding conservation benefit, as all members are currently protected 
to the fullest extent under the ESA as an endangered species. Section 7 
consultations already consider the effects of an action on individuals 
and populations to determine whether a Federal agency has insured that 
its action is not likely to jeopardize the continued existence of the 
species. Even if the species were disaggregated into DPSs, this change 
would not be expected to result in different substantive outcomes in 
consultations. In addition, focused conservation efforts have been, and 
will continue to be, applied at scales smaller than the species-level. 
For example, FWS' Marine Turtle Conservation Fund provides funding to 
partners in foreign nations to protect leatherback turtles and their 
nesting habitats; projects include efforts to monitor and protect 
leatherback turtles in Indonesia and Gabon (https://www.fws.gov/international/wildlife-without-borders/marine-turtle-conservation-fund.html). Similarly, Pacific leatherback turtles are highlighted 
under NMFS' Species in the Spotlight: Survive to Thrive initiative, 
which directs attention and resources to highly-at-risk species 
(https://www.fisheries.noaa.gov/topic/endangered-species-conservation#species-in-the-spotlight).
    For these reasons, the Services have determined that replacing the 
existing global listing with separate listings for individual DPSs is 
not warranted. Although the best available data indicates that the 
populations meet the criteria for significance and discreteness, we 
find that it would not further the purposes of the Act to recognize and 
list seven DPSs separately as endangered under the ESA. The current 
global listing of the species remains in effect.
    We conclude that the petitioned actions, to identify the NW 
Atlantic population as a DPS and list it as a threatened species under 
the ESA, are not warranted. This is a final action, and, therefore, we 
are not soliciting public comments.

Peer Review

    In December 2004, the Office of Management and Budget (OMB) issued 
a Final Information Quality Bulletin for Peer Review, establishing 
minimum peer review standards, a transparent process for public 
disclosure of peer review planning, and opportunities for public 
participation. The OMB Bulletin, implemented under the Information 
Quality Act (Pub. L. 106-554), is intended to enhance the quality and 
credibility of the Federal government's scientific information and 
applies to influential or highly influential scientific information 
disseminated on or after June 16, 2005. To satisfy our requirements 
under the OMB Bulletin, we obtained independent peer review of the 
Status Review Report by independent scientists with expertise in 
leatherback turtle biology, endangered species listing policy, and 
related fields. All peer reviewer comments were addressed prior to the 
publication of the Status Review Report and this finding.

References Cited

    A complete list of references is available upon request to the NMFS 
Office of Protected Resources (see ADDRESSES).

Authority

    The authority for this action is the Endangered Species Act of 
1973, as amended (16 U.S.C. 1531 et seq.).

Samuel D. Rauch III,
Deputy Assistant Administrator for Regulatory Programs, National Marine 
Fisheries Service.
Aurelia Skipwith,
Director, U.S. Fish and Wildlife Service.
[FR Doc. 2020-16277 Filed 8-7-20; 8:45 am]
BILLING CODE 3510-22-P