# L OGGERHEAD 

# S EA 

# T URTLE 

# (C ARETTA CARETTA 

# )

2009 S TATUS 

# R EVIEW


# U NDER THE 

# U.S. E NDANGERED 

# S PECIES 

# A CT

L OGGERHEAD 

## B IOLOGICAL 

## R EVIEW 

## T EAM



### Therese A. Conant, Peter H. Dutton, To moharu Eguchi, Sheryan P. Epperly, Christina C. Fahy, Matthew H. Godfrey, Sa ndra L. MacPherson, Ea rl E. Possardt, Barbara A. Schroeder, Jeffrey A. Seminoff, Melissa L. Snover, Carrie M. Upite, and Blair E. Witherington

August 2009iii 

## LOGGERHEAD BIOLOGICAL REVIEW TEAM

Therese A. Conant

National Marine Fisheries Service

Office of Protected Resources

Peter H. Dutton

National Marine Fisheries Service

Southwest Fisheries Science Center

Tomoharu Eguchi

National Marine Fisheries Service

Southwest Fisheries Science Center

Sheryan P. Epperly

National Marine Fisheries Service

Southeast Fisheries Science Center

Christina C. Fahy

National Marine Fisheries Service Southwest Regional Office

Matthew H. Godfrey

North Carolina Wildlife Resources Commission

Sandra L. MacPherson

U.S. Fish and Wildlife Service Earl E. Possardt

U.S. Fish and Wildlife Service

Barbara A. Schroeder

National Marine Fisheries Service

Office of Protected Resources

Jeffrey A. Seminoff

National Marine Fisheries Service

Southwest Fisheries Science Center

Melissa L. Snover

National Marine Fisheries Service

Pacific Islands Fisheries Science Center

Carrie M. Upite

National Marine Fisheries Service Northeast Regional Office

Blair E. Witherington

Florida Fish and Wildlife Conservation

Commission

Fish and Wildlife Research Instituteiv 

## ACKNOWLEDGEMENTS

The Biological Review Team gratefully acknow ledges the key input and assistance of the following individuals who served as peer reviewer s of this document: Dr. Alan Bolten (Archie Carr Center for Sea Turtle Research, University of Florida), Dr. Paolo Casale (MTSG-IUCN Italy), Dr. Nancy FitzSimmons (University of Canberra), Dr. Selina Heppell (Oregon State University), Dr. Yoshimasa Matsuzawa (Sea Tur tle Association of Japan), Dr. Jeffrey Moore (Duke University Marine Laboratory), Dr. Ronel Nel (Nelson Mandela Metropolitan University), Dr. Kartik Shanker (Indian Institute of Scien ce), and Dr. Manjula Tiwa ri (National Marine Fisheries Service contractor). We also thank the individuals who provided unpublished

information for use in the document.

This document should be cited as follows:

Conant, T.A., P.H. Dutton, T. Eguchi, S.P. Eppe rly, C.C. Fahy, M.H. Godfrey, S.L. MacPherson, E.E. Possardt, B.A. Schroeder, J.A. Seminoff, M.L. Snover, C.M. Upite, and B.E. Witherington. 2009. Loggerhead sea turtle (Caretta caretta) 2009 status review under the U.S. Endangered Species Act. Report of the Loggerhead Biol ogical Review Team to the National Marine Fisheries Service, August 2009. 222 pages.v 

## EXECUTIVE SUMMARY

The loggerhead sea turtle ( Caretta caretta ) was listed as threatened on July 28, 1978. The threatened listing was applied to wherever th e species occurs. In 2007, the National Marine Fisheries Service (NMFS) and the U.S. Fish and Wildlife Service (FWS) completed a 5-year review for the loggerhead. A 5-year review is an assessment of a listed species to determine whether its status has changed since the time of its listing such that it should be delisted or classified differently than its current status. The agencies concluded that new information available since the completion of the previous 5-year review in 1995 indicated a possible separation of populations by ocean basins but that a more in-depth analysis was needed to determine the application of the distinct population segment (DPS) policy. Based on the new information and the need for further analysis under the DPS policy, NMFS and FWS recommended that no change in listing status was warranted. However, they committed to fully assemble and analyze all relevant info rmation in accordance with the DPS policy.

On July 16, 2007, NMFS and FWS received a petiti on from the Center for Biological Diversity and Turtle Island Restoration Network requesting th at loggerhead turtles in the North Pacific be

reclassified as a DPS with e ndangered status and that criti cal habitat be designated. On November 16, 2007, NMFS and FWS received a pe tition from the Center for Biological Diversity and Oceana requesting that loggerhead turtles in the Northwes t Atlantic Ocean be

reclassified as a DPS with enda ngered status and that critical habitat be designated. NMFS and FWS determined that the July 16, 2007, Nort h Pacific petition and the November 16, 2007, Northwest Atlantic petition both presented substa ntial information that the petitioned actions may be warranted. In the published 90-day petition findings, NMFS and FWS committed to assess the loggerhead listing stat us on a global basis, whic h is consistent with the recommendation in the 2007 5-year review.

NMFS and FWS convened a biological review team (BRT) in February 2008 to review the best available scientific information, determine whether DPSs exist, and assess the extinction risk for each potential DPS. The BRT organized thei r evaluation by ocean basin: Pacific Ocean, Atlantic Ocean (including the Mediterranean Sea), and Indian Ocean. This approach was

consistent with the 2007 5-year review for the loggerhead and was chosen to facilitate data assemb ly and evaluation. It was not meant to preclude identification of DPSs on a broader or finer scale.

The NMFS and FWS 1996 DPS policy defines a population to be a “distinct population segment” if it is both discrete and significant relative to its taxon. Th e BRT evaluated genetic evidence, tagging (flipper and PIT tags) and satell ite telemetry data, demographics information, oceanographic features, and geographic barriers, a nd determined that there are at least nine

discrete population segments for loggerhead sea tur tles globally. These discrete population segments are markedly separated from each othe r as a consequence of ecological, behavioral, and oceanographic factors, and based on genetic evidence. Therefore, the BRT unanimously concluded that two discrete population segments ex ist in the Pacific Ocea n, three in the Indian Ocean, and four in the Atlantic Ocean/Mediterranean Sea.vi The BRT then considered whether each of the ni ne identified discrete population segments is significant relative to its taxon. The BRT determined that each of the nine discrete population segments were biologically and ecologically signif icant. They each represent a large portion of the species range, sometimes encompassing an en tire hemispheric ocean basin. The range of each discrete population segment represents a uni que ecosystem, influen ced by local ecological and physical factors. The loss of any single discrete population segment would result in a significant gap in the loggerhead’s range. E ach discrete population segment is genetically unique; the loss of any one discrete population segment would represen t a significant loss of genetic diversity. Therefore, the BRT concluded that the nine identified population segments are

both discrete from other conspecific population segments and significant to the species to which they belong, Caretta caretta

.

The BRT has identified the following nine loggerhead DPSs distributed globally: (1) North Pacific Ocean DPS, (2) South Pacific Ocean DPS, (3) North Indian Ocean DPS, (4) Southeast Indo-Pacific Ocean DPS, (5) Southwest Indian Ocean DPS, (6) Northwest Atlantic Ocean DPS, (7) Northeast Atlantic Ocean DPS, (8) Mediterranean Sea DPS, and (9) South Atlantic Ocean DPS.

The BRT next conducted two independent analyses to assess extinction ri sks of the identified DPSs. The first analysis used the diffusion a pproximation approach based on a time series of counts of nesting females or nests to provide a me tric, susceptibility to quasi-extinction (SQE). SQE is an increasing function of quasi-extinction threshold (QET). The second analysis focused on determining the effects of known anthropogeni c mortality on each potential DPS with respect to the vital rates of the species. This approach focused on how additional mortalities may affect the future growth and status of each DPS.

For three of five DPSs with sufficient data to conduct the SQE analysis (Northwest Atlantic Ocean, South Pacific Ocean, and North Pacific O cean), the threshold of SQE = 0.3 was reached at QET < 0.3, indicating high likelihood of quasi-ext inction over a wide ra nge of QET values. There were not enough data to c onduct the SQE analysis for the North Indian Ocean, Southeast Indo-Pacific Ocean, Northeast Atlantic Ocean, and Mediterranean Sea DPSs.

According to the analysis using experts’ opini ons in the matrix model framework, all loggerhead turtle DPSs have the potential to decline in the future. Although so me DPSs are indicating increasing trends at nesting beaches (Southwes t Indian Ocean and South Atlantic Ocean),

available information about anthr opogenic threats to juveniles and adults in neritic and oceanic environments indicate possible unsustainable addi tional mortalities. According to the threat matrix analysis, the potential for future decline is greatest for the North Indian Ocean, Northwest Atlantic Ocean, Northeast Atlantic Ocean, Medite rranean Sea, and South Atlantic Ocean DPSs.vii 

## TABLE OF CONTENTS

LOGGERHEAD BIOLOGICAL REVIEW TEAM

................................................................ iii

ACKNOWLEDGEME NTS

........................................................................................................ iv

EXE CUTIVE SUMMARY .......................................................................................................... v TABLE OF CONTENTS

........................................................................................................... vii

LIST OF ACRONYM S AND ABBREVIATIONS .................................................................... 1

SECTION 1—INTRO DUCTION AND BACKGROUND ........................................................ 3

1.1. ESA Overview ................................................................................................................... 3

1.1.1. Purpose ........................................................................................................................... 3

1.1.2. Definitions ...................................................................................................................... 3

1.1.3. Listing ............................................................................................................................. 3

1.1.4. Distinct Population Segment ........................................................................................ 4

1.2. History of ESA Listing, Stat us Review s, and Petitions .................................................. 4

1.2.1. ESA Listing .................................................................................................................... 4

1.2.2. ESA 5-year Review s ...................................................................................................... 4

1.2.3. Petitions .......................................................................................................................... 5

1.3. Approach to the Status Review ........................................................................................ 6

SECTION 2—SPECIE S OVERVIEW ....................................................................................... 7

2.1. Taxonomy .......................................................................................................................... 7

2.2. Physical Appearance ......................................................................................................... 7

2.3. Distribution and Habitat .................................................................................................. 8

2.4. Biological Characteristics ............................................................................................... 13

SECTION 3—DET ERMINATION OF DPS ........................................................................... 16

3.1. Overview of Information Used to Determine DPS ....................................................... 16

3.1.1. Discretenes s Determination ........................................................................................ 17

3.1.2. Significan ce Determination ........................................................................................ 30

3.2. Significan t Portion of it s Range (SPOIR) Assessment ................................................. 35

SECTION 4—ASSES SMENT OF EXTINCTION RISK ....................................................... 36

4.1. Description of Extinction Risk Assessment Approaches ............................................. 36

4.2. Computation of Susceptibility to Quasi-extinction (SQE) .......................................... 36

4.2.1. Methods ........................................................................................................................ 36

4.2.2. Count Data ................................................................................................................... 38

4.2.2.1. North Pacific Ocean DPS ....................................................................................... 40

4.2.2.2. South Pacific Ocean DPS ........................................................................................ 41

4.2.2.3. North Indian Ocean DPS ........................................................................................ 42

4.2.2.4. Southeast Indo-Pacific Ocean DPS ........................................................................ 42

4.2.2.5. Southw est Indian Ocean DPS ................................................................................ 42

4.2.2.6. Northw est Atlantic Ocean DPS .............................................................................. 43

4.2.2.7. Northeast Atlantic Oc ean DPS .............................................................................. 43

4.2.2.8. Mediterran ean Sea DPS ......................................................................................... 43

4.2.2.9. South Atlantic Ocean DPS ..................................................................................... 43

4.2.3. Results .......................................................................................................................... 46

4.3. Threat Matrix Analysis .................................................................................................. 53

4.3.1. General M odeling Approach ..................................................................................... 53

4.3.2. Survival Rates and Population Grow th Rate for the Base Models ........................ 53viii 4.3.3. Quantifying Known Threats ...................................................................................... 56

4.3.4. Combining Additional Mortalitie s and the Base Population Model ...................... 58

4.4. Results .............................................................................................................................. 59

4.4.1. Results—Base Population Model ............................................................................... 59

4.4.2. Results—Threat Matrix ............................................................................................. 67

4.4.2.1. North Pacific Ocean DPS ....................................................................................... 68

4.4.2.2. South Pacific Ocean DPS ........................................................................................ 69

4.4.2.3. North Indian Ocean DPS ........................................................................................ 70

4.4.2.4. Southeast Indo-Pacific Ocean DPS ........................................................................ 71

4.4.2.5. Southw est Indian Ocean DPS ................................................................................ 72

4.4.2.6. Northw est Atlantic Ocean DPS .............................................................................. 73

4.4.2.7. Northeast Atlantic Oc ean DPS .............................................................................. 74

4.4.2.8. Mediterran ean Sea DPS ......................................................................................... 75

4.4.2.9. South Atlantic Ocean DPS ..................................................................................... 76

4.4.3. DPS Status ................................................................................................................... 77

4.5. Synthesis ........................................................................................................................... 83

4.6. Supplemental Material—Sensitivity Analyses ............................................................. 85

4.6.1. Effects of AFR and CV of AFR on the Base Population Model ............................. 85

4.6.2. Effects o f AFR on the Threat Matrix Analysis ........................................................ 86

4.6.3. Proportion of Neritic Juvenile Stage Duration ........................................................ 86

SECTION 5—ANALYSIS OF FACTORS LI STED UNDER ES A SECTION 4(a)(1) ...... 100

5.1. General Description of Threats ................................................................................... 100

5.2. Threats Specific to DP S ................................................................................................ 110

5.2.1. North Pacific Ocean DPS ......................................................................................... 110

5.2.2. South Pacific Ocean DPS .......................................................................................... 118

5.2.3. North Indian Ocean DPS .......................................................................................... 122

5.2.4. Southeast Indo-Pacific Ocean DPS .......................................................................... 124

5.2.5 Southw est Indian Ocean DPS .................................................................................. 127

5.2.6. Northw est Atlantic Ocean DPS ................................................................................ 129

5.2.7. Northeast Atlantic Oc ean DPS ................................................................................ 138

5.2.8. Mediterran ean Sea DPS ........................................................................................... 145

5.2.9. South Atlantic Ocean DPS ....................................................................................... 157

SECTION 6—CONCL USIONS—SYNTHESIS & INTEGRATION .................................. 161

6.1. North Pacific Ocean DPS ............................................................................................. 161

6.2. South Pacific Ocean DPS .............................................................................................. 161

6.3. North Indian Ocean DPS .............................................................................................. 162

6.4. Southeast Indo-Pacific Ocean DPS .............................................................................. 162

6.5. Southw est Indian Ocean DPS ...................................................................................... 163

6.6. Northw est Atlantic Ocean DPS .................................................................................... 164

6.7. Northeast Atlantic Oc ean DPS .................................................................................... 164

6.8. Mediterran ean Sea DPS ............................................................................................... 165

6.9. South Atlantic Ocean DPS ........................................................................................... 165

SECTION 7—REFE RENCES ................................................................................................. 1671 

## LIST OF ACRONYMS AND ABBREVIATIONS

The following are standard abbreviations fo r acronyms and terms found throughout this

document:

AFR Age at First Reproduction

ASEAN Association of South East Asian Nations BRT Biological Review Team CBD Convention on Biological Diversity CCL Curved Carapace Length

CI Confidence Interval

CITES Convention on the Internati onal Trade in Endangered Specie s of Wild Fauna and Flora

CMS Convention on the Conservation of Migratory Species of Wild Animals

COFI FAO Committee on Fisheries

CPPS Commision Permanente del Pacifico Sur

CV Coefficient of Variation DNA Deoxyribonucleic Acid DPS Distinct Population Segment

ENSO El Niño Southern Oscillation ESA Endangered Species Act of 1973, as amended FAO Food and Agriculture Organi zation of the United Nations

FWS U.S. Fish and Wildlife Service

GBR Great Barrier Reef GCRU Greater Caribbean Recovery Unit

GIWA Global International Waters Assessment

IAC Inter-American Convention for the Cons ervation and Protection of Marine Turtles

ICCAT International Commission for th e Conservation of Atlantic Tunas

IOSEA Indian Ocean – South-East Asian Ma rine Turtle Memora ndum of Understanding IPCC Intergovernmental Panel on Climate Change LME Large Marine Ecosystem

MARPOL International Convention for th e Prevention of Pollution from Ships MOU Memorandum of Understanding MSA U.S. Magnuson-Stevens Fishery Co nservation and Management Act mtDNA Mitochondrial Deoxyribonucleic Acid NBSD Negative Binomial Stage Distribution NGMRU Northern Gulf of Mexico Recovery Unit

NMFS National Marine Fisheries Service NPF Northern Australian Prawn Fishery NRU Northern Recovery Unit PIT Passive Integrated Transponder PFRU Peninsular Florida Recovery Unit PVA Population Viability Analysis QET Quasi-extinction Threshold RFMO Regional Fishery Mana gement Organizations SCL Straight Carapace Length

SD Standard Deviation2 SEAFO South-East Atlantic Fisheries Organization SPAW Protocol Concerning Specially Protected Areas and Wildlife SPOIR Significant Portion of its Range SPREP South Pacific Regiona l Environment Programme SQE Susceptibility to Quasi-extinction STAJ Sea Turtle Association of Japan

TED Turtle Excluder Device

UNCLOS United Nations Convention on the Law of the Sea UNEP United Nations Environment Programme3 

## SECTION 1—INTRODUCTION AND BACKGROUND

1.1. ESA Overview

1.1.1. Purpose

The purpose of the Endangered Species Act of 1973, as am ended (ESA) (16 U.S.C. 1531 et seq .) is to provide a means to conserve ecosyste ms upon which endangered and threatened species depend, to provide a program for the conservation of endangered and threatened species, and to take appropriate steps to recover endangered and threatened species. The National Marine Fisheries Service (NMFS) and U.S. Fish and Wi ldlife Service (FWS) share responsibility for administering the ESA. NMFS and FWS are re sponsible for determining whether species, subspecies, or distinct populati on segments of vertebrate specie s are threatened or endangered under the ESA. FWS typically has the lead fo r terrestrial and freshw ater species, and NMFS typically has the lead for marine, estuarine, and anadromous species. NMFS and FWS share jurisdiction for recovering sea tu rtles – NMFS is responsible for sea turtles in their marine environment and FWS is responsible for s ea turtles in their terrestrial habitat.

1.1.2. Definitions

Species : includes any subspecies of fish or wild life or plants and any distinct population

segm ent of any species of vertebrate fish or wildlife which interbreeds when mature.

Endangered Species : any species which is in danger of extinction throughout all or a significant portion of its range.

Threatened Species : any species which is likely to beco me an endangered species within the foreseeable future throughout all or a significant portion of its range.

1.1.3. Listing

Section 4 of the ESA specifies a process upon which a species m ay be determined to be listed as threatened or endangered with extinction, change d in status, or removed from the list. The determination is based solely on the best available scientific and commercial data after reviewing the status of the species and taking into account conservation e fforts. NMFS and FWS must determine whether any species is an endangered sp ecies or a threatened species because of any of the following factors (section 4(a)(1)(A)-(E)): (A) the present or threatened, destruction, modi fication, or curtailment of its habitat or

range; (B) overutilization for commercial, recreational, scientific, or educational purposes; (C) disease or predation; (D) the inadequacy of existing regulatory mechanisms; and (E) other natural or manmade factors a ffecting its conti nued existence.

NMFS and FWS can begin the review for listing de terminations, or any interested person may petition for a listing determination unde r section 553(e) of U.S.C. title 5.4 1.1.4. Distinct Population Segment

In 1996, NMFS and FWS published a policy to defi ne the phrase “distinct population segment” (DPS) (FWS and NMFS 1996, 61 FR 4722). The policy defines a population to be a DPS if it is

both discrete and significant relative to its taxon. A population may be considered discrete if it

satisfies either one of the following conditions:

a. It is markedly separated from other populations of the same taxon as a consequence of physical, physiological, ecological, or beha vioral factors. Quantitative measures of genetic or morphological discontinuity ma y provide evidence of this separation.

b. It is delimited by international governmental boundaries within which differences in control of exploitation, management of hab itat, conservation stat us, or regulatory mechanisms exist that are significant in light of section 4(a)(1)(D) of the ESA.

If a population segment is considered discrete, NMFS and/or FWS must then consider whether the discrete segment is significan t relative to its taxon. Criteria that can be used to determine whether the discrete pop ulation segment is significant incl ude, but are not limited to, the following:

a. Persistence of the discrete population segm ent in an ecological setting unusual or unique for the taxon,

b. Evidence that loss of the disc rete population segment would result in a significant gap in the range of the taxon,

c. Evidence that the discrete population segmen t represents the only surviving natural occurrence of a taxon that may be more abundant elsewhere as an introduced population outside its historic range, or

d. Evidence that the discrete population se gment differs markedly from other populations of the species in its genetic characteristics.



### 1.2. History of ESA Listing, Status Reviews, and Petitions

1.2.1. ESA Listing

The loggerhead sea turtle ( Caretta caretta ) was listed as threatened on July 28, 1978 (FW S and NMFS 1978, 43 FR 32800). The threatened listing wa s applied to wherever the species occurs. The major factors contributing to its status included human encroachment and associated activities on nesting beaches; commercial harvest of eggs, subadults, and adults; predation; lack of comprehensive and consistent pr otective regulations; and incidental take in fisheries. Critical habitat has not been designated for the loggerhead sea turtle.

1.2.2. ESA 5-year Reviews

Under the E SA, FWS and NMFS are required to c onduct a review of liste d species under their jurisdiction at least once every 5 years. A 5-year review is an assessment of a listed species to determine whether its status has changed since the time of its listing such that it should be delisted or classified differently than its current status. The pur pose of a 5-year review is to ensure that a listed species has the appr opriate level of protection under the ESA.5 NMFS conducted the first 5-year review for the loggerhead in 1985 (Mager 1985). Data on population trends were limited and were base d largely on the number of nests and nesting females. Of 52 nesting populations examined throughout the Atlantic, Pacific, and Indian Oceans, 33 were thought to be declining, 18 were unknown, and only one – the southeast U.S. Atlantic – was thought to be increasing. Alt hough commercial harvest of eggs had decreased and the U.S. had implemented protective regulations, many threats continued both domestically and abroad. NMFS determined that, since the 19 78 listing, information was insufficient to assess whether a listing change was warranted.

FWS also conducted a 5-year review for the l oggerhead in 1991 (FWS 1991, 56 FR 56882). In this review, the status of many species wa s simultaneously evaluated with no in-depth assessment of the section 4(a)(1) factors as they pertained to the individu al species. The notice stated that FWS was seeking any new or additi onal information reflecting the necessity of a change in the status of the sp ecies under review. The notice i ndicated that if significant data were available warranting a change in a speci es’ classification, FWS wo uld propose a rule to modify the species’ status. No change in the loggerhead threatened listing classification was recommended following this 5-year review.

In 1995, NMFS and FWS conducted a joint 5-year review (Plotkin 1995). New information on population structure indicated the nesting assemblage along the sout heast U.S. coast may consist of two separate populations

– Florida and those nesting in Ge orgia and northward. Nesting in Georgia, South Carolina, and North Carolina was declining. The Florida nesting population was stable, but it was determined that human presence in areas of heavy nesting could impact the population in the future. Although no change in the loggerhead’s list ing classification was recommended, a need for further study of U.S. loggerhead population structure was identified. The 5-year review further stated that should additiona l research confirm th e existence of two separate nesting populations, and if the decline of the norther n nesting population continued,

careful consideration should be given to reclassification dur ing the next 5-year review.

The last 5-year review was completed in 2007 (NMFS and FWS 2007). Many technological advances and a diversity of re search occurred since the 1995 re view. Molecular markers (i.e., mitochondrial DNA and microsatellites) helped de fine the genetic struct uring within and among ocean basins, both at the nesting beaches and at foraging grounds. New information existed on demographic parameters such as age at first re production and survival rates and the biology of loggerheads, especially away from the nesting be ach. These data indica ted a possible separation of populations by ocean basins; however, a more in -depth analysis was needed to determine the application of the DPS policy. Based on the new information and the need for further analysis under the DPS policy, NMFS and FWS recommended that no change in listing status was warranted. They committed to fully assemble and analyze all relevant information in accordance with the DPS policy.

1.2.3. Petitions

Past Petitio ns In 2002, NMFS and FWS received a petition to recl assify the Northern (northeast Florida through North Carolina) and Flor ida Panhandle subpopulatio ns as DPSs, change their status to6 endangered, and designate critical habitat. NMFS found that the petition presented substantial scientific and commercial information indicating that the petitioned reclassification may be warranted (a “positive 90-day finding”) (NMFS and FWS 2002, 67 FR 38459). Based on further analysis of the best available data, NMFS a nd FWS found that although there was some degree of discreteness between the nesting assemblages, the separation was not highly rigid and did not qualify as DPSs (a “12-month determination of not warranted”) (F WS and NMFS 2003, 68 FR 53947). Therefore, the species’ li sting status was not changed.

Current Petitions On July 16, 2007, NMFS and FWS received a petiti on from the Center for Biological Diversity and Turtle Island Restoration Network requesting th at loggerhead turtles in the North Pacific be

reclassified as a DPS with e ndangered status and that criti cal habitat be designated. On November 16, 2007, NMFS and FWS received a pe tition from the Center for Biological Diversity and Oceana requesting that loggerhead turtles in the Northwes t Atlantic Ocean be

reclassified as a DPS with endangered status and that cr itical habitat be designated.

NMFS and FWS determined that the July 16, 2007, North Pacific petition and the November 16, 2007, Northwest Atlantic petition both presented substantial information that the petitioned actions may be warranted (NMFS 2007b, 72 FR 64585; NMFS 2008b, 73 FR 12941). In the published 90-day petition findings, NMFS and FWS committed to assess the loggerhead listing status on a global basis, which is consistent with the recomme ndation in the 2007 5-year review (see Section 1.2.2.).



### 1.3. Approach to the Status Review

NMFS and FWS convened a biological review team (BRT) in February 2008 to review the best available scientific information, determine whether DPSs exist, and assess the extinction risk for each potential DPS. The BRT organized thei r evaluation by ocean basin: Pacific Ocean, Atlantic Ocean (including the Mediterranean Sea), and Indian Ocean. This approach was consistent with the 2007 5-year review for the loggerhead and was chosen to facilitate data assembly and evaluation. It was not meant to preclude identification of DPSs on a broader or finer scale. The BRT primarily reviewed gene tic, tagging (flipper and PI T tags), and satellite telemetry datasets and examined oceanographic features and geographic barriers to determine discreteness and significance (S ection 3). The BRT conducted two independent analyses to assess extinction risks of potential DPSs (Sectio n 4). The first analysis used the diffusion approximation approach based on a time series of counts of nesting females or nests to provide a metric, susceptibility to quasi-ex tinction. The second analysis focused on determining the effects of known anthropogenic mortality on each potential DPS with respect to th e vital rates of the species. This approach focused on how additional mortalities may affect the future growth and status of potential DPSs.7 

## SECTION 2—SPECIES OVERVIEW

2.1. Taxonomy

Kingdom: Animalia Phylum: Chordata Class: Reptilia

Order: Testudines Family: Cheloniidae

Genus: Caretta Species: caretta Common name: Loggerhead sea turtle

The loggerhead was first described by Linnaeus in 1758 and named Testudo caretta . Over the next two centuries more than 35 names were app lied to the species (Dodd 1988), but there is now

agreement on Caretta caretta as the valid name. While Derani yagala described an Indo-Pacific

form as C

. gigas in 1933, he revised that view in 1939 to hold that gigas was a subspecies of C

. caretta (Deraniyagala 1933, 1939). The genus has been regarded as monotypic since that time. The subspecific designation of gigas has likewise been challenge d persuasively (Brongersma 1961, Pritchard 1979). Thorough synonymies and ta xonomic reviews of this form are given most recently by Pritchard and Trebbau (1984) and Dodd (1988). Subspecies assignments are not supported based on genetic evidence (Bowen 2003) or other diagnostic characters (Dodd 1998).



### 2.2. Physical Appearance

The carapace of adult and juveni le loggerheads is reddi sh-brown. The dorsa l and lateral head scales and the dorsal scales of the flippers ar e also reddish-brown, but with light to medium yellow margins. The unscaled areas of the inte gument (neck, shoulders, limb bases, inguinal area) are dull brown dorsally and light to medium yellow laterally and ventrally. The plastron is medium to light yellow, and the thick, bony carap ace is covered by non-overlapping scutes that meet along seam lines. There usually are 11 or 12 pa irs of marginal scutes, five pairs of costals, five vertebrals, and a nuchal (precen tral) scute that is in contact with the first pair of costal scutes. The plastron is composed of paired gular, humeral, pe ctoral, abdominal, femoral, and anal scutes and connected to the carapace by three pairs of poreless inframarginal scutes. Hatchlings vary from light to dark brown to dark gray dorsally and lack the reddish-brown coloration of adults and juveniles. Flippers are dark gray to brown above with distinct white margins. The ventral coloration of the plastron and other areas of the integument are generally yellowish to tan. The carapace has three keels a nd the plastron has two keels. At emergence, hatchlings average 45 mm in SCL and weigh approximately 20 g (Dodd 1988).8 

### 2.3. Distribution and Habitat

The loggerhead occurs throughout the temperate and tr opical regions of the Atlantic, Pacific, and Indian Oceans (Dodd 1988). However, the majority of loggerhead nesting is at the western rims of the Atlantic and Indian Oceans. The most recent reviews show that only two loggerhead nesting aggregations have greater than 10,000 fema les nesting per year: Peninsular Florida, United States and Masirah Island, Oman (Baldwin et al . 2003, Ehrhart et al . 2003, Kamezaki et
al . 2003, Limpus and Limpus 2003b, Margaritoulis et al . 2003). Nesting aggregations with 1,000 to 9,999 females nesting annually are Georgi a through North Carolina (U.S.), Quintana Roo and Yucatán (Mexico), Braz il, Cape Verde Islands (Cap e Verde), Western Australia (Australia), and Japan. Smalle r nesting aggregations with 10 0 to 999 nesting females annually

occur in the Northern Gulf of Mexico (U.S.), Dry Tortugas (U.S.), Cay Sal Bank (The Bahamas), Tongaland (South Africa), Mozambique, Arabian S ea Coast (Oman), Halaniyat Islands (Oman), Cyprus, Peloponnesus (Greece), Zakynthos (Greece) , Crete (Greece), Turkey, and Queensland (Australia). In contrast to determining population size on nesting beaches, determining population size in the marine envi ronment has been very localized (Bjorndal and Bolten 2000). At present, there are no data on population size in the oceanic habitat. Detailed information on

distribution and habitat by ocean basin follows.

Pacific Ocean Loggerheads can be found throughout tr opical to temperate waters in the Pacific; however, their breeding grounds include a restrict ed number of sites in the Nort h Pacific and South Pacific. Within the North Pacific, loggerhead nesting has been documented only in Japan (Kamezaki et
al . 2003), although low level nesting may occur outsi de of Japan in areas surrounding the South China Sea (Chan et al . 2007). In the South Pacific, nesti ng beaches are restricted to eastern Australia and New Caledonia and, to a much le sser extent, Vanuatu and Tokelau (Limpus and Limpus 2003b).

Important loggerhead nesting lo cations in Japan include Yaku shima Island, and Miyazaki, Minabe, and Atsumi beaches on the mainland. Appr oximately 40% of all loggerhead nesting in Japan occurs at three primary nesting beaches on Yakushima Island (Kamezaki et al . 2003). Based on tag-recapture studies, the East China Sea has been identified as the major habitat for post-nesting adult females (Iwamoto et al . 1985, Kamezaki et al . 1997, Balazs 2006), while satellite tracking of juve nile loggerheads indicates the Kurosh io Extension Bifurcation Region to be an important pelagic foraging area for juveniles (Polovina et al . 2006). Other important juvenile foraging areas have been identified off th e coast of Baja California Sur, Mexico (Pitman 1990, Peckham and Nichols 2006).

Nesting occurs along the mainland of Australia from South Stradbroke Is land to Bustard Head, and on the islands of the Capricorn Bunker Group and Swain Reefs, and on Bushy Island (Limpus and Limpus 2003b, Limpus 2009). Within this area, five rookeries account for 70% of

nests in eastern Australia: (1) Mon Repos, (2) Wreck Rock, (3 ) mainland and Wreck Island, (4) Erskine Island, and (5) Tryon Island (Limpus and Reimer 1994). Nesting females tagged on the

coast of eastern Australia ha ve been recorded foraging in New Caledonia; Queensland, New South Wales, and Northern Territory, Austra lia; S olomon Islands; Papua New Guinea; and9 Indonesia (Limpus and Limpus 2003b). Foraging Pacific loggerheads originating from nesting beaches in Australia are known to migrat e to Chile and Peru (Alfaro-Shigueto et al . 2004, 2008a; Donoso and Dutton 2006; Boyle et al . 2009).

Although nesting in the South Pacifi c is concentrated in eastern Au stralia, nesting has also been reported in New Caledonia, Vanuatu, and Tokela u (Limpus and Limpus 2003b). Nesting may occur in other areas of the South Pacific, but it remains unsubstantiated. In New Caledonia, the most substantial loggerhead nest ing has been reported on periphe ral small coral cays offshore from the main island of Île des Pins (Belo ff personal communication cited in Limpus and Limpus 2003b). The population in the Île des Pi ns area has been estimated at 10-100 females nesting annually (Limpus and Limpus 2003b). Based on aerial surveys of New Caledonia nesting beaches conducted between December 2 006 and January 2008, it was estimated that the nesting female population for this nesting seas on was approximately 200 individuals (World Wildlife Fund 2008). In Vanuatu, low density nes ting was reported at Malekula in 1993 (Atuary 1994 cited in Limpus and Limpus 2003b); however, the status of l oggerhead nesting is uncertain because most of Vanuatu has been poorly surv eyed. In 1981, nesting was reported in Tokelau, a territory of New Zealand that comprises three coral atolls in the South Paci fic, but is believed to be uncommon (Balazs 1983).

Indian Ocean In the Southwest Indian Ocean, l oggerhead nesting occurs on the southeastern coast of Africa, from the Paradise Islands in Mozambique southw ard to St. Lucia in South Africa, and on the

south and southwestern coasts of Madagascar (Baldwin et al . 2003). Foraging habitats are only known for post-nesting females from Tongaland, South Africa; tagging data show these loggerheads migrating eastward to Madagasc ar, northward to Mozambique, Tanzania, and Kenya, and southward to Cape Agulhas at the southernmost point of Africa and into the Atlantic Ocean (Baldwin et al . 2003, Luschi et al . 2006).

In the North Indian Ocean, Oman hosts the vast ma jority of loggerhead nes ting. The majority of the nesting in Oman occurs on Masirah Island, on the Al Halaniyat Islands, and on mainland beaches south of Masirah Island all the way to the Oman-Yemen bor der (IUCN - The World Conservation Union 1989a, 1989b; Salm 1991; Salm and Salm 1991). In addition, nesting probably occurs on the mainland of Yemen on th e Arabian Sea coast, and nesting has been confirmed on Socotra, an island off the coast of Yemen (Pilcher and Saad 2000). Limited information exists on the foraging habitats of North Indian Ocean loggerheads; however, foraging individuals have been reported off the southern coastline of Oman (Salm et al . 1993). Satellite telemetry studies conduc ted in Oman have revealed ne w information about post-nesting migrations of loggerheads nesting on Masira h Island (Environment Society of Oman and Ministry of Environment and Climate Change , Oman, unpublished data). Results reveal

extensive use of the waters off the Arabian Peninsul a, with the majority of telemetered turtles (15 of 20) traveling southwest, following the shor eline of southern Oman and Yemen, and circling well offshore in nearby oceanic waters. A minority traveled north as far as the western Persian (Arabian) Gulf (3 of 20) or followed the shorelin e of southern Oman and Yemen as far west as

the Gulf of Aden and the Bab-el -Mandab (2 of 20). These pre liminary data suggest that postnesting migrations and adult female foragi ng areas may be centered within the region10 (Environment Society of Oman and Ministry of Environment and Climate Change, Oman, unpublished data).

The only verified nesting beaches for loggerhea ds on the Indian subcontinent are found in Sri Lanka. A small number of nesting females use the beaches of Sri Lanka every year (Deraniyagala 1939, Kar and Bhaskar 1982, Dodd 1988); however, there are no records indicating that Sri Lanka has ev er been a major nesting area for loggerheads (Kapurusinghe 2006). No confirmed nesting occurs on the ma inland of India despite historical papers suggesting loggerhead sightings on mainland b eaches (Tripathy 2005, Kapurusinghe 2006). This discrepancy may be attributed to inaccurate identifica tion of nesting specie s, as loggerheads are sometimes confused with olive ridleys in the Indian Ocean (Tripathy 2005). In addition, the Gulf of Mannar provides foraging habitat for juveniles and post-nesti ng adults (Tripathy 2005, Kapurusinghe 2006). Although loggerheads have b een reported nesting in low numbers in Myanmar, these data may not be reliable because of misidentification of species (Thorbjarnarson

et al . 2000).

In the East Indian Ocean, western Australia hosts all known loggerhead nesting (Dodd 1988). Nesting distributions in west ern Australia span from the Shark Bay World Heritage Area northward through the Ningaloo Marine Park coas t to the North West Cape and to the nearby Muiron Islands (Baldwin et al . 2003). Nesting individuals fro m Dirk Hartog Island have been recorded foraging within Shark Bay and Exmouth Gulf, while other adults range much farther (Baldwin et al . 2003).

Atlantic Ocean In the Northwest Atlantic, the majority of loggerh ead nesting is concentrat ed along the coasts of the United States from southern Virginia thr ough Alabama. Additional nesting beaches are found along the northern and wester n Gulf of Mexico, eastern Yu catán Peninsula, at Cay Sal Bank in the eastern Bahamas (Addison and Mo rford 1996, Addison 1997), on the southwestern coast of Cuba (F. Moncada-Gavilán, pe rsonal communication, cited in Ehrhart et al . 2003), and along the coasts of Central America, Colombia, Ve nezuela, and the easter n Caribbean Islands. In the Southwest Atlantic, loggerh eads nest in significant numbers only in Brazil. In the eastern Atlantic, the largest nesting popul ation of loggerheads is in the Cape Verde Islands (L.F. LópezJurado, personal communication, cited in Ehrhart et al . 2003), and some nesting occurs along the West African coast (Fretey 2001).

As post-hatchlings, Northwest Atlantic loggerheads hatched on U.S. beaches migrate offshore

and become associated with Sargassum habitats, driftlines, and other convergence zones (Carr 1986, Witherington 2002). These Northwest Atlantic oceanic juveniles use the North Atlantic gyre and enter Northeast Atlantic waters (Carr 1987); they are also found in the Mediterranean Sea (Carreras et al . 2006, Eckert et al . 2008). In these areas, they overlap with animals originating from the Northeast Atlantic and the Mediterranean Sea (Laurent et al . 1993, 1998;

Bolten et al . 1998; LaCasella et al . 2005; Carreras et al . 2006, Monzón-Argüello et al . 2006;

Revelles et al . 2007; Eckert et al . 2008). The oceanic juvenile st age in th e North Atlantic has been primarily studied in the waters around the Azores and Madeira (Bolten 2003). In Azorean waters, satellite telemetry data and flipper tag returns suggest a long period of residency (Bolten 2003), whereas turtles appear to be moving th rough Madeiran waters (Dellinger and Freitas11 2000). Preliminary genetic analys es indicate that juvenile and subadult loggerheads found in Moroccan waters are of western Atlantic origin (M. Tiwari, NMFS, and A. Bolten, University of Florida, unpublished data). Other concentrations of oceanic juveniles exist in the Atlantic (e.g., in the region of the Grand Banks off Newfoundla nd). Genetic information indicates the Grand Banks off Newfoundland are foraging grounds for a mixture of loggerheads from all the North Atlantic rookeries (LaCasella et al . 2005, Bowen et al . 2005), and a large size range is represented (Watson et al . 2004, 2005).

After departing the oceanic zone, neritic juvenile loggerheads in the Northwest Atlantic inhabit continental shelf waters from Cape Cod Ba y, Massachusetts, south through Florida, The Bahamas, Cuba, and the Gulf of Mexico (neritic refers to the inshore marine environment from the surface to the sea floor where water depths do not exceed 200 meters). In the U.S., estuarine waters, including areas such as Long Island S ound, Chesapeake Bay, Pamlico and Core Sounds, Mosquito and Indian River Lagoons, Biscayne Bay, Florida Bay, and numerous embayments fringing the Gulf of Mexico, comprise important in shore habitat. Along the Atlantic and Gulf of Mexico shoreline, essentially all shel f waters are inhabited by loggerheads.

Habitat preferences of Northwest Atlantic non-nesting adult loggerh eads in the neritic zone differ from the juvenile stage in that relatively enclos ed, shallow water estuarine habitats with limited ocean access are less frequently used. Areas such as Pamlico Sound and the Indian River Lagoon in the U.S., regularly used by juveniles, ar e only rarely frequented by adult loggerheads. In comparison, estuarine areas with more open ocean access, such as Chesapeake Bay in the U.S. mid-Atlantic, are also regularly used by juveniles, as well as by adults primarily during warmer seasons. Shallow water habitats with large expa nses of open ocean access, such as Florida Bay, provide year-round resident foragi ng areas for significant numbers of male and female adult loggerheads. Offshore, adults primarily inhabit continental shelf waters, from New York south through Florida, The Bahamas, Cuba, and the Gulf of Mexico. Seasonal use of mid-Atlantic shelf waters, especially offshore New Jersey, Delaware, and Virginia during summer months, and offshore shelf waters, such as Onslow Ba y (off the North Carolina coast), during winter months has been documented (Hawkes et al . 2007a; Georgia Department of Natural Resources, unpublished data; South Carolina Department of Natural Resources, unpublished data). Shelf waters along the west Florida coast, The Bahama s, Cuba, and the Yucatán Peninsula have been identified, using satellite telemetry, as important resident areas for adult female loggerheads that nest in Florida (Foley et al . 2008; M. Lamont, Florida Cooperati ve Fish and Wildlife Research Unit, personal communication, 2009; M. Nic holas, National Park Service, personal communication, 2009). The southern edge of the Grand Bahama Bank is important habitat for loggerheads nesting on the Cay Sa l Bank in The Bahamas, but nesting females are also resident in the bights of Eleuthera, Long Island, and Ragge d Islands as well as Florida Bay in the U.S., and the north coast of Cuba (A. Bolten and K. Bjorndal, University of Flor ida, unpublished data).

Moncada et al . (2009) report the recapture in Cuban wa ters of five adult female loggerheads originally flipper tagged in Qu intana Roo, Mexico, indicating that Cuban shelf waters likely also provide foraging habitat for adult females that nest in Mexico.

In the Northeast Atlantic, satellite telemetry st udies of post-nesting females from Cape Verde identified two distinct dispersal patterns; larger individuals migrated to benthic foraging areas off

the northwest Africa coast and smaller indivi duals foraged primarily oceanically off the12 northwest Africa coast (Hawkes et al . 2006). Monzon-Arguello et al . (2009) conducted a mixed stock analysis of juvenile loggerheads samp led from foraging areas in the Canary Islands, Madeira, Azores, and Andalusia and concluded that while juveniles from the Cape Verde population were distributed among these four site s, a large proportion of Cape Verde juveniles appear to inhabit as yet unidentified foraging areas.

In the South Atlantic, relatively little is know n about the at-sea behavior of loggerheads originating from nesting beaches in Brazil. R ecaptures of tagged juveniles and nesting females have shown movement of animals up and down the coast of South America (Almeida et al . 2000,

Marcovaldi et al . 2000, Laporta and Lopez 2003, Almeida et al . 2007). Juvenile loggerheads, presumably of Brazilian origin, have also been captured on the high seas of the South Atlantic (Kotas et al . 2004, Pinedo and Polacheck 2004) and off th e coast of Atlantic Africa (Bal et al

. 2007, Petersen 2005, Petersen et al . 2007) suggesting that loggerheads of the South Atlantic may undertake transoceanic developmental migrations (Bolten et al . 1998, Peckham et al . 2007).

Mediterranean Sea Loggerhead turtles are widely distributed in the Me diterranean Sea. However, nesting is almost entirely confined to the eastern Mediterranean basin, with the main nesting concentrations in Cyprus, Greece, and Turkey (Margaritoulis et al . 2003). Preliminary surveys in Libya suggested nesting activity comparable to Greece and Turk ey, although a better quan tification is needed (Laurent et al . 1999). Minimal to moderate nesting also occurs in other countries throughout the Mediterranean including Egypt, Israel, Italy (sout hern coasts and islands), Lebanon, Syria, and Tunisia (Margaritoulis et al . 2003). Recently, isolated nesting ev ents have been recorded in the western Mediterranean basin, namely in Spain, Corsica (France), and in the Tyrrhenian Sea (Italy) (Bentivegna et al . 2005, Delaugerre and Ce sarini 2004, Tomás et al . 2002).

In Cyprus, nesting occurs mainly on beaches of the western coast and Chrysochou Bay (Demetropoulos and Hadjichristophorou 1989), as well as along the northern coast (Broderick and Godley 1996). Nesting occurs along the west ern and southern coasts of Greece and on the island of Crete, with the vast majority of nesting occurring on the island of Zakynthos (Margaritoulis 1987, 1998, 2005; Margaritoulis et al . 1995, 2003). Seventeen important loggerhead nesting sites have been identif ied on Turkey’s beaches (Margaritoulis et al . 2003). Nesting activity in Libya is spread throughout the entire coast, but may be less dense in western areas (Laurent et al . 1999).

Important neritic marine habitats have been sugg ested for the large continental shelves of: (1) Tunisia-Libya, (2) northern Adriatic Sea, (3) Egypt, and (4) Spain (Margaritoulis 1988, Argano

et al . 1992, Laurent and Lescure 1994, Lazar et al . 2000, Gomez de Segura et al . 2006,

Broderick et al . 2007, Casale et al . 2007b, Nada and Casale 2008). At least the first three

constitute shallow benthic habi tats for adults (including post-nesting females). Some other neritic foraging areas include Amvrakikos Bay in western Greece, Lakonikos Bay in southern Greece, and southern Turkey. Oceanic foraging areas for small juveniles have been identified in the south Adriatic Sea (Casale et al . 2005b), Ionian Sea (Deflorio et al . 2005), Sicily Strait (Casale et al . 2007b), and western Mediterran ean (Spain) (e.g., Caminas et al . 2006). In addition, tagged juveniles have been recorded cr ossing the Mediterranean fr om the eastern to the13 western basin and vice versa, as well as in the Eastern Atlantic (Argano et al . 1992, Casale et al

. 2007b).

Reproductive migrations have been confirmed by flipper tagging and satellite telemetry. Female loggerheads, after nesting in Greece, migrate prim arily to the Gulf of Gabès and the northern Adriatic (Margaritou lis 1988, Margaritoulis et al . 2003, Lazar et al . 2004, Zbinden et al . 2008). Loggerheads nesting in Cyprus mi grate to Egypt and Libya, exhibi ting fidelity in following the same migration route during subseq uent nesting seasons (Broderick et al . 2007). In addition, directed movements of juvenile loggerheads have been confirmed through flipper tagging (Argano et al . 1992, Casale et al . 2007b) and satellite tracking (Rees and Margaritoulis 2009).



### 2.4. Biological Characteristics

Loggerheads nest on ocean beaches and occasionally on estuarine shorelines. Although specific characteristics vary between rookeries, logge rhead nesting beaches tend to be wide, sandy beaches backed by low dunes and fronted by a flat, sandy approach from the water (Miller et al

. 2003). Nests are typically laid between the high tide line and th e dune front (Routa 1968, Witherington 1986, Hailman and Elowson 1992).

Sea turtle eggs require a high-humidity substrat e that allows for sufficient gas exchange and temperatures conducive to egg development (Miller 1997, Miller et al . 2003). Mean clutch size varies greatly between populat ions, but on average is approx imately 100-130 eggs per clutch (Dodd 1988). Loggerhead nests incubate for vari able periods of time. The length of the incubation period is inversel y related to nest temperat ure, such that between 26 o C and 32 o C, a change of 1 o C adds or subtracts approximately 5 da ys (Mrosovsky 1980). The warmer the sand surrounding the egg chamber, the faster the em bryos develop (Mrosovsky and Yntema 1980). Sand temperatures prevailing during the middle thir d of the incubation period also determine the sex of hatchlings (Mrosovsky and Yntema 1980). Incubation temperatures near the upper end of the tolerable range produce only fe male hatchlings while incubati on temperatures near the lower end of the tolerable range produce only male ha tchlings. The pivotal temperature (i.e., the incubation temperature that produces equal numbers of males and females) in loggerheads is approximately 29 o C (Limpus et al . 1983, Mrosovsky 1988, Marcovaldi et al . 1997). Moisture

conditions in the nest influence incubation period, hatching success, and hatchling size (McGehee 1990, Carthy et al . 2003).

Loggerhead hatchlings pip and escape from thei r eggs over a 1- to 3- day interval and move upward and out of the nest over a 2- to 4-day interval (Christens 1990). The time from pipping to emergence ranges from 4 to 7 days with an average of 4.1 days (Godfrey and Mrosovsky 1997). Hatchlings emerge from their nests en masse almost exclusively at night, and presumably using decreasing sand temperature as a cu e (Hendrickson 1958, Mr osovsky 1968, Witherington

et al . 1990). Moran et al . (1999) concluded that a lowering of sand temperatures below a critical threshold, which most typically occurs after night fall, is the most probable trigger for hatchling emergence from a nest. After an initial emergence, there may be secondary emergences on subsequent nights (Carr and Ogren 1960, With erington 1986, Ernest and Martin 1993, Houghton and Hays 2001).14 Hatchlings use a progression of orientation cues to guide their movement from the nest to the marine environments where they spend th eir early years (Lohmann and Lohmann 2003). Hatchlings first use light cues to find the ocean. On naturally lighted b eaches without artificial lighting, ambient light from the open sky creates a relatively bright horizon compared to the dark silhouette of the dune and vegetation landward of the nest. This contrast gui des the hatchlings to the ocean (Daniel and Smith 1947, Limpus 1971, Salmon et al . 1992, Witherington 1997, Witherington and Martin 1996, Stewart and Wyneken 2004).

Immediately after hatchlings emerge from the ne st, they begin a period of frenzied activity. During this active period, hatchl ings move from their nest to the surf, swim, and are swept through the surf zone (Carr and Ogren 1960; Carr 1962, 1982; Wyneken and Salmon 1992; Witherington 1995). Orientation cues used by hatc hlings as they crawl, swim through the surf, and migrate offshore are discussed in detail by Lohmann and Lohmann (2003).

Neonate loggerheads that have migrated away from land differ from swim frenzy stage hatchlings in that they are largely inactive and only exhib it infrequent low-energy swimming, and they have begun to feed, no longer relying on their retained yolk (Witherington 2002). As post-hatchlings, loggerheads are pelagic and ar e best known from neritic waters along the

continental shelf. This neritic post-hatchling stage is weeks or months long (Witherington 2002) and may be a transition to the oceanic stage that loggerheads enter as they grow and are carried within ocean currents (Bolten 2003).

In the northwest Atlantic, pos t-hatchling loggerheads inhab it areas where surface waters converge to form local downwellings (Witheri ngton 2002). These areas are characterized by linear accumulations of floating material, especially Sargassum , and are common between the Gulf Stream and the southeast U.S. coast, and between the Loop Current a nd the Florida coast inthe Gulf of Mexico. Post-hatchlings within this habitat are observed to be low-energy float-andwait foragers that feed on a wide variety of floating items (Witherington 2002). Witherington (2002) found that small animals commonly associated with the Sargassum community, such as hydroids and copepods, were mo st commonly found in esophagea l lavage samples. Posthatchling loggerheads from southeast U.S. nesting beaches may linger for months in waters just off the nesting beach or become transported by ocean currents within the Gulf of Mexico and

North Atlantic.

The oceanic juvenile stage begins when loggerhea ds first enter the oceanic zone (Bolten 2003). Juvenile loggerheads originating from nesting be aches in the Northwest Atlantic, West Indian, and West Pacific Oceans appear to use oceanic developmental habitats and move with the predominant ocean gyres for several years before returning to their neritic foraging and nesting habitats (Bolten 2003, Bowen et al . 1995, Hughes 1974a, Musick and Limpus 1997, Pitman 1990, Zug et al . 1995). However, the actual duration of th e oceanic juvenile stage varies with loggerheads leaving the oceanic zone over a wide size range (Bjorndal et al . 2000). In the Atlantic, Bjorndal and colleagues (Bjorndal et al . 2000, 2003a) estimated the duration of the oceanic juvenile stage to be betw een 7 and 11.5 years, with juvenile s recruiting to neritic habitats in the western Atlantic over a size range of 46-64 cm curved carapace length (CCL) (Bolten et
al . 1993, Turtle Expert Working Group 2009). Ho wever Snover (2002) suggests a much longer oceanic juvenile stage duration with a range of 9-24 years and a mean 14.8 years over similar15 size classes. However, in Japa n and Australia, juvenile loggerh eads do not disperse to neritic habitats until larger than ar ound 60 cm SCL (Y. Matsuzawa and S ea Turtle Association of Japan, unpublished data) and 70 cm CCL or larger (Limpus et al . 1994), respectively, while in the Mediterranean they do so at around 25 cm CCL (Casale et al . 2008a).

The neritic juvenile stage begins when loggerhea ds exit the oceanic zone and enter the neritic zone (Bolten 2003). After migrating to the ner itic zone, juvenile loggerheads continue maturing until they reach adulthood. Some juveniles ma y periodically move between the neritic and oceanic zones (Witzell 2002, Bolten 2003, Morre ale and Standora 2005, McClellan and Read 2007, Mansfield 2006, Eckert et al . 2008). The neritic zone al so provides important foraging habitat, inter-nesting habitat, a nd migratory habitat for adult logge rheads. Some adults may also periodically move between the neritic and o ceanic zones (Harrison and Bjorndal 2006). See

Schroeder et al . (2003) and Limpus and Limpus (2003a) for reviews of this life stage for the Atlantic and Pacific, respectively.

The duration of the adult stage can be reasonably estimated for females from tag return data at nesting beaches. For the Northwest Atlantic nesting assemblages, data from Little Cumberland Island, Georgia, show reproductive longevity, and hence duration of adult female stage, as long as 25 years (Dahlen et al . 2000). This is likely an underestim ate of the averag e reproductive life span given tag loss and incomplete surveys of ne sting beaches at night. Comparable data for

adult males do not exist.

Based on stable isotope analyses and satellite telemetry, Hatase et al . (2002a) demonstrated that some adult female loggerheads nesting in Japan inhabit oceanic habitats rather than neritic habitats. Satellite tagged adult loggerheads in we stern Africa have also been demonstrated to use oceanic foraging areas (Hawkes et al . 2006). Preliminary results from stable isotope analyses suggest that some loggerheads nesting in Florid a also may inhabit ocean ic habitats (Reich et al

. 2007). In both Japan and Florida, the females inhabiting oceanic habita ts were significantly

smaller than those in neritic habitats. The ex tent to which adult loggerheads occupy oceanic habitats needs to be evaluated, and effects on survival probabilities and reproductive output

should be assessed.

In both the oceanic and neritic zones, loggerh eads are primarily carnivorous, although they do consume some plant matter as well (s ee Bjorndal 1997 and Dodd 1988 for reviews). Loggerheads are able to exist on a wide variety of food items with ontogenetic and regional differences in diet. Loggerhead diets have been described from just a few coastal regions, and very little information is availabl e about differences or similarities in diet at various life stages. Very little is known of the diet of oceanic juveniles.16 

## SECTION 3—DETERMINATION OF DPS

3.1. Overview of Informatio n Used to Determine DPS

The BRT considered a vast array of informati on in assessing whether there are any loggerhead population segments that satisfy the Distinct Po pulation Segment criteria of being both discrete and significant.

First, the BRT discussed whether there were any loggerhead population segments that were discrete. As noted previousl y, joint NMFS/FWS policy defines a population to be a DPS if it is both discrete and significant relative to the taxon to which it belongs (FWS and NMFS 1996, 61 FR 4722). Under the policy, a population may be cons idered discrete if it sa tisfies either one of the following conditions: (1) it is markedly se parated from other populations of the same taxon as a consequence of physical, physiological, ecolo gical, or behavioral f actors; or (2) it is delimited by international govern mental boundaries within which differences in control of exploitation, management of hab itat, conservation status, or regul atory mechanisms exist that are significant in light of Secti on 4(a)(1)(D) of the ESA.

Data relevant to the discrete ness question include phys ical, physiological, ec ological, behavioral, and genetic data. Upon looking at the global loggerhead populat ion, the physical separation of ocean basins by continents was first considered. The result was an evaluation of the data for each ocean basin (Pacific Ocean, Atlantic Ocean, and Indian Ocean). This was not to preclude any larger or smaller DPS delinea tion, but to aid in data organization and assessment. The BRT then evaluated genetic information by ocean basin. The genetic data consisted of results from studies using maternally inhe rited mitochondrial DNA (mtDNA) and biparentally inherited nuclear DNA microsatellite markers. Next, ta gging data (both flipper and PIT tags) and telemetry data were reviewe d. Additional information, such as potential differences in morphology, was also evaluated. Fi nally, the BRT considered whet her the available information on loggerhead population segments was bounded by any oceanographic features (e.g., current systems) or geographic features (e.g., land masses).

In accordance with the DPS policy, the BRT also reviewed whether the population segments identified in the discreteness analysis were signi ficant. If a population segment is considered discrete, its biological and ecol ogical significance must then be considered. NMFS/FWS must

consider available scientific ev idence of the discrete segment’s importance to the taxon to which it belongs. Data relevant to the significance ques tion include the morphological, ecological, behavioral, and genetic data, as described above. The BRT c onsidered the following factors, listed in the DPS policy, in determining whet her the discrete population segments were significant: a) persistence of the disc rete segment in an ecological setting unusual or unique for the

taxon; b) evidence that loss of the discrete segment woul d result in a significan t gap in the range of the taxon; c) evidence that the discrete segment represents the only surviving natural occurrence of a taxon that may be more abundant elsewher e as an introduced population outside its historical range; and17 d) evidence that the discrete segment differs ma rkedly from other populations of the species in its genetic characteristics.

A discrete population segment needs to satisfy only one of these criteri a to be considered significant. The NMFS/FWS policy also allows fo r consideration of other factors if they are appropriate to the biology or ecology of the species . As will be described in subsequent sections,

the BRT evaluated the information and considered ite ms (a), (b) and (d), as noted above, to be most applicable to loggerheads.

3.1.1. Discreteness Determination

The loggerhead sea turtle ( Caretta caretta ) is p resent in all tropical and temperate ocean basins, and has a life history that involves nesting on coas tal beaches and foraging in neritic and oceanic habitats, as well as long-distan ce migrations between and within these areas. As with other globally distributed marine species, today’s gl obal loggerhead population has been shaped by a sequence of isolation events cr eated by tectonic and oceanogra phic shifts over geologic time scales, the result of which is populati on substructuring in many areas (Bowen et al . 1994, Bowen 2003). Globally, loggerhead turtle s comprise a mosaic of populat ions, each with unique nesting sites and in many cases possessing disparate dem ographic features (e.g., mean body size, age at first reproduction) (Dodd 1988). However, despite these differences, loggerheads from different populations often mix in comm on foraging grounds (Bolten and Witherington 2003), thus creating unique challenges when attempting to delineate distinct population segments for management or listing purposes.

Examining the phylogeography of loggerheads ac ross their global dist ribution through mtDNA sequence diversity, Bowen et al . (1994) found it to be si milar to green turtles ( Chelonia mydas ), with a separation of loggerheads in the Atlantic-Mediterranean basins from those in the IndoPacific basins since the Pleist ocene period. The divergence betw een these two primary lineages corresponds to approximately thre e million years, based on a molecular clock for control region sequences assessed originally for green turtles (2% per million years; Dutton et al . 1996,

Encalada et al . 1996). Geography and climate appear to ha ve shaped the evol ution of these two matriarchal lineages with the ons et of glacial cycles, the appearance of the Panama Ithsmus creating a land barrier betw een the Atlantic and eastern Pacifi c, and upwelling of cold water off southern Africa creating an oceanographic barri er between the Atlantic and Indian Ocean (Bowen 2003). Recent warm temperatures during interglacial periods allowed bi-directional invasion by the temperate-adapted loggerhead s into the respective basins (Bowen et al . 1994; J.S. Reece, Washington University, personal communication, 2008). Today, it appears that loggerheads within a basin are effectively isolated from populations in the other basin, but some dispersal from the Tongaland rookery in the I ndian Ocean into feeding and developmental

habitat in the South A tlantic is possible via the Agulhas current (G.R. Hughes, unpublished data,

cited in Bowen et al . 1994). In the Pacific, extensive mt DNA studies show that the northern loggerhead populations are isolated from the southern Pacific populations, and that juveniles from these distinct genetic populations do not disperse across the equator (Hatase et al . 2002a; Dutton 2007, unpublished data).18 Mitochondrial DNA data indicate th at regional turtle rookeries with in an ocean basin have been strongly isolated from one another over ecological tim escales (Bowen et al . 1994, Bowen and Karl 2007). These same data indicate strong fema le natal homing and suggest that each regional nesting population is an independent demographic unit (Bowen and Karl 2007) . It is difficult to determine the precise boundaries of these demogr aphically independent populations in regions, such as the eastern U.S. coast, where rookeries are close to each othe r and range along large areas of a continental coastline. There appear s to be varying levels of connectivity between proximate rookeries facilitated by imprecise natal homing and male mediated gene flow (Pearce 2001, Bowen 2003, Bowen et al. 2005). Regional genetic populat ions often are characterized by allelic frequency differences rather than fixed genetic differences. There is concern that current analytical tools are unable to identify discrete or demographically independent populations based on genetic data when the allelic frequency differe nces are slight, and there is a need for new approaches to identifying the appropriate uni ts to conserve (Taylor and Dizon 1999). For example, incorporating a metapopulation approach to regional population st ructure analysis may be informative for future work addressing c onnectivity and discretene ss of marine turtle populations (Crooks and Sanjayan 2006, Kritzer and Sale 2006, Dutton et al . 2007).

Through the evaluation of genetic data, tagging data, telemetry, and demography, the BRT

determined that there are at least nine discre te population segments for loggerhead sea turtles globally. These discrete population segments ar e markedly separated from each other as a consequence of ecological, behavioral, and oceanographic factors, and given the genetic evidence, the BRT has unanimously conclude d that each regional popul ation identified is discrete from other populations of loggerheads. Information considered by the BRT in its delineation of discrete p opulation segments is presented below by ocean basin.

Pacific Ocean Perhaps the most distinct and easily recognized of all loggerhead populations is that from the North Pacific Ocean. The primary nesting areas for this population are found along the southern Japanese coastline and Ryukyu Archipelago (Kamezaki et al . 2003). Loggerhead turtles hatching on Japanese beaches undertake extensiv e developmental migrations utilizing the

Kuroshio and North Pacific Cu rrents (Balazs 2006, Kobayashi et al . 2008), and some turtles reach the vicinity of Baja Ca lifornia in the eastern Pacifi c (Uchida and Teruya 1988, Bowen et
al . 1995, Peckham et al . 2007). After spending years foraging in the central and eastern Pacific, loggerheads return to their natal beaches for repro duction (Resendiz et al . 1998, Nichols et al

. 2000) and remain in the western Pacific for the remainder of thei r life cycle (Iwamoto et al. 1985, Kamezaki et al . 1997, Sakamoto et a l . 1997, Hatase et al . 2002c).

Despite the long-distance developmental movements of loggerheads in the North Pacific, current scientific evidence, ba sed on genetic analysis, flipper tag recoveries, and satellite telemetry, indicates that individuals originating from Japan remain in the North Pacific for their entire life cycle, never crossing the equator or mixing with individuals from the South Pacific (Hatase et al

. 2002a; Dutton 2007, unpublished data; LeRoux and Dutton 2006). Indeed, this apparent complete separation of two adjacent populations is unique and most likely results from: (1) the presence of two distinct Northern and Southern Gyre (current flow) systems in the Pacific (Briggs 1974), (2) near-passive movements of post -hatchlings in these gyres that initially move them farther away from areas of potential mi xing among the two populations along the equator,19 and (3) the nest-site fidelity of adult turtles that prevents tu rtles from returning to non-natal nesting areas.

Pacific loggerheads are further partitioned evol utionarily from other loggerheads throughout the world based on additional analyses of mtDNA. The haplotypes from both North and South Pacific loggerheads are disti nguished by a minimum genetic di stance (d) equal to 0.017 from other conspecifics, which indicates isolation of approximately one million years (Bowen 2003).

Within the Pacific, Bowen et al . (1995) used mtDNA to identify two genetically di stinct nesting populations in the Pacific – a northern hemisphe re population nesting in Japan and a southern hemisphere population nesting primarily in Aust ralia. This study also suggested that some loggerheads sampled as bycatch in the North Pacific might be from the Australian nesting population (Bowen et al . 1995). However, more extensive mtDNA rookery data from Japan (Hatase et al . 2002a) taken together with preliminary results from microsatellite (nuclear) analysis confirms that loggerheads inhabiting th e North Pacific actually originate from nesting beaches in Japan (P. Dutton, NMFS, unpublished data). LeRoux et al . (2008) report additional genetic variation in North Pacific loggerhead s based on analyses using new mtDNA primers designed to target longer mtDNA sequences, and suggest finer scale population structure in North Pacific loggerheads may be present.

Although these studies indi cate genetic distinctness between logge rheads nesting in Japan versus those nesting in Australia, Bowen et al . (1995) did identify individuals with the common Australian haplotype at foraging areas in the North Pacific, ba sed on a few individuals sampled as bycatch in the North Pacific. More recently, Hatase et al . (2002a) detected this common haplotype at very low frequenc y at Japanese nesting beaches. However, the presence of the common Australian haplotype does not preclude the genetic distinctiveness of Japanese and Australian nesting populations, and is likely the result of rare ge ne flow events occurring over geologic time scales.

The distinct status of loggerheads in the Nort h Pacific is further supported by the results from flipper tagging in the North Pacific. Flippe r tagging of loggerheads has been widespread throughout this region, occurring on adults nestin g in Japan and incidentally bycaught in the coastal pound net fishery (Y. Matsuzawa, Sea Turtle Association of Japan, personal communication, 2006), juveniles reared and released in Japan (Uchida and Teruya 1988, Hatase

et al . 2002a), juveniles foraging near Baja Ca lifornia, Mexico (Nichols 2003, Seminoff et al

. 2004), and loggerheads captured in and tagged from commercial fisheries platforms in the North Pacific high seas (NMFS, unpublished data). To date , there have been at l east three transPacific tag recoveries showing east-west and west-east movements (Uch ida and Teruya 1988; Resendiz

et al . 1998; W.J. Nichols, Ocean Conservancy, a nd H. Peckham, Pro Peninsula, unpublished data) and several recoveries of adu lts in the western Pacific (Iwamoto et al .1985

, Kamezaki et al

. 1997). However, despite the more than 30,000 mark ed individuals, not a si ngle tag recovery has been reported outside the North Pacific.

A lack of movements by loggerheads south acr oss the equator has also been supported by extensive satellite telemetry. As with flipper tagging, satellite telemetry has been a tool used widely in the North Pacific, w ith satellite transm itters being placed on adult turtles departing20 nesting beaches (Sakamoto et al . 1997; Japan Fisheries Reso urce Conservation Association 1999; Hatase et al . 2002b, 2002c), on adults and sub-adults bycaught in pound nets off the coast of Japan (Sea Turtle Association of Japan, unpublis hed data), on headstarte d juveniles released in Japan (Balazs 2006), on juvenile, subadult and a dult turtles bycaught in the eastern and central North Pacific (e.g., Kobayashi et al . 2008), and on juvenile and s ubadult turtles foraging in the eastern Pacific (Nichols 2003; Peckham et al . 2007; J. Seminoff, NMFS, unpublished data). Of the nearly 200 transmitters depl oyed on loggerheads in the No rth Pacific, none have moved south of the equator. These studies have dem onstrated the strong associ ation loggerheads show with oceanographic mesoscale features such as the Transition Zone Chlorophyll Front or the Kuroshio Current Bifurcation Region (Polovina et al . 2000, 2001, 2004, 2006; Etnoyer et al

. 2006; Kobayashi et al . 2008). Kobayashi et al . (2008) demonstrated that loggerheads strongly track these zones even as they shift in location, thus suggesting that str ong habitat specificity during the oceanic stage also contributes to the lack of mixing. Telemetry studies in foraging areas of the eastern Pacific, near Baja California, Mexico (Nichols 2003; Peckham et al . 2007; H. Peckham, Pro Peninsula, unpublished data) and Peru (J. Mangel, Pro Delphinus, unpublished data) similarly show a complete lack of long distance north or south movements.

The North Pacific population of l oggerheads appears to occupy an ecological setting distinct from other loggerheads, including those of the South Pacific populati on. In general, this is the only population of loggerheads to be found north of the equator in the Pacific Ocean, foraging in

the eastern Pacific as far south as Ba ja California Sur, Mexico (Seminoff et al . 2004, Peckham et
al . 2007) and in the western Pacific as far south as the Philippines (Limpus 2009) and the mouth of Mekong River, Vietnam (Sadoyama et al. 1996). Pelagic juveniles have been found to spend much of their time foraging in the central and eastern North Pacific Ocean. The Kuroshio Extension Current, lying west of the international date line, serves as the dominant physical and biological habitat in the North Pacific and contains high productivity, likely due to unique

features such as eddies and m eanders that concentrate prey and allow food webs to develop. Juvenile loggerheads originating from nesting be aches in Japan were found to exhibit high site fidelity to an area referred to as the Kuroshio Extension Bifurcation Region, an area with extensive meanders and mesoscale eddies (Polovina et al . 2006). Juveniles also were found to correlate strongly with areas of surface chlorophyll a levels in an area known as the Transition Zone Chlorophyll Front, an area concentra ting surface prey for loggerheads (Polovina et al

. 2001, Parker et al . 2005, Kobayashi et al . 2008). Another area found ecologically unique to the North Pacific population of loggerheads, likely becau se of the high density of pelagic red crabs ( Pleuronocodes planipes ), is lo cated off the Pacific coast of the Baja California Peninsula, Mexico, where researchers have documented a foraging area for juveniles based on aerial surveys and satellite telemetry (Seminoff et al . 2006, Peckham et al . 2007). Tag returns show post-nesting females migrating into the East China Sea off South Korea, China, and the Philippines, and the nearby coastal waters of Japan (Iwamoto et al. 1985; Kamezaki et al . 1997, 2003). Clearly, the North Pacific population of loggerheads is unique ly adapted to the ecological setting of the North Pacific Ocean and throughout its long life history serves as an important part of the ecosystem it inhabits.

Loggerheads inhabiting the North Pacific Ocean are derived from Japanese beaches, with the possible exception of rare waifs over evolutionary time scales. Fu rthermore, nesting colonies of Japanese loggerheads are found to be genetically distinct based on mtDN A analyses, and when21 compared to much larger and more genetically diverse loggerhead popula tions in the Atlantic and Mediterranean, Pacific loggerheads have likely experienced cr itical bottlenecks (in Hatase et
al . 2002a), underscoring the importan ce of management and protecti on in retaining this genetic population.

In the South Pacific Ocean, loggerhead turtles ne st primarily in Queensland, Australia, and, to a lesser extent, New Caledonia and Vanuatu (Limpus and Limpus 2003b, Limpus et al . 2006, Limpus 2009). Loggerheads from these rooke ries undertake an oceanic developmental migration, traveling to habitats in the central and southeastern Pacific Ocean where they may reside for several years prior to returning to the wester n Pacific for reproduction. Loggerheads in this early life history stage differ markedly fr om those originating fr om western Australia beaches in that they undertake long west-to-east migrations, likely using specific areas of the pelagic environment of the South Pacific Ocean. An unknown portion of these loggerheads forage off Chile and Peru, and preliminary ge netic information from foraging areas in the southeastern Pacific confirms that the haplot ype frequencies among immature turtles in these areas closely match those found at nesting beaches in eastern Australia (Alfaro-Shigueto et al

. 2004; Donoso and Dutton 2006, 2007; Boyle et al . 2009). Large immature and adult loggerheads generally remain in the western South Pacific, inhabiti ng neritic and oceanic foraging sites during non-nesting periods (Limpus et al . 1994, Limpus 2009).

Loggerheads from Australia and New Caledonia appare ntly do not travel north of the equator. Flipper tag recoveries from nesting females ha ve been found throughout the western Pacific, including sites north of Australia, the Torres Straight, and the Gulf of Carpentaria (Limpus 2009). Of approximately 1,000 (adult and immature ; male and female) lo ggerheads that have been tagged in eastern Australian feeding areas, on ly two have been recorded migrating to breed outside of Australia; both traveled to New Ca ledonia (Limpus 2009). Flipper tagging programs in Peru and Chile tagged approximately 500 loggerheads from 1999-2006, none of which have been reported from outside of the s outheastern Pacific (Alfaro-Shigueto et al . 2008a; S. Kelez, Duke University Marine Laboratory, unpublishe d data; M. Donoso, ONG Pacifico Laud - Chile, unpublished data). Limited satellite telemetry data (12 tags) in the area show a similar trend (J. Mangel, Pro Delphinus, unpublished data).

The spatial separation between the North Pacifi c and South Pacific logg erhead populations has

contributed to substantial differe nces in the genetic profiles of the nesting populations in these two regions. Whereas the dominant mtDNA haplot ypes among loggerheads nesting in Japan are CCP2 and CCP3 (equivalent to B and C respectively in Bowen et al . 1995 and Hatase et al

. 2002a; Leroux et al . 2008; P. Dutton, NMFS, unpublished da ta), loggerheads nesting in eastern Australia have a third haplotype (CCP1, previously A) which is dominant (98% of nesting females) (Bowen et al . 1994, FitzSimmons et al . 1996, Boyle et al . 2009). Further, preliminary genetic analysis using microsatellite markers (nuclear DNA) indicates genetic distinctiveness between nesting populations in the N orth vers us South Pacific (P. Dutton, NMFS, personal communication, 2008).

The separateness between nesting populations in eastern and west ern Australia is less clear, although these too are considered to be genetically dis tinct from one another (Limpus 2009). For example, mtDNA haplotype CCP1, which is th e overwhelmingly dominant haplotype among22 eastern Australia nesting females (98%), is al so found in western Australia, although at much lower frequency (33%) (FitzSimmons et al . 1996, 2003). The remaining haplotype for both regions was the CCP5 haplotype. Further, Fitz Simmons (University of Canberra, unpublished data) found significant differences in nuclear DNA microsatellite lo ci from females nesting in these two regions. Estimates of gene flow betw een eastern and western Australian populations was an order of magnitude less than gene flow within regions. Although some level of malemediated gene flow may occur within ecologi cal time frames, presumably during mixing at foraging areas near the Torres Stra ight and in the Gulf of Carp entaria, the result may also indicate homoplasy within the genetic markers in the two regions leading to a false impression of similarity. Male-mediated gene flow between ea stern and western Austra lia, therefore, may be insignificant, which, when considered in light of the substantial disp arity in mtDNA haplotype frequencies between these two regions, provi des further evidence of population separation between these two regions.

At present, there is no indicati on from genetic studies that th e loggerhead turtles nesting in eastern Australia are distinct from those nesting in New Caledonia. Of 27 turtles sequenced from New Caledonia, 93% carried the CCP1 haplotyp e and the remaining had the CCP5 haplotype; similar to eastern Australia (Boyle et al. 2009). There remains a need to analyze these populations using nuclear DNA microsatellites.

The South Pacific population of loggerheads occ upies an ecological setting distinct from other loggerheads, including the North Pacific popul ation. Much less is known regarding the ecosystem upon which oceanic juveniles and ocea nic adults depend. Sea surface temperature and chlorophyll frontal zones in the South Pacific have been s hown to dramatically impact the movements of green turtles, Chelonia mydas (Seminoff et al . 2008) and leatherback turtles, Dermochelys coriacea (Shillinger et al . 2008), and it is likely that loggerheads similarly benefit

from interactions with these mesoscale oceanographic features.

Loggerheads in the South Pacific are substant ially impacted by periodic environmental perturbations such as the El Niño Southern Oscillation (ENSO). Th is 3- to 6-year cycle within the coupled ocean-atmosphere system of the tr opical Pacific brings increased surface water temperatures and lower primary productivit y, both of which have profound biological consequences (Chavez et al . 1999). Loggerheads are presumab ly adversely impacted by the lower food availability that of ten results from ENSO events, although data on this subject are lacking. Although ENSO may last for only short pe riods and thus not have a long-term effect on loggerheads in the region, recent studies by Chaloupka et al . (2008) suggest that long-term increases in sea surface temperature within the South Pacific may influence the ability of the Australian nesting population to recove r from historic population declines.

Loggerheads originating from nesting beaches in the western South Pacific are the only population of loggerheads to be found south of the equator in the Pacific Ocean. As posthatchlings, they are generally swept sout h by the East Australian Current (Limpus et al . 1994), spend a large portion of time foraging in the oceanic South Pacific Ocean, and a fraction migrate

to the southeastern Pacific Ocean off the coasts of Peru and Chile as juven iles (Alfaro-Shigueto

et al . 2004, Donoso et al . 2000, Boyle et al . 2009). As large immatu res and adults, these loggerheads’ foraging range encompasses the eastern Arafura Sea, Gulf of Carpentaria, Torres23 Strait, Gulf of Papua, Coral Sea, and wester n Tasman Sea to southern New South Wales including the Great Barrier Reef, Hervey Bay, and Moreton Bay. The outer extent of this range

includes the coastal waters off eastern Indonesi a northeastern Papua Ne w Guinea, northeastern Solomon Islands, and New Caledonia (in Limpus 2009).

All loggerheads inhabiting the South Pacific Ocean are derived from beaches in eastern Australia

and a lesser known number of beaches in s outhern New Caledonia, Vanuatu, and Tokelau (Limpus and Limpus 2003b, Limpus 2009). Furtherm ore, nesting colonies of the South Pacific population of loggerheads are found to be genetica lly distinct from loggerheads in the North

Pacific and Indian Ocean.

Given the information presented above, the BRT has unanimously concluded that two discrete population segments exist in the Pacific Ocean as a consequence of ecological, behavioral, and oceanographic factors, and based on genetic evid ence: (1) North Pacific Ocean and (2) South

Pacific Ocean.

Indian Ocean Loggerhead sea turtles in the Indian Ocean have a life history that i nvolves nesting on coastal beaches, foraging in neritic and oceanic habita ts, and long-distance migrations between and within these areas. The distribution of loggerh eads in the Indian Ocean is limited by the Asian

landmass to waters south of 30 o N latitude. In comparison to poten tial loggerhead distributions in southern waters of the Atlantic and Pacific, In dian Ocean distributions east and west are not restricted by landmasses south of approximately 38 o S latitude.

Historical accounts of loggerhead turtles in the Indian Ocean were given by Smith (1849), who described the species in South Africa, and De raniyagala (1933, 1939) who described Indian Ocean loggerheads within the subspecies C

. c

. gigas . Hughes (1974a) argued th at there was little justification for this separation. This work by Deraniyagala pr ovided evidence for the significant

historical distribution of loggerheads around Sri Lanka.

Loggerhead nesting in the Southwest Indian O cean includes the southeas tern coast of Africa from the Paradise Islands in Mozambique southw ard to St. Lucia in South Africa, and on the

south and southwestern coasts of Madagascar (Baldwin et al . 2003). Foraging habitats are only known for the Tongaland, South Af rica adult female loggerheads. Returns of flipper tags describe a range that extends eastward to Mada gascar, northward to Mozambique, Tanzania, and Kenya, and southward to Cape Agulhas at th e southernmost point of Africa (Baldwin et al

. 2003). Four post-nesting loggerhead s satellite tracked by Luschi et al . (2006) migrated northward, hugging the Mozambique coast and remained in shallow shelf waters off Mozambique for more than 2 months. Only one post-nesting female from the Southwest Indian Ocean population (South Africa) has been docum ented migrating north of the equator (to southern Somalia) (Hughes and Bartholomew 1996).

In the North Indian Ocean, Oman hosts the vast majority of loggerhead nests. The largest nesting assemblage is at Masirah Isla nd, Oman, in the northern tropics at 21 o N (Baldwin et al

. 2003). Other key assemblages occur on the Al Ha laniyat Islands, and on mainland beaches south

of Masirah Island to the Oman-Yemen borde r (IUCN - The World Conservation Union 1989a,24 1989b; Salm 1991; Salm and Salm 1991). In a ddition, nesting probably occurs on the mainland of Yemen on the Arabian Sea coast, and nesting has been confirmed on Socotra, an island off the coast of Yemen (Pilcher and Saad 2000). Baldwin et al . (2003) list other major nesting assemblages (> approximately 400 nesting female s/year) at Oman’s Arabian Sea Coast (17-20 o S) and Al Halaniyat Islands, Oman (17 o S).

Outside of Oman, loggerhead nesting is rare in the North Indian Ocean. The only verified nesting beaches for loggerheads on the Indian subcontinent are f ound in Sri Lanka (Deraniyagala 1939, Kar and Bhaskar 1982, Dodd 1988, Kapurusinghe 2006). Reports of regular loggerhead nesting on the Indian mainland ar e likely to be from misiden tifications of olive ridleys ( Lepidochelys olivacea ) (Tripathy 2005, Kapurusinghe 2006). Although loggerheads have been reported nesting in low numbers in Myanmar, these data may not be reliable because of misidentification of species (Thorbjarnarson et al . 2000).

Limited information exists on foraging locations of North Indian Ocean loggerheads. Foraging individuals have been reported off th e southern coastline of Oman (Salm et al . 1993) and in the Gulf of Mannar, between Sri Lanka and Indi a (Tripathy 2005, Kapurusinghe 2006). Satellite telemetry studies conducted in Oman have rev ealed new information on post-nesting migrations of loggerheads nesting on Masirah Island (Envi ronment Society of Oman and Ministry of Environment and Climate Change, Oman, unpublished da ta). Results reveal extensive use of the waters off the Arabian Peninsula, with the majori ty of telemetered turtles (15 of 20) traveling southwest, following the shoreline of southern Oman and Yemen, and circling well offshore in nearby oceanic waters. A minority traveled north as far as the western Persian (Arabian) Gulf (3 of 20) or followed the shoreline of southern Oman and Yemen as far west as the Gulf of Aden and the Bab-el-Mandab (2 of 20). These prelimin ary data suggest that post-nesting migrations and adult female foraging areas may be centere d within the region (E nvironment Society of Oman and Ministry of Environment and Climate Change, Oman, unpublished data). No tag returns or satellite tracks indicated that loggerheads nesting in Oman traveled south of the

equator.

In the East Indian Ocean, western Australia hosts all known loggerhead nesting (Dodd 1988). Nesting distributions in west ern Australia span from the Shark Bay World Heritage Area northward through the Ningaloo Marine Park coas t to the North West Cape and to the nearby Muiron Islands (Baldwin et al . 2003). Nesting individuals fro m Dirk Hartog Island have been recorded foraging within Shark Bay and Exmouth Gu lf, while other adults ra nge into the Gulf of Carpentaria (Baldwin et al . 2003). At the eastern extent of th is apparent range, there is possible overlap with loggerheads that ne st on Australia’s Pacific coast (L impus 2009). However, despite extensive tagging at principal nesting beaches on Australia’s Indian Ocean and Pacific coasts, no exchange of females between nesting beaches has been observed (Limpus 2009).

The available genetic information relates to c onnectivity and broad evol utionary relationships

between ocean basins. There is a lack of genetic information on population structure among

rookeries within the Indian Ocean. Bowen et al . (1994) described mtDN A sequence diversity among eight loggerhead nesting a ssemblages and found one of two principal branches in the Indo-Pacific basins. Using additional publishe d and unpublished data, Bowen (2003) estimated divergence between these two lineages to be approximately three million years. Bowen points25 out evidence for more recen t colonizations (12,000-250,000 year s ago) between the Indian

Ocean and the Atlantic-Mediterranean. For ex ample, the sole mtDNA haplotype (among eight samples) observed by Bowen et al . (1994) at Masirah Island, Oma n, is known from the Atlantic and suggests some exchange between oceans some 250,000 years ago. The other principal Indian Ocean haplot ype reported by Bowen et al . (1994) was seen in all loggerheads sampled (n=15) from Natal, South Africa. Encalada et al . (1998) reported that this haplotype was common throughout the North Atlantic and Medite rranean, thus suggesting a similar exchange

between Atlantic and Indian oceans as recently as 12,000 years ago (Bowen et al . 1994). Bowen (2003) speculated that Indian-A tlantic Ocean exchanges took place via the temperate waters south of South Africa and became rare as the ocean shifted to cold temperate conditions in this region.

In estimates of loggerhead gene flow in and out of the Indian Ocean, J.S. Reece (Washington University, personal communica tion, 2008) factored 100 samples from Masirah Island, 249 from Atlantic rookeries (from Encalada et al . 1998), and 311 from Pacific rookeries (from Hatase et
al . 2002a and Bowen et al . 1995). Reece used lineage coalescence methods to estimate that gene flow, expressed as number of effective migran ts, or exchanges of breeding females between

Indian Ocean rookeries and those fro m the Atlantic or Pacific occurre d at the rate of less than 0.1 migrant per generation. Reece estimated gene flow based on coalescence of combined mtDNA and nuclear DNA data to be approximately 0.5 migrants per generation. These unpublished results, while somewhat theoretical , may indicate that there is restri cted gene flow into and out of the Indian Ocean. The low level of gene flow mo st likely reflects the hist orical connectivity over geological timescales, rather than any contemporar y migration, and is consistent with Bowen’s hypothesis that exchange occurred most r ecently over 12,000-3,000,000 years ago, and has been restricted over recent ecological timescales.

The discrete status of three loggerhead groups in the Indian Ocean is primarily supported by observations of tag returns and satellite telemetry. The limited genetic information currently available based on mtDNA sequences does not allo w for a comprehensive analysis of genetic population structure analysis for I ndian Ocean rookeries, although Bowen et al . 1994 indicate the Oman and South African rookeries are genetical ly distinct, and once sequencing studies are completed for these rookeries, it is likely that they will also be distinct from the rookeries in western Australia. Based on multiple lines of evid ence, discrete status is supported for the North Indian Ocean, Southwest Indian Ocean, and Sout heast Indo-Pacific Ocean. Although there is not a sufficiently clear picture of gene flow betw een these regions, we propose that significant

vicariant barriers exist between these three Indian Ocean groups that would prevent migration of

individuals on a time scale relativ e to management and conservati on efforts. These vicariant barriers are the oceanographic pheno mena associated with Indian Ocean equatorial waters, and the large expanse between continents in the South Indian Ocean w ithout suitable benthic foraging habitat.

Given the information presented above, the BRT has unanimously concluded that three discrete population segments exist in the Indian Ocean as a consequence of ecological, behavioral, and oceanographic factors, and given the genetic ev idence: (1) North Indian Ocean, (2) Southwest Indian Ocean, and (3) Southeast Indo-Pacific Ocean.26 Atlantic Ocean and Mediterranean Sea Within the Atlantic Ocean, loss and re-colonizati on of nesting beaches over evolutionary time scales has been determined by climate, nata l homing, and rare dispersal events (Encalada et al

. 1998, Bowen and Karl 2007). At times, temperate beaches were too cool to incubate eggs and nesting could have continued only on tropical be aches. Thus, the contemporary distribution of nesting is the product of colonization events from the tropical refugi a during the last 12,000 years. Apparently, turtles from the Northwest Atlantic colonized the Mediterranean and at least two matrilines were involved (Schroth et al . 1996); these rookeries became isolated from the Atlantic populations in th e last 10,000 years (Encalada et al . 1998). A similar colonization event appears to have populated the Northeast Atlant ic (C. Monzón-Argüello, Instituto Canario de Ciencias Marinas - Spain, personal communication, 2008).

Nesting in the western South Atla ntic occurs primarily along the mainland coast of Brazil from Sergipe south to Rio de Janeiro, with peak concen trations in northern Bahia, Espírito Santo, and northern Rio de Janeiro (Marcovaldi and Chaloupka 2007). In the eastern South Atlantic, diffuse nesting may occur along the mainland coast of Africa (Fretey 2001), with more than 200 loggerhead nests reported for Rio Longa beach in central Angola in 2005 (Brian 2007). However, other researchers have been unable to confirm nesting by loggerheads in the last decade anywhere along the south Atlantic coast of Africa, including Angola (Fretey 2001, Weir

et al . 2007). There is the possibility that repo rts of nesting loggerheads from Angola and Namibia (Márquez M. 1990, Brian 2007) may have ar isen from misidentifie d olive ridley turtles (Brongersma 1982, Fretey 2001). At the current time, it is not possible to confirm that regular, if any, nesting of loggerheads occu rs along the Atlantic coast of Africa, south of the equator.

Genetic surveys of loggerheads have revealed th at the Brazilian rookeri es have a unique mtDNA haplotype (Encalada et al . 1998, Pearce 2001). The Brazilian mtDNA haplotype, relative to North Atlantic haplotypes, indica tes isolation of South Atlantic loggerheads from North Atlantic loggerheads on a scale of 0.25-0.5 million years ago, and microsatellite DNA results show divergence on the same time scale (Bowen 2003). Brazil’s unique haplotype has been found only in low numbers in foraging pop ulations of juvenile loggerhead s of the North Atlantic (Bass

et al . 2004). Other lines of evidence support a de ep division between loggerheads from the South Atlantic and from the North Atlantic, includ ing: (1) a nesting seas on in Brazil that peaks in the austral summer around December/January (Marcovaldi and Laurent 1996), as opposed to the May-August nesting season in the southeast U.S. in the nor thern hemisphere (Witherington et
al . 2009); and (2) no observations of tagged logg erheads moving across the equator in the Atlantic, except a single case of a captive-reared animal that wa s released as a juvenile from Espírito Santo and was recaptured 3 years later in the Azores (Bolten et al . 1990). Post-nesting females from Espírito Santo moved either north or south along the coast, but remained between 10°S and 30°S (Projeto TAMAR, unpublished data).

Relatively little is known about the at-sea beha vior of loggerheads originating from nesting beaches in Brazil. Recaptures of tagged juveni les and nesting females ha ve shown movement of animals up and down the coast of South America (Almeida et al . 2000, Marcovaldi et al . 2000, Laporta and Lopez 2003, Almeida et al . 2007). Juvenile loggerheads, presumably of Brazilian origin, have also been captured on the high seas of the South Atlantic (Kotas et al . 2004, Pinedo and Polacheck 2004) and off the coast of Atlantic Africa (Petersen 2005, Petersen et al . 2007,27 Weir et al . 2007) suggesting that, like th eir North Pacific and Northwest Atlantic counterparts, loggerheads of the South Atlantic may undertake transoceanic developmental migrations (Bolten

et al . 1998, Peckham et al . 2007). There is also the possibility that some loggerhead individuals from the Southwest Indian Ocean population se gment are swept around the Cape of Good Hope by the Benguela Current (Baldwin et al . 2003), into the Southwest A tlantic Ocean, where they overlap with loggerheads from the S outh Atlantic Ocean population segment.

The mean size of reproductive female loggerh eads in Brazil is 92.9 cm SCL, which is comparable to the size of nesters in the Northw est Atlantic, but larger than nesters of the

Northeast Atlantic and Mediterranean (Tiwari and Bjornda l 2000, Margaritoulis et al . 2003,

Varo Cruz et al . 2007). Egg size and mass of Brazilian l oggerheads are smaller than those from the Northwest Atlantic, but larg er than those of the Mediterra nean (Tiwari and Bjorndal 2000).

Within the Northwest Atlantic, the majority of nesting activity o ccurs from April through September, with a peak in June and July (Williams-Walls et al . 1983, Dodd 1988, Weishampel et
al . 2006). Nesting occurs within the Northwest A tlantic along the coasts of North America, Central America, northern South America, the Antill es, and The Bahamas, but is concentrated in the southeastern U.S. and on the Yucatán Peninsula in Mexico (Sternberg 1981, Ehrhart 1989,

Ehrhart et al . 2003, NMFS and FWS 2008). Many nesting beaches within the Northwest Atlantic have yet to be sampled for genetic an alysis. Five recovery units (subpopulations) have been identified based on genetic differences a nd a combination of geographic distribution of nesting densities and geographic separation. These recovery units are: Northern Recovery Unit (Florida/Georgia border through southern Virginia), Peninsular Florida Recovery Unit (Florida/Georgia border through Pinellas County, Fl orida), Northern Gulf of Mexico Recovery Unit (Franklin County, Florida, through Texas) , Greater Caribbean Recovery Unit (Mexico through French Guiana, The Bahamas, Lesser Antil les, and Greater Antilles), and Dry Tortugas Recovery Unit (islands located west of Key West, Florida) (NMFS and FWS 2008). There is limited exchange of nesting females among these recovery units (Encalada et al . 1998, Foote et
al . 2000, J. Richardson personal communi cation cited in NMFS 2001, Hawkes et al . 2005). Based on the number of haplotypes, the highest le vel of loggerhead mtDNA genetic diversity in

the Atlantic has been observed in females of the Greater Caribbean Recovery Unit that nest at Quintana Roo, Mexico (Encalada et al . 1999, Nielsen et al . in press). Howeve r, genetic diversity should be evaluated further using haplotype and nucleotide diversit y calculated similarly for each recovery unit. Genetic data are not availabl e for all the nesting a ssemblages in the region, including a key nesting assemblage in Cuba . New genetic markers have recently been developed, including primers that produce addi tional mtDNA sequence data (Abreu-Grobois et
al . 2006, LeRoux et al . 2008), and an array of micros atellite m arkers (Shamblin et al . 2008) that will enable finer resolution of population boundaries.

Loggerheads in the Northwest A tlantic display complex popula tion structure based on life history stages. Based on mtDNA, oceanic juveniles show no struct ure, neritic j uveniles show moderate structure, and nesting colo nies show strong structure (Bowen et al . 2005); however, a recent study suggests some structure may exist among oceanic juveniles, at least in the Northeast Atlantic oceanic foraging grounds (Monzón-Argüello et al . 2009), and is related to the latitudinal

distributions of the Northwest A tlantic rookeries. In contrast, a survey using microsatellite (nuclear) markers showed no significant populat ion structure among nes ting populations (Bowen28 et al . 2005), indicating that while females exhibit strong philo patry, males may provide an avenue of gene flow between nesting colonies in this region. However, the power to detect structure with the nuclear markers used in this study may have been limited due to the few markers used and small sample sizes. Additiona l work with larger sample sizes and better nuclear markers is underway (B. Shamblin, University of Georgia, personal communication, 2008). Nevertheless, Bowen et al . (2005) argued that male-mediated gene flow within the

Northwest Atlantic does not detr act from the classification of breeding areas as independent populations (e.g., recovery units) because the pr oduction of progeny depends on female nesting

success. All Northwest Atlantic recovery units are reproductivel y isolated from populations within the Northeast Atlantic, Sout h Atlantic, and Mediterranean Sea.

As oceanic juveniles, loggerheads from the Northwest Atlantic use the North Atlantic gyre and

often are associated with Sargassum communities (Carr 1987); th ey also are found in the Mediterranean Sea. In these areas, they overlap with animals originat ing from the Northeast Atlantic and the Mediterranean Sea (Laurent et al . 1993, 1998; Bolten et al . 1998; Bowen et al

. 2005; LaCasella et al . 2005; Carreras et al . 2006, Monzón-Argüello et al . 2006; Revelles et al

. 2007). In the western Mediterranean , they tend to be associated with the waters off the northern

African coast and the northeastern Balearic archipelago, areas gene rally not inhabited by turtles of Mediterranean origin (Carreras et al . 2006, Revelles et al . 2007, Eckert et al . 2008). As larger, neritic juveniles, they s how more structure and tend to inha bit areas closer to their natal origins (Bowen et al . 2004), but some do move to and from oceanic foraging grounds throughout this life stage (Mansfield 2006, McClellan and Read 2007), and some continue to use the Mediterranean Sea (Casale et al . 2008b, Eckert et al . 2008). Adult populations are highly structured with no overlap in distribution among adult loggerheads from the Northwest Atlantic, Northeast Atlantic, South Atlantic, and Medite rranean. Carapace epibionts suggest the adult females of different subpopulations use differe nt foraging habitats (Caine 1986). In the Northwest Atlantic, based on satellite telemetry studies and flipper tag returns, non-nesting adult females from the Northern Recovery Unit resi de primarily off the east coast of the U.S.; movement into the Bahamas or the Gulf of Mexico is rare (Bell and Richardson 1978, Williams and Frick 2001, Mansfield 2006, Turtle Expert Working Group 2009). Adult females of the Peninsular Florida Recovery Unit are distribu ted throughout eastern Florida, the Bahamas, Greater Antilles, Cuba, the Yucat án Peninsula, and the Gulf of Mexico, as well as along the

Atlantic seaboard of th e U.S. (Meylan 1982, Meylan et al . 1983, Foley et al . 2008; Turtle Expert Working Group 2009). Adult females from the Northern Gulf of Mexico Recovery Unit

rem ained in the Gulf of Mexico, including off th e Yucatán Peninsula, based on satellite telemetry and flipper tag returns (Foley et al . 2008; Turtle Expert Working Group 2009; M. Lamont, Florida Cooperative Fish and Wildlife Res earch Unit, personal communication, 2009; M. Nicholas, National Park Servi ce, personal communication, 2009).

Nesting in the Northeast Atlantic is concentrated in the Cape Verde Archipelago, with some nesting occurring on most of the islands, and the highest concentration on the beaches of Boa Vista Island (López-Jurado et al . 2000, Varo Cruz et al . 2007, Loureiro 2008). On mainland Africa, there is minor nesting on the coasts of Mauritania to Senega l (Brongersma 1982, Arvy et
al . 2000, Fretey 2001). Earlier reports of loggerh ead nesting in Morocco (Pasteur and Bons 1960) have not been confirmed in recent years (Tiwari et al . 2001). Nesting has not been reported from Macaronesia (Azores, Madeira Archipelago, The Selvagens Islands, and the29 Canary Islands), other than in the Cape Verde Archipelago (Brongersma 1982). In Cape Verde, nesting begins in mid June and extends into October (Cejudo et al . 2000), which is somewhat later than when nesting occurs in the Northwest Atlantic.

Based on an analysis of mtDNA of 196 nesting females from Boa Vista Island, the Cape Verde nesting assemblage is genetica lly distinct from other studie d rookeries (C. Monzón-Argüello, Instituto Canario de Ciencias Marinas - Sp ain, personal communicati on, 2008; Monzón-Argüello

et al . 2009). The results also indica te that despite the close prox imity of the Mediterranean, the Boa Vista rookery is most closely related to the rookeries of th e Northwest Atlantic.

The distribution of juveniles from the Northeast Atlantic is largely unknown but they have been found on the oceanic foraging grounds of the North Atlantic (A. Bolten, University of Florida, personal communication, 2008, based on Bolten et al . 1998 and LaCasella et al . 2005; MonzónArgüello et al . 2009; M. Tiwari, NMFS, and A. Bolten, Un iversity of Florida, unpublished data)

and in the western and central Mediterranean (A. Bolten, University of Florida, personal communication, 2008, based on Carreras et al . 2006), along with small juveniles from the Northwest Atlantic. Loggerheads of both juvenile and adult si ze also have been reported captured incidental to fishing operations off th e coasts of Europe and Africa and have been reported stranded and in the ma rkets (Maigret 1983; Tiwari et al . 2001; Benhardouze 2004;

Benhardouze et al . 2004; Duguy et al . 2004, 2005; Witt et al . 2007). The size of nesting females in the Northeast Atlantic is comparable to t hose in the Mediterranean (average 72-80 cm SCL; Margaritoulis et al . 2003) and smaller than those in the Nort hwest Atlantic or the South Atlantic; 91% of the nesting turtles are < 86.5 cm CCL (Hawkes et al . 2006) and nesting females average 77.1 cm SCL (Cejudo et al . 2000). Post-nesting females demons trated two behaviors, depending on size (Hawkes et al . 2006). The larger tu rtles (>93 cm) foraged in coastal waters along northwest Africa and the smaller animals (<87 cm) foraged oceanically, mostly between Cape

Verde and the African shelf from Mauritania to Guinea Bissau. Loureiro (2008) only observed the smaller size class of reproductive female loggerheads nesting on Santiago Island, Cape

Verde.

Nesting occurs throughout the central and eastern Mediterranean on the shores of Italy, Greece, Cyprus, Turkey, Syria, Lebanon, Israel, the Sina i, Egypt, Libya, and Tunisia (Sternberg 1981, Margaritoulis et al . 2003, SWOT 2007). Sporadic nesting al so has been reported in the western Mediterranean on Corsica (Delaugerre and Cesarini 2004), southwestern Italy (Bentivegna et al

. 2005), and on the Spanish Mediterranean coast (Tomás et al . 2003, 2008). Nesting in the

Mediterranean is concentrated between June and early August (Margaritoulis et al . 2003).

W ithin the Mediterranean, a recent study of mitochondrial and nuclear DNA in nesting assemblages from Greece to Israel indicated genetic structuring, philopatry by both females and males, and limited gene flow between assemblages (Carreras et al . 2007). Genetic

differentiation based on mtDNA indi cated that there are at leas t four independent nesting subpopulations within the Mediterranean and us ually they are characterized by a single haplotype: (1) mainland Greece and the adjoining I onian Islands, (2) eastern Turkey, (3) Israel, and (4) Cyprus. There is no evidence of adult female exchange among these four subpopulations (Carreras et al . 2006). In studies of the foraging grounds in the western and central Mediterranean, seven of the 17 distinct hapl otypes detected had not yet been described,30 indicating that nesting be ach data to describe the natal origins of juveniles exploiting the western Mediterranean Sea are incomplete (Carreras et al . 2006, Casale et al . 2008b). Gene flow among

the Mediterranean rookeries estimated from nuclear DNA was significan tly higher than that calculated from mtDNA, consistent with the scenar io of female philopatr y maintaining isolation between rookeries, offset by male-mediated gene flow. Nevertheless, the nuclear data show there was a higher degree of subs tructuring among Mediterranean r ookeries compared to those in the Northwest Atlantic (Bowen et al . 2005; Carreras et al . 2007).

Small oceanic juveniles from the Mediterranean Sea use the eastern basin (defined as inclusive of the central Mediterranean, Ioni an, Adriatic, and Aegean Seas) and the western basin (defined as inclusive of the Tyrrhenian Sea) along the European coast (Laurent et al . 1998, Margaritoulis

et al . 2003, Carreras et al . 2006, Revelles et al . 2007). Larger juvenile s also use the eastern Atlantic and the eastern Mediterranean, especially the Tunisia-Libya shelf and the Adriatic Sea (Laurent et al . 1993, Margaritoulis et al . 2003, Monzón-Argüllo et al . 2006, Revelles et al

. 2007). Adults appear to forage closer to the nesting beaches in the eastern basin; most tag recoveries from females nesting in Greece have occurred in the Adriatic Sea and off Tunisia (Margaritoulis et al . 2003, Lazar et al . 2004).

Straight carapace lengths of loggerheads nesting in the Mediterranean were significantly smaller than the lengths of loggerheads nesting in the Northwest Atlantic and the South Atlantic, and

within the Mediterranean showed some vari ability among subpopulations. Carapace lengths ranged from 58 to 95 cm SCL (Margaritoulis et al . 2003). Greece’s loggerheads averaged 77-80 cm SCL (Tiwari and Bjor ndal 2000, Margaritoulis et al . 2003), whereas Turkey’s loggerheads averaged 72-73 cm SCL (Margaritoulis et al . 2003). The Greece turtles also produced larger clutches (relative to body size) than those pr oduced by Florida or Braz il nesters (Tiwari and Bjorndal 2000). The authors suggested that sea turtles in the Mediterranean encounter environmental conditions significantly diffe rent from those experienced by populations

elsewhere in the Atlantic Ocean basin.

Given the information presented above, the BRT has unanimously concluded that four discrete population segments exist in the Atlantic Oc ean/Mediterranean Sea as a consequence of ecological, behavioral, and oceanographic factors, and based on genetic evidence: (1) Northwest Atlantic Ocean, (2) Northeast Atlantic Ocean, (3 ) South Atlantic Ocean, and (4) Mediterranean Sea. These population segments ar e genetically distinct (Bowen et al . 1994; Encalada et al

. 1998; Pearce 2001; Carerras et al . 2007; C. Monzón-Argüello, Inst ituto C anario de Ciencias Marinas - Spain, personal communication, 2008; Monzon-Arguello et al . 2009) and, although they may comingle on oceanic foraging grounds as juve niles, adults apparently are isolated from each other; they also differ demographically.

3.1.2. Significance Determination

In Section 3.1.1., the BRT identified nine discrete population segm ents. As described below by ocean basin, each of the nine discrete populat ion segments is biologically and ecologically significant. They each represent a large porti on of the species range, sometimes encompassing an entire hemispheric ocean basin. The range of each discrete population segment represents a unique ecosystem, influenced by local ecological and physical factors. The loss of any single31 discrete population segment would result in a significant gap in th e loggerhead’s range. Each discrete population segment is genetically unique, often identif ied by unique mtDNA haplotypes, and the BRT believes that these unique haplotypes could represent adaptive differences; the loss of any one discrete population segment would repr esent a significant loss of genetic diversity. Therefore, the BRT concludes that these nine population segments are both discrete from other conspecific population segments and significan t to the species to which they belong, Caretta

caretta

.

The geographic delineations given below for each discrete population segment were determined primarily based on nesting beach locations, gene tic evidence, oceanographic features, thermal tolerance, fishery bycatch data, and informati on on loggerhead distribution and migrations from satellite telemetry and flipper tagging studies. With rare exception, adults from discrete population segments remain within the delineated boundaries. In some cases, juveniles from two or more discrete population segments may mix on foraging areas and therefore, their distribution and migrations may extend beyond the geographic boundaries delineated below for each discrete population segment (e.g., juveniles from the Nort hwest Atlantic Ocean, Northeast Atlantic Ocean, and Mediterranean Sea disc rete population segments share foraging habitat in the western Mediterranean Sea).

Pacific Ocean The BRT considers 60oN latitude and the equa tor as the north-south boundaries of the North Pacific population segment based on oceanographi c features, loggerhead sightings, thermal tolerance, fishery bycatch data, and informa tion on loggerhead distribution from satellite telemetry and flipper tagging studies. The lo ss of the North Pacific population segment would result in a significant gap in th e range of the taxon. There is no evidence or reason to believe that female loggerheads from South Pacific nest ing beaches would repopulate the North Pacific nesting beaches should those nesting assemblages be lost. Tagging studies show that the vast majority of nesting females return to the sa me nesting area. As summarized by Hatase et al

. (2002a), of 2,219 tagged nesting females from Japa n, only five females relocated their nesting sites. In addition, flipper tag a nd satellite telemetry research, as described in detail in Section 3.1.1., shows no evidence of north-south movement of loggerheads across the equator. The BRT

concludes that the North Pacifi c population segment is both disc rete from other conspecific population segments and significant to the taxon to which it belongs, and therefore, it satisfies the DPS policy criteria.

The BRT considers the equator and 60oS la titude as the north-south boundaries and 67 o W longitude and 139oE longitude as the east-west boundaries of the South Pacific population segment based on oceanographic features, loggerh ead sightings, thermal tolerance, fishery bycatch data, and information on loggerhead dist ribution from satellite telemetry and flipper tagging studies. The loss of the South Pacific population segment would result in a significant gap in the range of the taxon. Th e South Pacific population is th e only population of loggerheads found south of the equator in the Pacific Ocean and there is no evidence or reason to believe that female loggerheads from North Pacific nesti ng beaches would repopulate the South Pacific nesting beaches should those nesting assemblages be lost. In addition, fl ipper tag and satellite telemetry research, as described in detail in Section 3.1.1., shows no evidence of north-south movement of loggerheads across th e equator. We also do not e xpect that recolonization from32 Indian Ocean loggerheads would occur in easter n Australia within eco logical time frames. Despite evidence of foraging in the Gulf of Carpenta ria by adult loggerheads from the nesting populations in eastern Australia (South Pacific population se gment) and western Australia (Southeast Indo-Pacific Ocean popul ation segment), the nesting females from these two regions are considered to be genetically distinct fr om one another (Limpus 2009). In addition to a substantial disparity in mtDNA haplotype frequencies between these two populations, FitzSimmons (University of Canberra, unpublished data) found significant differences in nuclear

DNA microsatellite loci between females nesting in these tw o regions, indicating separation

between the South Pacific Ocean and the Sout heast Indo-Pacific Ocean population segments. Long term studies show a high degree of site fide lity by adult females in the South Pacific, with most females returning to the same beach with in a nesting season and in successive nesting seasons (Limpus 1985, 2009; Limpus et al . 1994). This has been docum ented as characteristic of loggerheads from various rookeries throughout the world (Schroeder et al . 2003). The BRT

concludes that the South Pacifi c population segment is both disc rete from other conspecific population segments and significant to the taxon to which it belongs, and therefore, it satisfies the DPS policy criteria.

Indian Ocean

The BRT considers 30 o N latitude and the equator as the north-south boundary of the North Indian Ocean population segment based on o ceanographic features, loggerhead sightings, thermal tolerance, fishery bycatch data, and info rmation on loggerhead distribution from satellite telemetry and flipper tagging studies. The lo ss of the North Indian Ocean population segment would result in a significant gap in the range of the ta xon. Genetic studies ha ve shown that adult populations are highly structured with no overlap in distribution among a dult loggerheads in the North Indian Ocean, Southwest Indian Ocean, and Southeast Indo-Pacific. There is no evidence or reason to believe that female loggerheads from the Southwest Indian O cean or Southeast IndoPacific would repopulate the North Indian Ocean nesting beaches should those populations be

lost. The BRT concludes that the North Indian Ocean population segment is both discrete from other conspecific population segments and signi ficant to the taxon to which it belongs, and therefore, it satisfies the DPS policy criteria.

The BRT considers the equator and 60oS lati tude as the north-sout h boundaries and 20oE longitude at Cape Agulhas on the southern tip of Africa and 80oE longitu de as the east-west

boundaries of the Southwest Indi an Ocean population segment ba sed on oceanographic features, thermal tolerance, fishery bycatch data, and info rmation on loggerhead distribution from satellite telemetry and flipper tagging studies. The loss of the Southwest Indian Ocean population segment would result in a significa nt gap in the range of the ta xon. Genetic studies have shown that adult populations are highly structured in the North Indian Ocean, Southwest Indian Ocean, and Southeast Indo-Pacific. Only one post-nest ing female from the Southwest Indian Ocean population (South Africa) has been documented mi grating north of the equator (to southern Somalia) (Hughes and Bartholomew 1996). There is no evidence or reason to believe that female loggerheads from the North Indian Ocea n or Southeast Indo-Paci fic would repopulate the Southwest Indian Ocean nesting beaches should those populations be lost. There is also no

evidence of movement of adult Southwest Indian Ocean loggerheads west of 20oE longitude at Cape Agulhas, the southernmost point on the Afri can continent or east of 80oE longitude within

the Indian Ocean. The BRT concludes that th e Southwest Indian Ocean population segment is33 both discrete from other consp ecific population segments and signi ficant to the taxon to which it belongs, and therefore, it satis fies the DPS policy criteria.

The BRT considers the equator and 60oS lati tude as the north-south boundaries and 139oE longitude and 80oE longitude as the east-west boundaries of the South east Indo-Pacific Ocean population segment based on oceanographic features , thermal tolerance, fishery bycatch data, and information on loggerhead distribution from sa tellite telemetry and flipper tagging studies. The loss of the Southeast Indo-Pacific Ocean pop ulation segment would result in a significant gap in the range of the taxon. Genetic studies have shown that adult populations are highly structured with no overlap in distribution among adult loggerheads in the North Indian Ocean, Southwest Indian Ocean, and Southe ast Indo-Pacific. There is no evidence or reason to believe that female loggerheads from the North Indi an Ocean or Southwest Indian Ocean would repopulate the Southeast Indo-Pacific nesting beaches should those populations be lost. There is

also no evidence of movement of adult Southeas t Indo-Pacific Ocean loggerheads west of 80oE longitude within the Indian Ocean. Despite evid ence of foraging in the Gulf of Carpentaria by adult loggerheads from the nesting populations in eastern Australia (S outh Pacific population segment) and western Australia (Southeast Indo-Paci fic Ocean population segment), the nesting females from these two regions are considered to be genetically distinct from one another (Limpus 2009). In addition to a substantial di sparity in mtDNA haplotype frequencies between these two regions, FitzSimmons (University of Canberra, unpublished data) found significant

differences in nuclear DNA microsatellite loci from females nesting in these two regions, indicating separation between the South Pacifi c population segment and the Southeast IndoPacific Ocean population segment. The BRT conc ludes that the Southeast Indo-Pacific Ocean population segment is both discrete from other conspecific population segments and significant to the taxon to which it belongs, and theref ore, it satisfies the DPS policy criteria.

Atlantic Ocean and Mediterranean Sea

The BRT considers 60 o N latitude and the equator as th e north-south bou ndaries and 40oW longitude as the east boundary of the Northwest Atlantic Oc ean population segment based on oceanographic features, loggerhead sightings, th ermal tolerance, fish ery bycatch data, and information on loggerhead distribution from satelli te telemetry and flipper tagging studies. The loss of the Northwest Atlantic population segment would result in a signif icant gap in the range

of the taxon. Genetic studies have shown that adult populations are high ly structured with no overlap in distribution among adu lt loggerheads from the Northwest Atlantic, Northeast Atlantic, South Atlantic, and Mediterranean. There is no evidence or reason to believe that female loggerheads from the Northeast Atlantic, Mediterra nean Sea, or South Atlantic nesting beaches would repopulate the Northwest Atla ntic nesting beaches should these populations be lost. Data from satellite telemetry studies and flipper tag re turns show that the vast majority of nesting

females from the Northwest Atlantic return to th e same nesting area; they reveal no evidence of movement of adults south of the equator or eas t of 40oW longitude. The BRT concludes that the Northwest Atlantic population segment is both discrete from other conspecific population segments and significant to the taxon to which it belongs, and therefore, it satisfies the DPS policy criteria.

The BRT considers 60 o N latitude and the equator as th e north-south bou ndaries and 40oW longitude as the west boundary of the Northeas t Atlantic Ocean population segment. The BRT34 considers the boundary between th e Northeast Atlantic Ocean a nd Mediterranean Sea population segments as 5o 36’W longitude (Strait of Gibraltar). These boundaries are based on oceanographic features, loggerhead sightings, th ermal tolerance, fish ery bycatch data, and information on loggerhead distribution from satelli te telemetry and flipper tagging studies. The loss of the Northeast Atlantic population segm ent would result in a significant gap in the range of the taxon. Genetic studies have shown th at adult populations are highly structured with no overlap in distribution among adult loggerheads from the No rthwest Atlantic, Northeast Atlantic, South Atlantic, and Mediterranean. Th ere is no evidence or reason to believe that female loggerheads from the Northwest Atlantic, Mediterranean Sea, or South Atlantic nesting beaches would repopulate the Northeast Atlantic nesting beaches should these populations be

lost. There is also no evidence of movement of Northeast Atlantic adults west of 40oW longitude

or east of the Strait of Gibraltar (5o 36’W longitude). The BRT concludes that the Northeast Atlantic population segment is both discrete fr om other conspecific population segments and significant to the taxon to which it belongs, and therefore, it sa tisfies the DPS policy criteria.

The BRT considers the Mediterranea n Sea west to 5o 36’W longitude (Strait of Gibraltar) as the boundary of the Mediterranean Sea populati on segment based on oceanographic features, loggerhead sightings, thermal tolerance, fisher y bycatch data, and information on loggerhead distribution from satellite telemetry and flipper tagging studies. The loss of the Mediterranean Sea population segment would result in a signific ant gap in the range of the taxon. Genetic studies show that adult populati ons are highly structured with no overlap in distribution among adult loggerheads from the Nort hwest Atlantic, Northeast At lantic, South Atlantic, and

Mediterranean. There is no evidence or reason to believe that female loggerheads from the Northwest Atlantic, Northeast At lantic, or South Atlantic nes ting beaches would repopulate the Mediterranean Sea nesting beaches should these populations be lost . As previously described, adults from the Mediterranean Sea population segm ent appear to forage closer to the nesting beaches in the eastern basin, and most flipper tag recoveries from females nesting in Greece have

occurred in the Adriatic Sea and off Tunisia (Margaritoulis et al . 2003, Lazar et al . 2004). There

is no evidence of moveme nt of adult Mediterranean Sea logg erheads west of the Strait of Gibraltar (5o 36’W longitude). The BRT conc ludes that the Mediterranean Sea population segment is both discrete from other conspecifi c population segments a nd significant to the taxon to which it belongs, and therefore, it satisfies the DPS policy criteria.

The BRT considers the equator and 60 o S latitude as the north -south boundaries and 20oE longitude at Cape Agulhas on th e southern tip of Africa and 67 o W longitude as the east-west boundaries of the South Atlantic Ocean populat ion segment based on oceanographic features, loggerhead sightings, thermal tolerance, fisher y bycatch data, and information on loggerhead

distribution from satellite teleme try and flipper tagging studies. The loss of the South Atlantic population segment would result in a significant gap in the range of the taxon. Genetic studies show that adult populations are highly structured with no overl ap in distribution among adult loggerheads from the Northwest Atlantic, Northeas t Atlantic, South Atlantic, and Mediterranean.

There is no evidence or reason to believe that female loggerheads from the Northwest Atlantic, Northeast Atlantic, or Mediterranean Sea nesti ng beaches would repopulate the South Atlantic nesting beaches should these populations be lost. The BRT concludes that the South Atlantic population segment is both discrete from other conspecific population segments and significant to the taxon to which it belongs, and theref ore, it satisfies the DPS policy criteria.35 In summary, the BRT has identified nine logge rhead DPSs distributed globally: (1) North Pacific Ocean DPS, (2) South P acific Ocean DPS, (3) North Indi an Ocean DPS, (4) Southeast Indo-Pacific Ocean DPS, (5) Southwest Indian O cean DPS, (6) Northwest Atlantic Ocean DPS, (7) Northeast Atlantic Ocean DPS, (8) Mediterr anean Sea DPS, and (9) South Atlantic Ocean

DPS.

3.2. Significant Portion of it s Range (SPOIR) Assessment

The BRT determ ined that the range of each DPS c ontributes meaningfully to the conservation of the DPS and that populations that may contribute more or less to the conservation of each DPS throughout a portion of its range ca nnot be identified due to the highly migratory nature of the listed entity.

The loggerhead sea turtle is highly migratory and crosses multiple domestic and international geopolitical boundaries. Depending on the life stage, they may occur in oceanic waters or along

the continental shelf of landma sses, or transit back and fort h between oceanic and neritic habitats. Protection and management of both the te rrestrial and marine envi ronments is essential to recovering the listed entity. Management me asures implemented by any State, foreign nation, or political subdivision likely would only affect indi vidual sea turtles duri ng certain stages and seasons of the life cycle. Management measur es implemented by any State, foreign nation, or political subdivision may also affect individu als from multiple DPSs because juveniles from disparate DPSs can overlap on foraging grounds or migratory corridors (e.g., Northwest Atlantic, Northeast Atlantic, and Mediterranean Sea DPSs). The “significant” term in SPOIR refers to the contribution of the population(s) in a portion of the range to the c onservation of the listable entity being considered. The BRT was unable to s ubdivide the geographic range of loggerhead populations below the DPS level in which threats and conservation efforts specific to any portion of its range can be identified and analyzed for listing status without di sregarding the impacts of

these threats and conservation efforts to th e listed entity throu ghout all of its range.36 

## SECTION 4—ASSESSMENT OF EXTINCTION RISK

4.1. Description of Extinction Risk Assessment Approaches

To assess extinction risks of loggerhead DPSs, we conducted two independent analyses. The first analysis used the diffusion approximation a pproach based on time series of counts of nesting females (Lande and Orzack 1988, Dennis et al . 1991, Holmes 2001, Snover and Heppell 2009). This analysis provided a metric (susceptibility to quasi-extinction or SQE) to determine if the probability of a population’s risk of quasi-extinction is high enough to warrant a particular status listing (Snover and Heppell 2009). This approach is based on stoc hastic projections of observed trends and variability in the numbers of mature females at various nesting beaches. The second approach used a deterministic stage-based popul ation model that focused on determining the effects of known anthropogenic mo rtalities on each DPS with resp ect to the vital rates of the species. Anthropogenic mortalitie s were added to natural morta lities and possible ranges of population growth rates were computed as anothe r metric of population health. Because this approach is based on matrix mode ls, we call it a threat matrix an alysis. This approach focused on how additional mortalities may af fect the future growth and recovery of a loggerhead turtle DPS. The first approach (SQE) was solely based on the available time-series data on the numbers of nests at nesting beaches, whereas th e second approach (threat matrix analysis) was based on the known biology of the species and anthropogenic mortalities independent of observed nesting beach data.



### 4.2. Computation of Susceptibility to Quasi-extinction (SQE)

4.2.1. Methods

Estim ates of quasi-extinction ri sk are known to have high degrees of uncertainty due to the stochastic nature of populations and their environments and the e rror involved in data collection and subsequent parameter estimation (Holmes et al . 2007). However, there are ongoing needs for management to classify populations in terms of their status when only limited data, often with high observation error rates, are available. To address this need, Snover and Heppell (2009) presented a quasi-extinction risk index called susceptibility to quasi -extinction (SQE) that can be used to classify sea turtle nesting populations based on relative risks using only nesting beach census data. This index integrates parameter unc ertainty and stochasticity in extinction risk forecasting while allowing managers to balance the risk of making Type I (considering a population to be not at risk when it is) and Type II (considering a population to be at risk when it is not) errors when making decisi ons. While they showed that th e method is robust in assessing actual risk (in terms of a binary metric of “at risk” or “not at risk”) using population simulations, they clarify that SQE values are not indicative of a true probability of quasi-extinction because they assume constant distributions of trend and variance over long time frames (three generations) and ignore density dependence. Rather , the index serves as a tool for classifying populations by relative status.

The technique involves standard methods of diffusion approximation (Lande and Orzack 1988,

Dennis et al . 1991). These methods are based on a de nsity-independent e xponential model in a randomly varying environment:37 ) exp( 1 ε μ + = + tt NN , (1)

where ) ,0(~ 2 σε N

, N is the population size, and t is time (Dennis et al . 1991, Holmes 2001).
Assuming that the lognormal distribution can be used to compute the probability that the population will be of a certain size, two parameters estimated by this method are μ , the

arithmetic mean of the log population growth rate, and 2 σ , the variance of the log population growth rate, which accounts for sources of vari ability, including environmental and demographic stochasticity, and observation error (Holmes 2001, Morris and Do ak 2002 - Chapter 5). These parameters were estimated using the regression analysis, whereas the confidence intervals were estimated using the method of Dennis et al . (1991) and Morris and Doak (2002). While not as statistically robust as the methods presente d by Holmes (2001), the methods of Dennis et al

. (1991) and Morris and Doak (2002) are simpler and more tractable for managers. For the lengths of time series typically available for s ea turtle populations (i.e., <20 years), Snover and Heppell (2009) demonstrated that both methods performed equally well in determining SQE. These estimated parameters were used to make inferences on total popul ation growth rates and quasi-extincti on probability.

The data available for this analysis were time seri es of annual counts of nest s or nesting females. When the data were in units of nests per yr, we converted these numbers to estimates of nesting females (see Section 4.2.2 for details). For each nesting beach or region, we followed the recommendations in Snover and Heppell (2009), a nd used a running-sum of 3 yr. This data smoothing provides a more accurate reflection of population change by re ducing year-to-year

fluctuations in the number of nests that affect our translation of the number of nests to the number of adult females and actual population pr ocesses. Current adult female population size,

n 0 , was estimated as the sum of the last 3 yr of data. We used the parametric bootstrap estimation procedure from Morris and Doak (2002) to compute the μ and σ 2 distributions required to calculate susceptibility to quasi-extinction (SQE) and a wide range of quasi-extinction thresholds (QETs). We used a range from 2.5 to 97.5% of the current abundance of nesting females as potential QETs. Because loggerhead turtles are likely to mature at > 30 yr (Snover 2002), we used the time period of 100 yr to compute QETs, which is consistent with the IUCN criteria (3 generations or 100 years, whicheve r is shorter). To incorporat e the uncertainty of parameter
estimates in determining SQE, we used 95% confidence limits of μ ˆ and 2 ˆ σ , using the
parametric bootstrapping met hod of Morris and Doak (2002).

The following steps were used to obtain SQEs:

1.
i μ and 2

i σ were drawn randomly and independent ly from respective distributions, {
i μ ~

,

2 ~

i σ }.

2. Each pair {
i μ ~

, 2 ~

i σ } was used to compute the probability of reaching a QET within the

next 100 years.

3. Steps 1 and 2 were repeated 5000 times ( i = 1, ..., 5000) to create a distribution of probabilities of reaching the QET.

4. The SQE metric for the QET is the proportion of the probabilities that are >0.9 (cut-off probability).38 5. Steps 1 through 4 were repeated for a range of QETs between 2.5% and 97.5% of the current abundance of nesting females.

Using simulations, Snover and He ppell (2009) demonstrated that SQE values greater than 0.4 indicated a population has >0.9 pr obability of quasi-extinction. At this critical value (SQE = 0.40), Type I and Type II errors are minimized simultaneously at approximately 10%. Reducing the critical value to 0.3 lessens the ‘Type I’ e rror rate but increases the ‘Type II’ error rate (Snover and Heppell 2009). The choice of 0.9 as th e cut-off probability was arbitrary and values other than 0.9 could be used. However, new critical values other th an 0.4 needed to be established for different values of the cut-off pr obability. Qualitatively, the results would not differ if a value other than 0.9 was used (Snover and Heppell 2009). In this assessment, we used the cut-off probability of 0.9 as in Snover and Heppell (2009) and a critica l value for the SQE of 0.30, which reduced the ‘Type I’ error (a DPS is consid ered to be not at risk when in fact it is). SQE values greater than 0.30, theref ore, indicate the DPS is at risk.

4.2.2. Count Data

The following datasets were used for the computa tions of SQEs. Ideally, the SQE analysis is conducted on the numbers of nesting females over time, which represent the temporal change in the number of reproductive females. The statisti c, however, is difficult to determine because of the required effort to mark and id entify individual turtles. Cons equently, data are often collected for the number of nests per nesting season. To estimate the number of females from the observed number of nests, we use the average nu mber of nests per female per nesting season for that region. Because the average is treated as a constant over time for each DPS (Table 1) and SQE is calculated as probability of a proportiona l reduction in population si ze, the SQE analysis is unaffected by the choice of unit (i.e., the number of females or nests). In the following, we use the number of females as the unit of analysis . Snover and Heppell ( 2009) suggest that time series at least 15 years in length are needed for diffusion approximation analysis, which limits the populations that have adequate data for applying this approach. We did include one additional nesting recovery unit in the Northwes t Atlantic Ocean DPS, the Northern Gulf of Mexico Recovery Unit, with only 12 years of data in order to characterize as much of that region as possible. Results from that recovery unit sh ould be interpreted with caution, although the general trends observed are consistent with region as a whole.39 Table 1. Life history parameters used for the ni ne DPSs. Juvenile survival rates (oceanic and neritic) were computed by using the negative binomial stage durat ion model and fixing survival rates of adults and the first year and fecundity. S ee text for details. Also please consult the excel file posted on the NMFS Office of Protected Resources website [ http://www.nmfs.noaa.gov/pr/species/statusreviews.htm ] to access information used in identifying the life history parameters.

North

Pacific South

Pacific Indian* NW

Atlantic NE

Atlantic Med South

Atlantic Fecundity Remigration interval (yrs) 2.7 3.8 3 3 3 2.6 3

Clutch freq. (yr -1 ) 3 3.4 5 5 5 2 5

Clutch size 112 127 101-118 115 85 95 123 Sex ratio (% female) 50 50 50 50 50 50 50 Emergence success 0.61 0.71 0.54 0.54 0.319 0.67 0.67

Habitat use

Juveniles Prop. neritic 0.14 0.50 0.655 0.655 0.655 0.50 0.655

Yrs oceanic 27.0 14.5 10 10 14.5 14.5 10

Yrs neritic 2.0 14.5 19 19 14.5 14.5 19 Mean AFR (yrs) 30 30 30 30 30 30 30

SD AFR 5 5 5 5 5 5 5

Adults Prop. neritic (non-nesting) 0.82 0.85 0.58 0.95 0.30 0.66 0.95

Prop. neritic (nesting) 0.82 0.85 0.73 0.95 0.30 0.66 0.95

Survival rates Aquatic hatchlings 0.400 0.400 0.400 0.400 0.400 0.400 0.400 Oceanic juvenile

λ 0 = 1.05

λ 0 = 1.10 0.858

0.914 0.828

0.894 0.796

0.877 0.794

0.875 0.858

0.933 0.856

0.933 0.787

0.863 Neritic juvenile

λ 0 = 1.05

λ 0 = 1.10 0.928

0.955 0.911

0.949 0.895

0.933 0.893

0.932 0.922

0.956 0.921

0.954 0.883

0.923

Oceanic adult 0.950 0.950 0.950 0.950 0.950 0.950 0.950

Neritic adult 0.950 0.950 0.950 0.950 0.950 0.950 0.950 *All parameters are identical for three DPSs in the Indian Ocean.40 4.2.2.1. North Pacific Ocean DPS

All loggerhead nesting in the North Pacific occurs in Japan. Nesti ng data for this region are from two sources, the Sea Turtle Association of Japan (STAJ; unpublished data provided to the

Western Pacific Fisheries Mana gement Council) and Kamezaki et al . (2002). Data from the STAJ represent total counts for Japan from 1998-2007. Snover (2008) combined these datasets to achieve a time series from Japan from 1990-2007 (Figure 1). Due to the nature of the STAJ data, the data from Japan are represented by a single time series, rather than by nesting beaches.

Figure 1. Change in the number of nesting fe males at nesting beaches for the North Pacific

Ocean DPS. The number of nes ting females was computed from the observed number of nests divided by the mean clutch frequency (3 yr -1 ; Table 1).41 4.2.2.2. South Pacific Ocean DPS

We used nesting census data for index beaches in eastern Australia (Limpus 2009; Figure 2). These include mainland beaches, Wreck Rock beaches, Great Barrier Reef (GBR) Coral Cays, and Wreck and Tyron Islands. Each beach was analyzed separately.

Figure 2. The change in the numbers of nesti ng females at four nesti ng beaches for the South

Pacific Ocean DPS.42 4.2.2.3. North Indian Ocean DPS

No adequate time series of nesting b each data was available for this DPS.

4.2.2.4. Southeast Indo-Pacific Ocean DPS

No adequate time series of nesting b each data was available for this DPS.

4.2.2.5. Southwest Indian Ocean DPS

Count data for the Southwest Indian Ocean DPS were obtained from Baldwin et al . (2003; Figure 3).

Figure 3. Change in the number of nesting females for the Southwest Indian Ocean DPS.43 4.2.2.6. Northwest Atlantic Ocean DPS

The Recovery Plan for the Northwest Atlantic Population of the Loggerhead Sea Turtle (NMFS and FWS 2008) recognized five recovery units (subpopulations) of loggerhead turtles w ithin the

Northwest Atlantic:

1. Northern Recovery Unit (southern VA through FL/GA border)

2. Peninsular Florida Recovery Unit (F L/GA border through Pinellas County, FL)

3. Dry Tortugas Recovery Unit (islands located west of Key West, FL)

4. Northern Gulf of Mexico Recovery Unit (Franklin County, FL, through TX)

5. Greater Caribbean Recovery Unit (Mexico th rough French Guiana, The Bahamas, Lesser Antilles, and Greater Antilles) Of these recovery units, four have adequate time series data for a pplying this analysis (Figure 4). Data for the Northern Recovery Unit and the Peni nsular Florida Recovery Unit comprise sums of

numerous individual beaches. For the Northern Gulf of Mexico Recovery Unit data, we used a

time series from the Florida Panhandle. We used nesting beach data from the Yucatan Peninsula (J. Zurita, personal communication, 2008) to repr esent the Greater Caribbean Recovery Unit.

4.2.2.7. Northeast Atlantic Ocean DPS

Although nesting data are available for the Cape Verde Archipelago, the time series is not long enough for the diffusion approximation analysis.

4.2.2.8. Mediterranean Sea DPS

Representative nesting beach data were not ava ilable for this DPS. Although nesting data are available for Greece and Turkey, the BRT doe s not think these data are sufficiently representative of the entire region given the pot ential differences in trends among nesting areas.

4.2.2.9. South Atlantic Ocean DPS

Data for the South Atlantic Ocean DPS were obtained from Marcovaldi and Chaloupka (2007; Figure 5).44 Figure 4. Changes in the numbers of nesting females at nesting beaches for the Northwest Atlantic Ocean DPS. The number of nesting females was computed from the observed number of nests divided by the mean clutch frequency (5 yr -1 ; Table 1). NRU = Northern Recovery Unit, PFRU = Peninsular Florida Recovery Unit, NGMRU = Northern Gulf of Mexico Recovery Unit, and GCRU = Greater Caribbean Recovery Unit.45 Figure 5. Change in the number of nesting fe males at nesting beaches for the South Atlantic

Ocean DPS. The number of nes ting females was computed from the observed number of nests divided by the mean clutch frequency (5 yr -1 ; Table 1).46 4.2.3. Results

The metric (susceptibility to quasi-extinction or SQE) is an increas ing function of quasiextinction threshold (QET). Unless a DPS is increasing, the likelihood of the population reaching some level of QET, as measured in th e proportion of current abundance, increases with QET. For example, if 95% of the current abundance is used as the QET, the likelihood of a declining population reaching the QET is high. For severely declini ng populations, the QET needs to be set very low to reach the defined SQE value of 0.3. Only the Southwest Indian

Ocean and South Atlantic Ocean D PSs indicated increasing trends ( μ > 0; Table 2). Large variability in the count data resulted in wide confidence intervals; all but three confidence intervals included zero (Table 2).

For the North Pacific Ocean DPS, SQE = 0.3 was reached at approximately 3% of the current female abundance, indicating the high likelihood of quasi-extinction for almost all levels of QET (Figure 6). This was caused by the recent decline of nesting females at the majority of nesting beaches in Japan (Snover 2008).

All monitored nesting beaches for the South Pacific Ocean DPS indicated high likelihood of SQE (Figure 7). As it was expected from observed counts (Figure 2), nesting beaches on

Mainland Australia indicated a bett er chance of persisting than othe r sites (Figure 7). For Wreck Rock, Great Barrier Reef (GBR), and Wreck a nd Tyron nesting beaches, quasi-extinction was certain for all values of QET.

For the recovery units with sufficient data w ithin the Northwest Atlantic Ocean DPS, the likelihood of quasi-extinction was highest for th e Northern Gulf of Mexico Recovery Unit (NGMRU), where SQE was greater than 0.3 for all values of QET (Figure 8) . For the other three recovery units, SQE = 0.3 was reached at QET < 0.3 (Figure 8).

Two DPSs indicated low likeli hoods of SQE; the Southwest I ndian Ocean (data from South Africa) and South Atlantic Ocean DPSs (Figures 9 and 10). Because of the observed increases in the nesting females in both time series (Baldwin et al . 2003, Figure 3; Marcovaldi and Chaloupka 2007, Figure 5), the lik elihood of quasi-extinctions ar e negligible for these DPSs using the SQE analysis.47 Table 2. Results of the diffusion approximation an alysis for each nesting beach or region. The

parameter μ is the arithmetic mean of the log population growth rate, and 2 σ is the variance of the
log population growth rate. CI is confidence in terval. PFRU = Peninsular Florida Recovery Unit, NRU = Northern Recovery Unit, NGMRU = No rthern Gulf of Mexico Recovery Unit, and GCRU = Greater Caribbean Recovery Unit.

DPS μ 95% CI μ σ 2 95% CI σ 2

North Pacific

All Japan

South Pacific -0.032 [-0.111, 0.046] 0.020 [0.011, 0.048]

Mainland Australia -0.013 [-0.047, 0.021] 0.009 [0.006, 0.016]

Wreck Rock,

Australia -0.038 [-0.107, 0.032] 0.023 [0.014, 0.048]

Great Barrier Reef,

Australia -0.050 [-0.116, 0.017] 0.026 [0.016, 0.049]

Wreck and Tyron Islands, Australia -0.075 [-0.150, -0.001] 0.030 [0.018, 0.058]

Southwest Indian 0.029 [0.003, 0.054] 0.005 [0.004, 0.009]

Northwest Atlantic

PFRU -0.026 [-0.065, 0.013] 0.005 [0.003, 0.013]

NRU -0.012 [-0.079, 0.055] 0.021 [0.012, 0.043]

NGMRU -0.049 [-0.121, 0.022] 0.009 [0.004, 0.029]

GCRU -0.012 [-0.068, 0.043] 0.010 [0.006, 0.025]

South Atlantic

Bahia and Espirito Santa, Brazil 0.046 [0.023, 0.068] 0.001 [0.001, 0.004]48 Figure 6. Susceptibility to quasi-extinction (S QE) as a function of qua si-extinction threshold (QET) for the North Pacific Ocean DPS. QET is defined as the proportion of the current female abundance. The dotted line indicates SQE = 0.3, which was adapted as the threshold for the analysis.49 Figure 7. Susceptibility to quasi-extinction (S QE) as a function of qua si-extinction threshold (QET) for the South Pacific Ocean DPS. QET is defined as the proportion of the current female abundance. Dotted lines indicate SQE = 0.3, whic h was adapted as the threshold for the analysis.50 Figure 8. Susceptibility to quasi-extinction (S QE) as a function of qua si-extinction threshold (QET) for the Northwest Atlantic Ocean DPS. QE T is defined as the proportion of the current female abundance. Dotted lines indicate SQE = 0. 3, which was adapted as the threshold for the analysis. NRU = Northern Recovery Unit, PFR U = Peninsular Florida Recovery Unit, NGMRU = Northern Gulf of Mexico Recovery Unit, and GCRU = Greater Caribbean Recovery Unit.51 Figure 9. Susceptibility to quasi-extinction (S QE) as a function of qua si-extinction threshold (QET) for the Southwest Indian O cean DPS. Data are for South Africa nesting beaches. QET is defined as the proportion of the current female abundance. Dotted lines indicate SQE = 0.3, which was adapted as the threshold for the analysis.52 Figure 10. Susceptibility to qua si-extinction (SQE) as a functi on of quasi-extinction threshold (QET) for the South Atlantic Ocean DPS. QET is defined as the proportion of the current female abundance. The dotted line indicates SQE = 0.3, which was adapted as the threshold for the analysis.53 

### 4.3. Threat Matrix Analysis

4.3.1. General Modeling Approach

The second approach to our risk analysis was based on a metric that indicates whether or not known threats may be sufficient to keep a DPS from recovering. Using as much information on the biology of loggerhead turtles as possible, a discrete-time st age-structured population model was constructed for each DPS. First, the model was parameterized to represent the maximum plausible vital rates and populati on growth rate, which may be attained for a recovering DPS (base population model hereafter). We did not incorporate demographic and environmental

stochasticities in the base matrix. Known an thropogenic threats to each life stage of a DPS, measured as additional annual mortality, were quantified using availabl e data and experts’ opinions, where the stage-specific additional annual mortality was summarized in a matrix format (threat matrix). The base population model and threat matrix then were multiplied to

determine effects of the additional mortalitie s on the base population model. We use the dominant eigenvalue ( λ ) of the product of the two matrices as an index of DPS status, where λ >1 indicates a healthy population, whereas λ <1 indicates possible decline of the population in the

future. Because the base matrix of a DPS wa s developed on hypothesized dominant eigenvalues,

the actual value of λ should not be considered as the populat ion growth rate of a DPS. We also provide the proportion of λ > 1. This proportion increases as the additional mortalities decrease.

Briefly, the risk analysis for each DPS consisted of four stages. First, a base population model was constructed. Second, anthropogenic threats we re quantified. Third, effects of the threats were calculated via the dominant eigenvalue of a linear time-invariant model. Finally, a range of dominant eigenvalues was determined according to the ranges of threat le vels. Details of each step are described in the following sections. Note that this analysis is similar to a demographic population viability analysis (PVA). However, we did not consider environmental or demographic stochasticity, dens ity dependence, autocorrelations in vital rates, or sampling variations. Consequently, this analysis should not be considered as a complete PVA. The two metrics (the dominant eige nvalue and the proportion of λ > 1) are indices of potential population decline of a DPS, considering the existing in formation about the biology of the species and experts’ opinions on known anthro pogenic threats to the DPS. The analysis does not provide estimates for the likelihood or probability of extinction.

4.3.2. Survival Rates and Population Grow th Rate for the Base Models

Using knowledge of life history of loggerhead turtles and following the previous work of others (Crouse et al . 1987, Crowder et al . 1994, Heppell et al . 2003), we constructed a stage-based population model for each DPS. The model cons isted of four stages; first year, oceanic juveniles, neritic juveniles, and adults. The dura tions of juvenile stages were modeled using the negative binomial stage di stribution model (NBSD mode l; Caswell 2001, pp. 164-165).

The NBSD model does not rely on the assumption of stable age distribution within a stage to compute the transition probability, as other methods would. Detailed desc riptions are available in Caswell (2001). Briefly, for a series of k identical pseudo-stages within a juvenile stage, we assign the transition probability ( p ) of moving from one pseudo-stag e to the next pseudo-stage.54 Because the total time requi red to step through all k pseudo-stages is equal to the time required

for the k th success in a series of identical Bernoulli trials with probability p , the total time ( T ) can be computed with a negative binomial distribution:



## ( ) kt k pp k t tT −

− ⎟ ⎟ ⎠ ⎞ ⎜ ⎜ ⎝ ⎛

− −

== 1 1 1 ) Pr(

, (1)

where kt ≥

, , ,2,1

L = k t , and 1 0 <

< p . The mean (E[T]) and variance (var[T]) of this

distribution are:

p k TE = ] [ and (2)

2 )1( ]var[

p pk T −

= . (3)

These equations can be solved for p and k by rearranging the above expressions:

][]var[ ][

TET TE p + = and (4)



## ( )

][]var[ ][ 2

TET TE k

+

= . (5)

Survival rates and transition probabilities among pseudo-stages are assumed identical. Therefore, each element of a ps eudo-stage is multiplied by the stage specific survival rate. In other words, the k diagonal elements for a juvenile stage are (1-p) φ , whereas the sub-diagonal

elements are p φ , where φ is the stage specific survival rate. For the same mean stage duration, a larger variance implies fewer pseudo stages ( k ) with smaller p , and vice versa. This method allowed us to construct projecti on matrices without assuming the stable age distribution within each juvenile stage or a fixed stage duration, while acknowledging the insufficient data to construct age-based models w ith variable growth rates.

The NBSD model requires the mean and variance of durations of stages. The age at first reproduction (AFR), however, has not been esti mated directly for loggerhead turtles. Skeletochronological studies have indicated that loggerhead turtle s in the western Atlantic may reach their first reproduction at approximately 30 years (Snover 2002). The experts of the team agreed on a standard deviation for the AFR of 5 years. Assuming a negative binomial distribution with a mean 30 years and a standard deviation of 5 y ears, 95% of AFRs are between 21 years and 41 years. These values were deemed reasonable by the experts. For computing the mean and variance for each stage of juveniles, i.e., oceanic and neritic, we used the same coefficient of variation (CV = 5/30 = 0.17). For ex ample, if juveniles of a DPS spend 15 years in oceanic and 14 years in neritic habitat, standard deviations for these stages are 2.6 and 2.4, respectively. We conducted a sens itivity analysis to evaluate th e effects of these parameters on

our conclusions.55 The stage-based model requires estimates for st age-specific fertility, mean and variance of durations of juvenile stages, and survival rates. Information from various sources provided data for fertility and hatching success of eggs (Table 1). The post- breeding census model was used for computing the fertility.

The DPS-specific matrices ( A ) have the following basic format:

⎥ ⎥ ⎥ ⎥ ⎥ ⎥ ⎥ ⎥ ⎥

⎦ ⎤

⎢ ⎢ ⎢ ⎢ ⎢ ⎢ ⎢ ⎢ ⎢

⎣ ⎡

− − −

A A

p

p p

p p

p ffp

φφ φ φ φ φ φ φ φφ

22

22 22

22 11

11 0 22

0 0)1( 0 0 0 0)1( 00 0 00 0)1( 0 000

LMMM O MMM M O MM MM O MMMLO ML L

,

where φ is survival rate, p is transition probability, f is the fertility (i.e., the number of female hatchlings that emerge per adult female), and subscripts indicate the first year (0), oceanic juveniles (1), neritic juve niles, (2), and adults ( A ). The numbers of sub-stages for juveniles vary according to the NBSD model described previously.

For survival rates, the vast majority of informa tion comes from studies at nesting beaches, where hatching success (egg survival) is evaluated routinely. However, the incubation period consists of only approximately 6.5 week s of the first year (Miller et al . 2003). Consequently, the total survival rate during the first year of their life is still unavailable. In the past, 0.4 was used for the Kemp’s ridley turtle ( Lepidochelys kempii ), based on a model fit to the observed numbers of nests and hatchling producti on for that species (Heppell et al . 2005). Because of the morphological and behavioral similarities betwee n the two species during the first year, we use the same survival rate for loggerh ead turtles. Adult survival rate s have been estimated for some loggerhead DPSs. Studies have in dicated that the survival rates of adult loggerhead turtles are generally greater than 0.8/yr (0.81; Fraze r 1983, 0.88; Chaloupka and Limpus 2002, 0.85; Hedges 2007). These estimates, however, like ly include anthropogenic mortalities. Consequently, the available estimates are negativ ely biased as the natural survival rate. Similarly for juveniles, available estimates for juvenile survival rate s include anthropogenic mortality rates (Bjorndal et al . 2003b, Braun-McNeill et al . 2007b, Sasso and Epperly 2007).

The relationship between possibl e ranges of juvenile survival rates and maximum population growth rates was evaluated graphically. The asym ptotic population growth rate of a stage-based matrix model is a function of fertility, survival rates, and transition probabilities. Consequently, the relationship among the average juve nile survival rates, first year survival rates, and dominant eigenvalues can be plotted. Si milar approaches have been used to evaluate the relationships among juvenile survival rates, adul t survival rates, fecundity, and average age at sexual maturity for the Blanding’s turtle ( Emydoidea blandingii ; Congdon et al . 1993). We considered adult survival rates from 0.80 to 0.99 per year, juvenile survival rates from 0.70 to 0.95 per year, and the first year survival from 0.01 to 0.50 per year . To constrain the parameter space, we made an56 assumption that the average juvenile survival rate was less than the average adult survival rate. Examinations of relationships among these parame ters allowed us to find the plausible maximum population growth rate ( λ ).

To compute the juvenile surviv al rates, a numerical minimization method was used. By fixing the dominant eigenvalue ( λ 0 ), the first year survival rate, adult survival rate, and fertility, the following equation was solved for two juvenile surviv al rates with the constr aint that the oceanic juvenile survival rate ( φ 1 ) is less than the neritic juvenile survival rate ( φ 2 ):

)(),( 021 A eig f −= λφφ

,

where eig () is the characteristic equation to fi nd the dominant eigenvalue of a matrix A and | | is the absolute value. The general form of the matrix A was shown previously.

In this analysis, several life history parameters were determined from experts’ knowledge or based on estimates from small sample sizes (S D of AFR, AFR, proportion of time spent in particular habitat). To determine how these para meters would affect the result of the analysis, sensitivity analyses were conducted. Results of these sensitivity anal yses are provided as supplemental material (Section 4.6).

4.3.3. Quantifying Known Threats

We quantified experts’ knowledge about the existing anth ropogenic threats on loggerhead turtles, which subsequently were combined with the proj ection model described in the previous section (see Tables 3-11 in Sections 4.4.2.1.-4.4.2.9. for the results of the experts’ threats assessments; also consult the excel file posted on the NMFS Office of Protected Resources website [ http://www.nmfs.noaa.gov/pr/species/statusreviews.htm ] to access information used to derive the estimated threats levels). Because the levels of the existing anthropogenic threats differ among habitats (e.g., high seas vs. coastal oceanic shelf), juveniles and adults were further separated into neritic and oceanic stages. Threat s, therefore, were dete rmined for the following six stages: (1) eggs/hatchlings, (2 ) neritic juveniles, (3) oceanic juveniles, (4) neritic adults, (5) oceanic adults, and (6) nesting females. For each of the following four factors, experts were asked to categorize known threats for each stage in high, medium, low, or very low, according to the level of additional annual mortality ( m ), where high = 0.20 < m ≤ 0.25, medium = 0.10 < m ≤ 0.20, low = 0.01 < m ≤ 0.10, and very low = 0 ≤ m ≤ 0.01. The four factors are:

• The present or threatened, destruction, modificat ion, or curtailment of its habitat or range (habitat).

• Overutilization for commercial, recreational, scientific, or educational purposes (overuse).

• Disease or predation.

• Other natural or manmade factors affec ting its continued existence (other).

The estimated threat levels were based on the best information available. Justifications and references for each threat are provided in Sect ion 4.4.2. and in the online threats matrix spreadsheets.57 These four factors corresponded to four of the five factors in section 4(a)(1) of the ESA. Another factor (inadequacy of ex isting regulatory mechanisms) was excluded from this analysis

because the effects of the factor were not cons idered to be reducing the survival rates of loggerhead turtles directly.

For two DPSs (South Pacific and North Pacifi c), experts were not able to distinguish anthropogenic mortalities from the natural mortalit ies. In other words, the provided mortalities included natural and anthropogenic mortalities. Consequently, an thropogenic mortalities were computed using the assumed base matrix model for each DPS, which was described in the previous section. Let m n = annual natural mortality, m a = annual anthropogenic mortality, and m = annual total mortality. In the discrete time scale, the total mortality is:

) 1)(1(1

n a mm m

− − − = . (6)

Consequently, the annual anthropoge nic mortality can be computed from the total and natural

mortalities:

1 1 1

− − +=

n a m m m (7)

To use these threat levels in calculations, we us ed the range limits of mortality values. For example, when the effect of a factor to a stage is categorized as ‘Low,’ 0.01 and 0.10 were used in the calculations as the lower and upper limits, respectively. To compute the total mortality,

mortalities from four factors were summed, which bounded the upper limit of “High” at 0.25 so

that the sum would not exceed one. The use of lower and upper limits provided the best and

worst case scenarios. The best case scenario was given by the sum of lower limits, whereas the upper limits were used to compute the worst case scenario.

For the three adult stages, threats for three habitats (neritic, oceani c, and nesting) were pooled to conform to the projection model. Because adu lt loggerhead turtles may use multiple habitats within a year, where they are exposed to diffe rent threats, we computed the total annual anthropogenic mortality rate from a threat ta ble using the following two equations. During a nesting year, the total anthropogenic mortality of nesting females is:

) 1()1()1(1

, ,

, lterrestria

adult T oceanic

adult T neritic

adult nesting adult m m m m nestingO nestingN

− − −−= . (8)

where m adult,nesting = annual anthropogenic mortality rate for adult females during the nesting year,

h

adult m = annual adult anthropogenic mortality in the habitat h from a threat table, where h is

either oceanic, neritic, or nesting, and T N,j , and T O,j are the average durations in years adult loggerhead turtles spend their time annually in neriti c and oceanic habitats in years, respectively,

and j is either nesting or non-nesting. The total anthropogenic mortality rate during a nonnesting year is:

nesting nonO nesting nonN T oceanic

adult T neritic

adult nesting nonadult m m m − −

− −−=

− , , )1()1(1

, , (9)58 Finally, with an estimated mean remigration interval ( R ), we calculated the average annual anthropogenic mortality rate of adults by:



# ( ) ( ) )1(

, , 1 11 +

− − −−= R nesting adult R

nesting nonadult adult m m m . (10)

Using the pre-defined mortality thresholds for each level of threats (very low, low, medium, and high), the total anthropogenic mortalities were co mputed for each of six stages. Although these thresholds may not represent the actual mort alities and may overestimate the anthropogenic mortalities, comparisons of mortalities among life stages within a DPS, as well as comparisons of mortalities among DPSs of a life stage, were conducted to pr ovide relative severity of anthropogenic mortalities among life stages. Finally , to evaluate the eff ects of these additional anthropogenic mortalities on a DPS, these addition al mortality rates were combined with the base matrix model and annual mortality rates for all stage classes computed.

4.3.4. Combining Additional Mortalitie s and the Base Population Model

Anthropogenic mortalities, quantified by the methods described in the previous section, were

combined with the ba se population model ( A ) from the first section via the “harvest” model described in Caswell (2001; Chapter 18). In the harvest model, the proportion of each stage group that survives the harvest is used to construct a “harvest matrix.” Because many of these anthropogenic threats to loggerhead turtle DPSs are not considered as harvest for human consumptions, we call them “threat survival matri ces.” A threat survival matrix and the base population model for each DPS are multiplied toge ther to construct a projection model with anthropogenic mortalities (Caswell 2001).

A threat survival matrix is a diagonal matrix: S = diag(s 1 , ..., s c ) , where s i is the proportion of

stage i surviving the threats ( s i = 1 – μ i ) and c is the number of stages. We assume a linear

model:

) ()1( t t SAn n = + , (11)

where t indicates time and n is a vector of stage-specific abundances (Caswell 2001). The dominant eigenvalue of SA ( λ ) is the asymptotic growth rate of the population with additional anthropogenic mortalities. Using the ranges of anthropogenic mortalities for each threat matrix, we computed the best ( λ B ) and worst ( λ W ) case scenarios of the dominant eigenvalues for each DPS, using the lower and upper limits of threat levels, respectively.

Finally, we provide the proportion of λ > 1.0 in percent (P λ ), which is computed by P λ = 100×( λ B – 1.0)/( λ B - λ W ), when λ B > 1.0, and P λ = 0.0, when λ B ≤ 1.0. This metric (P λ ) is a function of several components; the assumed maximum popu lation growth rate, assumed life history parameters, thresholds of threat levels, and st ate of knowledge about pr ecision of the existing threats. Assumed life history parameters and the maximum popul ation growth rate define how

much additional mortalities can be tolerated be fore the dominant eigenvalue becomes < 1.0.

Assumed natural survival rates of various life stages can affect the computed proportion; stages with high natural surviv al rates and low reproductive values can tolerate greater additional59 mortalities than those with low natural survival rates and high reproductive values. Changes in threshold values of threat matrices also can change the results. If the thresholds for each threat level were decreased, e.g., 0.01 to 0.05 for low, rather than from 0.01 to 0.10 and so forth for other levels, the difference between λ B and λ W would decrease and the pr oportion would increase. Decreasing the uncertainty in additional mortalities would also decrease the difference (and increase the proportion). Consequently, P λ can be considered as an index of the future population health, where it ranges from 0 to 100, conditional on the specific assumptions. The management goal for each DPS is to maximize P λ , which may be attained by reducing additional mortalities and increasing the pr ecisions of estimated life history parameters and additional

mortalities.

To determine how a “very low” level of threats may affect the computations of the population growth rate, an analysis was conduc ted with all threat categories set to “very low.” This analysis provided the baseline with which the other results can be compar ed. Because experts’ opinions on mortalities were provided eith er with natural mortality (South Pacific and North Pacific) or without natural mortality, the effect of this di fference also was determined with analysis using this hypothetical threat matrix.



### 4.4. Results

4.4.1. Results—Base Population Model

For plausible ranges of adult surviv al rates (0.80-0.99/yr), first year survival rates (0.01-0.50/yr), and juvenile survival rates (0.70-0.95/yr), the maximum population growth rates appeared to be less than 10%/yr for a ll DPSs. We found many combinations of the parameters to be unlikely because of the necessarily high average j uvenile survival rates (Figures 11-17).

Except for the Mediterranean Sea DPS (Figure 16), differences among DPSs were negligible. This exception was due to the differences in fe rtility parameters, where the Mediterranean Sea DPS had low average eggs per clutch (95 eggs per clutch) and low number of clutches per female (2, Table 1).

Parameters that defined habitat use were sim ilar among all DPSs, except the North Pacific Ocean DPS (Table 1). For the North Pacific Ocean DPS, the proportion of time juveniles spend in the neritic habitat (off the coast of Baja Peninsula, Mexico, and in coastal areas of the western Pacific) was less (14%) than for other DPSs ( 50-65.5%). The difference for this DPS comes from the experts’ opinions that a large proportion of the juvenile loggerhead turtles of this DPS remain in the pelagic habitat. Further, thos e juveniles along the coast of Baja Peninsula are thought to remain in the area for many years.

As expected, the sensitivity analysis indicated the change in the mean AFR can affect the relationship among the survival rates and asym ptotic population growth rates (Supplemental Figures S1-S7). For the same value of CV, ol der age at first reproduction resulted in higher average survival rates of juvenile s and adults to sustain an asym ptotic population growth rate. Effects of CVs were less than those of the mean AFR. For the following analyses, we use the

mean AFR = 30 and CV = 0.17 (or SD of AFR = 5 years).60 Figure 11. Relationships among adult survival rate, j uvenile survival rate, fi rst year survival rate, and the dominant eigenvalue for the North Paci fic Ocean DPS. Contour lines indicate the dominant eigenvalues, whereas th e diagonal lines indicate equal annual survival rates between juveniles and adults. Dashed horizontal lines in dicate the assumed adult survival rate for the base matrix model. The mean age at first reproduction was assumed at 30 and its standard

deviation 5.61 Figure 12. Relationships among adult survival rate, j uvenile survival rate, fi rst year survival rate, and the dominant eigenvalue for the South Paci fic Ocean DPS. Contour lines indicate the dominant eigenvalues, whereas th e diagonal lines indicate equal annual survival rates between juveniles and adults. Dashed horizontal lines in dicate the assumed adult survival rate for the base matrix model. The mean age at first reproduction was assumed at 30 and its standard

deviation 5.62 Figure 13. Relationships among adult survival rate, juvenile survival rate, first year survival rate, and the dominant eigenvalue for the North Indian Ocean, Southeast Indo-Pacific Ocean, and Southwest Indian O cean DPSs. Contour lines indicate the dominant eigenvalues, whereas the diagonal lines indicat e equal annual survival rates between juve niles and adults.

Dashed horizontal lines indicate the assumed adult su rvival rate for the base matrix model. The mean age at first reproduction was assumed at 30 and its standard deviation 5.63 Figure 14. Relationships among adult survival rate, j uvenile survival rate, fi rst year survival rate, and the dominant eigenvalue for the Northwest Atlantic Ocean D PS. Contour lines indicate the dominant eigenvalues, whereas th e diagonal lines indicate equal annual survival rates between juveniles and adults. Dashed horizontal lines in dicate the assumed adult survival rate for the base matrix model. The mean age at first reproduction was assumed at 30 and its standard

deviation 5.64 Figure 15. Relationships among adult survival ra te, juvenile survival rate, total first year survival rate, and the dominant eigenvalue for th e Northeast Atlantic Ocea n DPS. Contour lines indicate the dominant eigenvalu es, whereas the diagonal lines indicate equal annual survival rates between juveniles and adults. Dashed horiz ontal lines indicate the assumed adult survival rate for the base matrix model. The mean ag e at first reproduction was assumed at 30 and its

standard deviation 5.65 Figure 16. Relationships among adult survival rate, j uvenile survival rate, fi rst year survival rate, and the dominant eigenvalue for the Mediterr anean Sea DPS. Contour lines indicate the dominant eigenvalues, whereas th e diagonal lines indicate equal annual survival rates between juveniles and adults. Dashed horizontal lines in dicate the assumed adult survival rate for the base matrix model. The mean age at first reproduction was assumed at 30 and its standard

deviation 5.66 Figure 17. Relationships among adult survival rate, j uvenile survival rate, fi rst year survival rate, and the dominant eigenvalue for the South Atlant ic Ocean DPS. Contour lines indicate the dominant eigenvalues, whereas th e diagonal lines indicate equal annual survival rates between juveniles and adults. Dashed horizontal lines in dicate the assumed adult survival rate for the base matrix model. The mean age at first reproduction was assumed at 30 and its standard

deviation 5.67 4.4.2. Results—Threat Matrix

This analysis indicated potential overestimati ons by the experts of an thropogenic mortalities (Tables 3-11). The threat matrices indicated th at many DPSs might be in critical condition with high mortalities for all stages. One reviewer suggested using a different set of mortality thresholds for building the threat matrices. The suggested change, however, was arbitrary and expresses a false sense of certain ty, although our matrices may provide a false sense of urgency.

Because our threat levels were determined from available information on anthropogenic mortalities and abundance of logge rhead turtles, we retained the thresholds. The potential implications of the differences in opinions w ill be discussed later. In the following nine subsections, we provide threat levels for si x life stages of each DPS based on available information and experts’ opinions. We also provid e evidence for each threat. Further, using the threshold values presented previ ously, we provide numerical values as sums of all threats for each life stage under the four categories.68 4.4.2.1. North Pacific Ocean DPS

For the North Pacific Ocean DPS, the greatest anthropogenic threats exist for the neritic juveniles and neritic adults, where the coastal fish eries in Baja California, Mexico, and Japan kill significant number of turtles (Table 3) (Kamezaki et al . 2003, Peckham et al . 2007). A large number of loggerhead turtles (~ 1,000 annually) die in Baja botto m-set gillnet/longline fisheries (Koch et al . 2006, Peckham et al . 2007, Peckham et al . 2008). Fishery bycatch also is present in West Pacific neritic foraging areas (Cheng and Ch en 1997). Recent investigations show a major threat off coastal Japan due to pound net fisheries (T. Ishihara, Sea Turtle Association of Japan, personal communication, 2007). Coas tal development, such as the building of seaw alls, erosion control structures, beachfront lighting, vehicular and pedestrian traffic, sand extraction, beach pollution, removal of native vegetation, and pl anting of non-native vegeta tion, has resulted in reduction of available habitat for the eggs and hatchlings of this DPS (Suganuma 2002,

Kamezaki et al . 2003, Kudo et al . 2003). Beach debris also contri butes to deaths of hatchlings and nesting adults (H. Peckham, ProPeninsula, pers onal communication, 2009). Although no quantitative studies have been conducted, low hatching success has been documented at important nesting beaches in recent years. Thr eat levels for this DPS include anthropogenic and

natural mortalities.

Table 3. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the North Pacific Ocean DPS. Threat le vels were estimated by life stages and habitats, and were grouped into four categories based on four of the five factors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be re ducing the survival rates directly. The column with “Best” indicates the total mortality using the lowest threshold values for all factors, whereas the column with “Worst ” indicates the total mort ality using the highest

threshold value for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings M VL L M 0.21 0.51 Neritic juveniles L L L M 0.13 0.50 Oceanic juveniles VL L VL L 0.02 0.22

Neritic adults L L VL M 0.12 0.41

Oceanic adults VL VL L L 0.02 0.22 Nesting females VL VL VL VL 0.00 0.0469 4.4.2.2. South Pacific Ocean DPS

For the South Pacific Ocean DPS, the greatest threats come from bycatch in the oceanic environment for juveniles and adults (Table 4). Threats from numerous fisheries have been reported (Limpus and Reimer 1994; Poiner and Harris 1996; Robins et al . 2002a, b; Kelez et al

. 2003; Alfaro-Shigueto et al . 2006; Donoso and Dutton 2006; Alfaro-Shigueto et al . 2008b; Limpus 2009). At nesting beaches, an averag e of 40 adult females are taken annually in aboriginal harvest (Limpus 2009). Threat levels for this DPS include anthropogenic and natural

mortalities.

Table 4. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the South Pacific Ocean DPS. Threat le vels were estimated by life stages and habitats, and were grouped into four categories based on four of the five factors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be re ducing the survival rates directly. The column with “Best” indicates the total mortality using the lowest threshold values for all factors, whereas the column with “Worst ” indicates the total mort ality using the highest

threshold value for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings L L L L 0.04 0.40 Neritic juveniles L L L L 0.04 0.40 Oceanic juveniles VL VL VL M 0.10 0.23

Neritic adults L L L L 0.04 0.40

Oceanic adults VL VL L M 0.11 0.32 Nesting females VL M VL VL 0.10 0.2370 4.4.2.3. North Indian Ocean DPS

Threat levels for the North Indi an Ocean DPS were “very low” to “low” for all factors except for the fourth factor (“other”) in th e neritic environment (Table 5). The relatively high mortalities in the neritic environment result from fishery bycatch . Consistent low threats across all factors for eggs and hatchlings may affect a significant porti on of this stage. These threats include coastal development and light pollution, beach habitat degradation from overgrazing and beach driving, and bycatch from coastal fisheries.

Table 5. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the North Indian Ocean DPS. Threat le vels were estimated by life stages and habitats, and were grouped into four categories based on four of the five factors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be re ducing the survival rates directly. The column with “Best” indicates the total mortality using the lowest threshold values for all factors, whereas the column with “Worst ” indicates the total mort ality using the highest

threshold value for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings L L L L 0.04 0.40 Neritic juveniles L VL/L L M 0.12 0.41 Oceanic juveniles VL VL VL L 0.01 0.13

Neritic adults L VL/L L M 0.12 0.41

Oceanic adults VL VL VL L 0.01 0.13 Nesting females L VL VL L 0.02 0.2271 4.4.2.4. Southeast Indo-Pacific Ocean DPS

Anthropogenic effects are very low to low at all li fe stages for the Southeast Indo-Pacific Ocean DPS (Table 6). Cumulatively, however, a larg e proportion of eggs and hatchlings (~30%) may be affected by human activities.

Table 6. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the Southeast Indo-Paci fic Ocean DPS. Threat levels were estimated by life stages and habitats, and were grouped into four categories base d on four of the five f actors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be reducing the survival rates directly. The column with “Best” indicate s the total mortality using the lowest threshold values for all factors, whereas the column with “Worst” indicates the total mortality using the highest threshold valu e for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings VL/L VL L L 0.02 0.31 Neritic juveniles L VL VL L 0.02 0.22 Oceanic juveniles VL VL VL L 0.01 0.13

Neritic adults L VL VL L 0.02 0.22

Oceanic adults VL VL VL L 0.01 0.13 Nesting females VL VL VL VL 0.00 0.0472 4.4.2.5. Southwest Indian Ocean DPS

For the Southwest Indian Ocean DPS, wide ranges of total mortalities were found for eggs/hatchlings, neritic juveniles, and neritic adults (Table 7). These wide ranges result from

conservative estimation based on limited in formation on anthropogenic mortalities.

Table 7. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the Southwest Indian Ocean DPS. Threat levels were estimated by life stages and habitats, and were grouped into four categories base d on four of the five f actors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be reducing the survival rates directly. The column with “Best” indicate s the total mortality using the lowest threshold values for all factors, whereas the column with “Worst” indicates the total mortality using the highest threshold valu e for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings VL/L L L L 0.03 0.40 Neritic juveniles L VL/L VL L/M 0.02 0.40 Oceanic juveniles VL VL VL L 0.01 0.13

Neritic adults L VL/L VL L/M 0.02 0.41

Oceanic adults VL VL VL L 0.01 0.13 Nesting females VL VL VL VL 0.00 0.0473 4.4.2.6. Northwest Atlantic Ocean DPS

High cumulative threat levels were estimated fo r neritic and oceanic juveniles and neritic and oceanic adults (Table 8). These mortalities pr imarily result from fishery bycatch. Significant mortality occurs in longline fish eries, bottom and mid-water trawl fisheries, dredge fisheries, gillnet fisheries, and pot/trap fisheries. Although total mortality from all fisheries has not been estimated, the combined mortalities are likely sign ificant. Entanglement in marine debris is an additional threat. Further, boat strikes are another growing anthropogenic source of mortality in neritic waters (NMFS, unpublished data; Florid a Fish and Wildlife Conservation Commission, unpublished data).

Table 8. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the Northwest Atlantic Ocean DPS. Threat levels were estimated by life stages and habitats, and were grouped into four categories base d on four of the five f actors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be reducing the survival rates directly. The column with “Best” indicate s the total mortality using the lowest threshold values for all factors, whereas the column with “Worst” indicates the total mortality using the highest threshold valu e for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings L VL L L 0.03 0.31 Neritic juveniles L L L M 0.13 0.50 Oceanic juveniles VL VL VL M/H 0.10 0.28

Neritic adults L L L M 0.13 0.50

Oceanic adults VL VL VL M/H 0.10 0.28 Nesting females VL VL VL VL 0.00 0.0474 4.4.2.7. Northeast Atlantic Ocean DPS

For the Northeast Atlantic Ocean DPS, relatively high threats are re ported for terrestrial, neritic, and oceanic habitats (Table 9). The largest s ource of anthropogenic mortality is deliberate hunting of nesting females. Other threats include harvest of eggs and hatchlings, incidental capture, particularly in fisherie s, propeller and boat strikes, exposure to hydrocarbons, marine debris ingestion, and bioaccumulation of metals (see Section 5.2.7.). Sand mining has impacted the terrestrial habitats in some areas and change s in trophic structure likely have occurred in both the neritic and oceanic environments of the regi on. Shifts in ocean currents, and ecosystem shifts in prey distribution and abundan ce, could occur with climate change.

Table 9. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the Northeast Atlantic Ocean DPS. Threat levels were estimated by life stages and habitats, and were grouped into four categories base d on four of the five f actors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be reducing the survival rates directly. The column with “Best” indicate s the total mortality using the lowest threshold values for all factors, whereas the column with “Worst” indicates the total mortality using the highest threshold valu e for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings L L VL L 0.03 0.31 Neritic juveniles L L VL M 0.12 0.41 Oceanic juveniles VL VL VL M/H 0.10 0.28

Neritic adults L L VL M 0.12 0.41

Oceanic adults VL VL VL M/H 0.10 0.28 Nesting females VL H VL L 0.21 0.3775 4.4.2.8. Mediterranean Sea DPS

Fishery bycatch was determined to be the greatest threat for the Mediterranean Sea DPS (Table 10). For neritic and oceanic juveniles and adults, these threats were considered to be “medium” or “high.” Bycatch of juveniles and adults in trawl, pelagic and bottom longline, set net (gillnet and trammel nets), and driftnet fisheries is considered significan t (Laurent 1991, Argano et al

. 1992, Di Natale 1995, Lazar and Tvrtkovic 1995, Laurent et al . 1996, Camiñas 1997, Gerosa and Casale 1999, Silvani et al . 1999, Lazar et al . 2000, Laurent et al . 2001, Camiñas 2004, Carreras

et al . 2004, Casale et al . 2004, Tudela et al . 2005, Casale et al . 2007a, Jribi et al . 2007, Casale 2008). See Section 5.2.8.5. for capture leve ls/citations. Other notable concerns are pollution/debris and vessel collisions. The other significan t threat comes from habitat destruction that affects eggs and hatchlings at nesting beaches.

Table 10. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the Mediterranean Sea DPS. Threat le vels were estimated by life stages and habitats, and were grouped into four categories based on four of the five factors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be re ducing the survival rates directly. The column with “Best” indicates the total mortality using the lowest threshold values for all factors, whereas the column with “Worst ” indicates the total mort ality using the highest

threshold value for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings M L L L 0.13 0.50 Neritic juveniles L L VL M/H 0.12 0.46 Oceanic juveniles VL L VL M/H 0.11 0.37

Neritic adults L L L M/H 0.13 0.55

Oceanic adults VL L VL M/H 0.11 0.37 Nesting females VL L VL VL 0.01 0.1376 4.4.2.9. South Atlantic Ocean DPS

Significant anthropogenic threats exist for juveni le stages of the South Atlantic Ocean DPS (Table 11). For these stages, the most significant factor is fishing bycatch, where coastal and oceanic fishery operations interact with loggerheads in the area (Bugoni et al . 2008). Further, bycatch in longline fisheries appe ar to have significant effects on juvenile loggerhead turtles in the high seas (Pinedo and Polacheck 2004, Kotas et al . 2004).

Table 11. Estimated levels and quantificati on of anthropogenic threat s affecting loggerhead

turtles in the South Atlantic Ocean DPS. Threat levels were estimated by life stages and habitats, and were grouped into four categories based on four of the five factors in section 4(a)(1) of the ESA. Another factor (inadequacy of existing regulatory mechanisms) was excluded from this analysis because the effects of the factor were not considered to be re ducing the survival rates directly. The column with “Best” indicates the total mortality using the lowest threshold values for all factors, whereas the column with “Worst ” indicates the total mort ality using the highest

threshold value for all factors.

Habitat Overuse Disease/

Predation Other Best Worst

Eggs/hatchlings L L VL L 0.03 0.31 Neritic juveniles L VL L M 0.12 0.41 Oceanic juveniles VL VL L H 0.21 0.37

Neritic adults L VL L L 0.03 0.31

Oceanic adults VL VL L L 0.02 0.22 Nesting females VL VL VL VL 0.00 0.0477 Although the experts were asked to provide qualitative threat levels (very low, low, medium, or high) based on threshold values for each level (see Section 4.3.3 for details), it was sometimes a difficult task to assign an anthr opogenic mortality level to a partic ular stage class when few or no quantitative data existed. Because precise esti mates of anthropogenic mortalities do not exist, we used the precautionary principle for character izing the threat level based on best available information and/or expert opinion.

Even though threats were expressed in a quali tative manner, every entry was derived from available data and experts’ knowledge. The co mpilation of known threats for all DPSs provided an opportunity to evaluate the potential negativ e impacts of these thr eats on each DPS in the future. Regardless of the threshold values for ea ch threat level (per suggestion of a reviewer), comparisons among life stages within a DPS, as well as comparisons of a life stage among DPSs, are meaningful. For example, the North Pacifi c Ocean and Mediterranean Sea DPSs were shown to have high anthropogenic mortalities for the eggs and hatchling stage, relative to other DPSs. Within the North Pacific Ocean DPS, for another example, neritic habitats (along the west coast of Baja California, Mexico, and the nearshore wa ters of Japan) indica ted greater anthropogenic mortalities than other life stages. Finally, the indices of DPS health ( λ and P λ ) also provided metrics that combined the status and uncertainty of threats levels for each DPS. Interpretations

of these indices in terms of the DPS stat us are explained in the following section.

4.4.3. DPS Status

The analysis with a hypothetical threat matrix with all threat categories set to “very low” indicated the obvious effects of combining anthropogenic and natu ral mortalities in a threat matrix. For those threat matrices of DPSs that included natural mortalities (North Pacific and South Pacific), computed anthropogenic mortal ities were necessarily less than the assumed thresholds. Consequently, ranges of possible λ values for those DPSs were less than those DPSs that included only anthropogenic mort alities in their threat matrices (Figures 18 and 19). Even when all threat levels were set at “very low,” P λ was not 100% for the DPSs that included only anthropogenic mortalities in th e threat matrices and the a ssumed maximum population growth rate was 5% annum (Figure 18). “Very low,” ther efore, may not have been an appropriate term for this category. As it was stated in the met hod section, management decisions should focus on maximizing P λ

.78 Figure 18. Ranges of dominant eigenvalues for nine DPSs when anthropogenic mortalities were considered to be “very low” for all factors and the maximum natural population growth rates were assumed to be 5% per year ( λ 0 = 1.05). The mean age at first reproduction (AFR) was assumed to be 30 years and its standard devia tion (SD) 5. The values above the dotted line indicates the proportion of the bars that are above λ = 1.0 (P λ )79 Figure 19. Ranges of dominant eigenvalues for nine DPSs when anthropogenic mortalities were considered to be “very low” for all factors and the maximum natural population growth rates were assumed to be 10% per year ( λ 0 = 1.10). The mean age at first reproduction (AFR) was assumed to be 30 years and its standard devia tion (SD) 5. The values above the dotted line indicates the proportion of the bars that are above λ = 1.0 (P λ ).80 For computing P λ , the BRT did not consider functional forms of threat levels. In other words, all values between the best and worst case scenarios were considered equally likely. This assumption was fundamental fo r computing the proportion of λ > 1.0. With more information on threat levels, it may be possible to assign particular distributions to these threat levels, thereby increasing the precision and accuracy of P λ

.

Among the nine DPSs in this analysis, an index of future population health (P λ = the proportion

of λ > 1.0) was 0 for North Indian Ocean, Northwes t Atlantic Ocean, Northeast Atlantic Ocean, Mediterranean Sea, and South Atlantic Ocean D PSs, when the maximum population growth rate was assumed to be 5% per year (Figure 20). Even if the maximum population growth rate was 10% per year, P λ for the Northwest Atlantic Ocean, No rtheast Atlantic Ocean, Mediterranean Sea, and South Atlantic Ocean DPSs was zer o (Figure 21). When the maximum population growth rate was assumed to be 5% per year (Figure 20), P λ ranged from 0.0% to 16.9% (North Pacific Ocean DPS). The large value for the North Pacific Ocean DPS is caused by precise estimates of threats for the DPS, where the length of the bar is less than for the other DPSs. For

those DPSs that had P λ = 0 when the maximum population growth rate was assumed to be 5% per year, only the North Indian Ocean DPS resulted in P λ > 0 when the maximum population growth rate was assumed to be 10% per year. P λ ranged from 0.0% to 40.4% (Southeast IndoPacific Ocean DPS), when the maximum populati on growth rate was assumed to be 10% per year (Figure 21). According to P λ , all DPSs are faced with a potenti al decline in the future due to the existing additional mortalities, relative to th e assumed natural survival rates and fertility.81 Figure 20. Ranges of dominant ei genvalues for nine DPSs with anthropogenic mortalities that were provided by the experts. For this figure, the maximum population growth rates for all DPSs were assumed to be 5% per year ( λ 0 = 1.05). The mean age at first reproduction (AFR) was assumed to be 30 years and its standard devia tion (SD) 5. The values above the dotted line indicates the proportion of the bars that are above λ = 1.0 (P λ ).82 Figure 21. Ranges of dominant ei genvalues for nine DPSs with anthropogenic mortalities that were provided by the experts. For this figure, the maximum population growth rates for all DPSs were assumed to be 10% per year ( λ 0 = 1.10). The mean age at first reproduction (AFR) was assumed to be 30 years and its standard devia tion (SD) 5. The values above the dotted line indicates the proportion of the bars that are above λ = 1.0 (P λ ).83 

### 4.5. Synthesis

For three of five DPSs with sufficient data to conduct the SQE analysis (Northwest Atlantic Ocean, South Pacific Ocean, and North Pacific O cean), the threshold of SQE = 0.3 was reached at QET < 0.3, indicating high likelihood of quasi-e xtinction over a wide range of QET values. Similarly, the threat matrix analysis indicated th at all other DPSs have the potential for a severe decline in the future (Figur es 20 and 21). There were not enough data to conduct the SQE analysis for the Northeast Atlantic Ocean, Mediterranean Sea, North Indian Ocean, and

Southeast Indo-Pacific Ocean DPSs.

It should be noted that the SQE analysis is base d on nesting beach trend data. Trends at nesting

beaches are often used to assess the status of marine turtle populations , because nesting beach studies are usually more easily implemented than in-water studies and, in many cases, have been ongoing for many years. However, because of the pr otracted age to maturity in marine turtles, mortality that is acting on younger lif e stages will not yet be reflect ed in nest or nesting female trends, and may take many years to be detected at the nesting beach. Conversely, mortality acting on older life stages, including adults, will be reflected more rapidl y in nest or nesting female trends at nesting beaches. The SQE appr oach does not incorporat e different mortalities among life stages. The threat matrix approach was used to determine effects of known anthropogenic mortalities on each DPS.

According to the threat matrix analysis, using ex perts’ opinions in the matrix model framework, all loggerhead turtle DPSs have the potential to decline in the future. Although some DPSs are indicating increasing trends at nesting beaches (Southwest Indian Ocean and South Atlantic Ocean), available information about anthropogenic th reats to juveniles and adults in neritic and oceanic environments indicate possible unsustainab le additional mortalitie s (Figures 20 and 21). According to the threat matrix analysis, the pote ntial for future decline is greatest for the North Indian Ocean, Northwest Atlantic Ocean, Nort heast Atlantic Ocean, Me diterranean Sea, and

South Atlantic Ocean DPSs. We emphasize, however, that these indices were used to measure the negative effects of known anth ropogenic mortalities on the overall health of each DPS and not to estimate the actual population growth rates of these DPSs.

Our approach to the risk analysis presented several important points. First, the lack of precise estimates of age at first repr oduction hindered precise assessment of the status of any DPS. Within the range of possible ages at first repr oduction of the species, however, some DPSs could decline rapidly regardless of the exact age at first reproduction because of high anthropogenic mortality.

Second, the lack of precise estimates of anthr opogenic mortalities resulted in a wide range of possible status. For the best case scenario, a DPS may be healthy, whereas the same DPS may be considered as declining rapidly for the wors t case scenario (Figures 20 and 21). The precise prognosis of each DPS relies on obtaining precise es timate of anthropogenic mortality and vital

rates.

Third, the assessment of a population without the information on natu ral and anthropogenic mortalities is difficult. Because of the longevi ty of the species, loggerh ead turtles require high84 survival rates throughout their life to maintain a population. Anthropogenic mortality on the species occurs at every stage of their life, wher e the exact magnitude of the mortality is often unknown. As described in this document, the u pper end of natural mortality can be computed from available information. In the future, as more comprehensive and precise information on anthropogenic threats is gathered and availabl e, improvements in model outcomes will result.

Nesting beach count data for the North Paci fic Ocean DPS have indicated a decline of loggerhead turtle nesting in the last 20 year s (Figure 1). The SQE approach reflected the observed decline (Figure 6). However, in the threat matrix analysis, the range of λ values contains 1.0, indicating the large uncertainty about the future of th e DPS. As stated elsewhere, fishery bycatch along the coast of the Baja Penins ula and the nearshore wa ters of Japan are the main known sources of mortalities (Table 3). Mo rtalities in the high-se as, where a large number of juvenile loggerhead tu rtles reside (Kobayashi et al . 2008), from fishery bycatch are still

unknown.

The SQE approach indicated that, based on nest count data for the past three decades, the South Pacific Ocean DPS is at risk and thus likely to decline in the foreseeable future. These results were based on recently published nesting census data for loggerhead turtles at index beaches in eastern Australia (Limpus 2009). The threat matrix approach provided uncerta in results: in the case of the lowest anthropogenic threats, the So uth Pacific Ocean DPS may recover, but in the worst-case scenario, the DPS may substantially d ecline in the foreseeable future. These results are largely driven by the ongoing th reats to juvenile and adult l oggerheads from fishery bycatch that occur throughout the South Pacific Ocean and th e uncertainty in estimated mortalities (Table 4).

For the North Indian Ocean DPS, there were no ne sting beach data availa ble to conduct the SQE analysis. The threat matrix analysis indicated a decline of the DPS in the foreseeable future, primarily as a result of fishery bycatch in neriti c habitats. Cumulatively, substantial threats may exist for eggs/hatchlings (Table 5). Because of the lack of precise estimates of bycatch, however, the range of possible λ values was large (Figures 20 and 21).

Similar to the North Indian Ocean DPS, no nesti ng beach data were available for the Southeast Indo-Pacific Ocean DPS. The level of anthropoge nic mortalities is low for the Southeast IndoPacific Ocean DPS, based on the best available information, resulting in relatively large P λvalues and a narrow range (Figures 20 and 21). The greatest threats for the Southeast IndoPacific Ocean DPS exist for the fi rst year of the life stages (egg s and hatchlings; Table 6). For the Southwest Indian Ocean DPS, the SQE approach , based on a 37-year time series of nesting female counts at Tongaland, South Africa (1963- 1999), indicated this segment of the population, while small, has increased, and the likelihood of quasi -extinction is negligible . The threat matrix approach, on the other hand, provided a wide range of results: in the best case scenario, the DPS would grow slowly, whereas in the worst ca se scenario, the DPS would decline in the

foreseeable future. The results of the threat matrix approach were driven by uncertainty in anthropogenic mortalities in the neritic envir onment and the eggs/hatch ling stage (Table 7, Figures 20 and 21).85 The SQE analysis indicated diffe rences in SQEs among recovery units within the Northwest Atlantic Ocean DPS (Figure 8). Although the Northe rn Gulf of Mexico Recovery Unit indicated the worst result among the four recovery units unde r consideration (Figure 8), the length of the time series was shortest (12 data points; Figure 4). Other recovery units , however, appeared to show similar declining trends, which were also indicated through the SQE approach. The threat matrix analysis indicated a likely decline of the DPS in the foreseeable future. The greatest threats to the DPS result from cumulative fishery bycatch in neritic and oceanic habitats (Table 8).

The high likelihood of the predicted decline of the North east Atlantic Ocean DPS from the threat matrix analysis is largely dr iven by the ongoing harvest of nes ting females, low hatchling and emergence success, and mortality of juveniles an d adults from fishery bycatch throughout the Northeast Atlantic Ocean (Table 9). The thre at matrix analysis indicated a consistently pessimistic future for the DPS (Figures 20 and 21).

Nesting beach data for the Mediterranean S ea DPS were not available to conduct the SQE analysis. The threat matrix analysis indicated the DPS is likely to decl ine in the foreseeable future (Figures 20 and 21). The primary thr eats are fishery bycatch in neritic and oceanic habitats (Table 10).

The two approaches for determining risks to th e South Atlantic Ocean DPS provided different, although not incompatible, results. The SQE approach indicated that, based on nest count data for the past two decades, the population was unlikely to decline in the foreseeable future (Figure 10). These results were based on recently publis hed nesting beach trend analyses by Marcovaldi and Chaloupka (2007) and this QET analysis was consistent with their conclusions. However, the SQE approach was based on past performanc e of the DPS, specifically only nesting beach data, and did not address ongoing or future threats to segments of the DPS that might not have been nor might not yet be reflected by nest count data. The threat matrix approach indicated that

the South Atlantic Ocean DPS is likely to decline in the foreseeab le future (Figures 20 and 21). These results were largely driv en by the ongoing mortality threat s to juveniles from fishery bycatch that occurs throughout the South Atla ntic Ocean (Table 11). Although conservation efforts by national and international groups in th e South Atlantic are cu rrently working toward mitigating bycatch in the South Atlantic, it is unlik ely that this source of mortality can be greatly reduced in the near future, largely due to inad equate funding and knowledge gaps that together inhibit implementation of large-s cale management actions (Domingo et al . 2006).



### 4.6. Supplemental Material



## — Sensitivity Analyses

Because our threat matrix analysis was based on a deterministic model, and a few life history parameter values were determined from little em pirical data, we conducte d sensitivity analyses

to evaluate the effects of these life history parameters on the results of the threat matrix analysis.

4.6.1. Effects of AFR and CV of AFR on the Base Population Model

Figures S1-S7 indicate relations hips among AFR, CV of AFR, av erage adult survival rates, average juvenile survival rates, and the dominan t eigenvalues. Effects of AFR on the dominant86 eigenvalues were greater than t hose of CV of AFR (differences are greater among columns than rows in Figures S1-S7). Differences among DPSs were small as was expected from the small differences in vital rates among the DPSs.

For a deterministic age or stage based matrix model, the asymptotic population growth rate (or the dominant eigenvalue) is a functi on of survival rates, transition probabilities, and fertility. Consequently, for reasonable ranges of surv ival rates, the upper bound of the dominant eigenvalue can be found. Fertility parameters of the base matrices were obtained from averages (clutch size, remigration intervals, and nests per female). Conseque ntly, if we were to construct a stochastic demographic model, our base matrix would be treated as the average matrix. The dominant eigenvalue of the average matrix is gr eater than or equal to the average dominant eigenvalue from stochastic matrices (Jensen’ s inequality). Cons equently, the dominant eigenvalue from our matrix, based on the average life history parameters used in this analysis, may be considered as the maximum possible.

4.6.2. Effects of AFR on the Threat Matrix Analysis

Figures S8-S11 indicate changes in results of the threat matrix analysis due to the assumed AFR. For S8 and S9, AFR was assumed to be 25 years, whereas the CV of AFR was fixed at 15%.

Because of the small difference cau sed by CV of AFR from the prev ious analysis, results of the sensitivity analysis for CV of AFR are not shown. Greater proportions of dominant eigenvalues

are above 1 as the AFR decreases.

4.6.3. Proportion of Neriti c Juvenile Stage Duration

To incorporate different levels of anthropoge nic mortalities between neritic and oceanic environment in the threat matrix analysis, the ma trix model required the durations of neritic and oceanic juvenile stages. Because these estimate s are based on a combination of published data, extrapolations, and experts’ opinions, and because empirical data were not available for several DPSs, we used the Northwest Atlantic Ocean DPS ( 19 years in neritic habitat) as a proxy for five DPSs (Table 1). To determine the sensitivity of th e default value to the results, the threat matrix analysis was conducted while the proportion of ti me juveniles spend in neritic habitat was changed from 10% to 90%.

Lower bounds of the dominant eigenvalues decreased as juveniles spent more time in the neritic habitat (Figures S12-S13). This result was cau sed by the higher anthrop ogenic mortalities in

neritic habitats for all DPSs than in ocean ic habitats. The upper bounds of the dominant eigenvalues did not change as much as the lower bound, except the South Atlantic Ocean DPS.87 Figure S1. Relationships among adult survival rate, juvenile survival rate, age at first reproduction (AFR), coefficient of variation (CV) of AFR, and the dominant eigenvalue for the

North Pacific Ocean DPS. Contour lines indi cate the dominant eigenvalues, whereas the diagonal lines indicate equa l survival rates between juveniles and adults. Dashed horizontal lines

indicate the assumed adult survival rate for the base matrix model. The oceanic survival rate during the first year was fixed at 0.4/year.88 Figure S2. Relationships among adult survival rate, juvenile survival rate, age at first reproduction (AFR), coefficient of variation (CV) of AFR, and the dominant eigenvalue for the

South Pacific Ocean DPS. Contour lines indi cate the dominant eigenvalues, whereas the diagonal lines indicate equa l survival rates between juveniles and adults. Dashed horizontal lines

indicate the assumed adult survival rate for the base matrix model. The oceanic survival rate during the first year was fixed at 0.4/year.89 Figure S3. Relationships among adult survival rate, juvenile survival rate, age at first reproduction (AFR), coefficient of variation (CV) of AFR, and the dominant eigenvalue for the North Indian Ocean, Southeast Indo-Pacific Ocean, and Southwest Indian Ocean DPSs. Contour lines indicate the dominant eigenvalues, whereas th e diagonal lines indicate equal survival rates between juveniles and adults. Dashed horizontal lines indicate the assumed adult survival rate

for the base matrix model. The oceanic survival rate during the first year was fixed at 0.4/year.90 Figure S4. Relationships among adult survival rate, juvenile survival rate, age at first reproduction (AFR), coefficient of variation (CV) of AFR, and the dominant eigenvalue for the

Northwest Atlantic Ocean DPS. Contour lines indicate the domi nant eigenvalues, whereas the diagonal lines indicate equa l survival rates between juveniles and adults. Dashed horizontal lines

indicate the assumed adult survival rate for the base matrix model. The oceanic survival rate during the first year was fixed at 0.4/year.91 Figure S5. Relationships among adult survival rate, juvenile survival rate, age at first reproduction (AFR), coefficient of variation (CV) of AFR, and the dominant eigenvalue for the

Northeast Atlantic Ocean DPS. Contour lines indicate the dominant eigenvalues, whereas the diagonal lines indicate equa l survival rates between juveniles and adults. Dashed horizontal lines

indicate the assumed adult survival rate for the base matrix model. The oceanic survival rate during the first year was fixed at 0.4/year.92 Figure S6. Relationships among adult survival rate, juvenile survival rate, age at first reproduction (AFR), coefficient of variation (CV) of AFR, and the dominant eigenvalue for the

Mediterranean Sea DPS. Contour lines indicat e the dominant eigenvalues , whereas the diagonal lines indicate equal survival ra tes between juveniles and adults. Dashed horizontal lines indicate

the assumed adult survival rate for the base matr ix model. The oceanic survival rate during the first year was fixed at 0.4/year.93 Figure S7. Relationships among adult survival rate, juvenile survival rate, age at first reproduction (AFR), coefficient of variation (CV) of AFR, and the dominant eigenvalue for the

South Atlantic Ocean DPS. Contour lines in dicate the dominant eigenvalues, whereas the diagonal lines indicate equa l survival rates between juveniles and adults. Dashed horizontal lines

indicate the assumed adult survival rate for the base matrix model. The oceanic survival rate during the first year was fixed at 0.4/year.94 Figure S8. Ranges of dominant ei genvalues for nine DPSs with anthropogenic mortalities when the maximum population growth rates are assumed to be 5% per year. The mean age at first reproduction was assumed to be 25 years and its st andard deviation 3.75 (CV=15%). The values above the dotted line indicates the pro portion of the bars that are above λ = 1.0.95 Figure S9. Ranges of dominant ei genvalues for nine DPSs with anthropogenic mortalities when the maximum population growth rates are assumed to be 10% per year. The mean age at first reproduction was assumed to be 25 years and its standard deviation 3.75 (CV = 15%). The

values above the dotted line indicates the proportion of the bars that are above λ = 1.0.96 Figure S10. Ranges of dominant eigenvalues for nine DPSs with anthropogenic mortalities when the maximum population growth rates are assumed to be 5% per year. The mean age at first reproduction was assumed to be 35 years and its standard deviation 5.25 (CV = 15%). The

values above the dotted line indicates the proportion of the bars that are above λ = 1.0.97 Figure S11. Ranges of dominant eigenvalues for nine DPSs with anthropogenic mortalities when the maximum population growth rates are assumed to be 10% per year. The mean age at first reproduction was assumed to be 35 years and its standard deviation 5.25 (CV = 15%). The

values above the dotted line indicates the proportion of the bars that are above λ = 1.0.98 Figure S12. Effects of the changes in the propor tion of time juveniles spend in neritic habitat on results of the threat matrix analysis. For each DPS, vertical lines i ndicate the range of the dominant eigenvalues for the best and worst thre at scenarios. Each vertical line is for a proportion of neritic juvenile stage, where the proportion is from 10% to 90% with 10% increments (from left to right). Empty boxes are based on the values in Table 1. The mean age at first reproduction was assumed to be 30 year s and its standard deviation 5 (CV=17%). The asymptotic population growth rate was assumed to be 5% per year.99 Figure S13. Effects of the propor tion of time juveniles spend in neritic habitat on results of the risk analysis. For each DPS, vertical lines indi cate the range of the dominant eigenvalues for the

best and worst threat scenarios. Each vertical line is for a proportion of neritic juvenile stage, where the proportion is from 10% to 90% with 10 % increments (from left to right). Empty

boxes are based on the values in Table 1. The m ean age at first reproduction was assumed to be 30 years and its standard deviation 5 (CV=17%). The asymptotic population growth rate was assumed to be 10% per year.100 

## SECTION 5—ANALYSIS OF FACTORS LISTED UNDER ESA SECTION 4(a)(1)

Under Section 4(a)(1) of the ESA, NMFS and FWS are required to determine whether any species is an endangered or threatened species because of any of the following factors. These

factors are: (A) the present or threatened destruction, modification, or curtailment of its habitat or

range; (B) overutilization for commercial, recreati onal, scientific, or educational purposes; (C) disease or predation; (D) the inadequacy of existing regulatory mechanisms; or (E) other natural or manmade factors affecting its continued existence.

The following is an analysis of these factors as they relate to the current status of the nine loggerhead DPSs:

1. North Pacific Ocean

2. South Pacific Ocean

3. North Indian Ocean

4. Southeast Indo-Pacific Ocean

5. Southwest Indian Ocean

6. Northwest Atlantic Ocean

7. Northeast Atlantic Ocean

8. Mediterranean Sea

9. South Atlantic Ocean

In Section 5.1., a general descrip tion of the threats that occur fo r all DPSs is presented under the relevant section 4(a)(1) factor to avoid duplication. In Section 5.2, threats that are specific to a particular DPS are presented by D PS under each section 4(a)(1) factor.



### 5.1. General Description of Threats

5.1.1. Present or Threatened Destruction, Modi fication, or Curtailment of Its Habitat or Range

Terrestrial Zone Destruction and modification of loggerhead ne sting habitats are occu rring worldwide throughout the species range. The main an thropogenic threats impacting logge rhead nesting habitat include coastal development/construction, placement of eros ion control structures and other barriers to nesting, beachfront lighting, vehicular and pede strian traffic, sand extraction, beach erosion, beach sand placement, beach pollution, removal of native vegetation, and planting of non-native vegetation (Baldwin 1992; NMFS a nd FWS 1998, 2008; Margaritoulis et al . 2003).

Worldwide, development near coastal areas conti nues and is a major problem as more and more people are moving to or visiting coastal areas. Coastal development may include, but is not limited to, the construction of roads, highways, public infrastructure, hotels, condominiums, houses, harbors, nearshore shorelin e stabilization structures (e.g., groins, jetties, breakwaters),101 and seawalls and other forms of coastal armori ng. All of these various forms of coastal construction alter nesting habitat to one degree or another, typi cally making it less suitable for nesting females, egg incubation, and hatchling emergence. This development can also lead to deleterious effects of animals at-sea through incr eased fishing activity near important foraging areas (Peckham et al . 2007), along migratory corridors, and coastally, near the nesting beaches (e.g., Kamezaki et al . 2003).

Erosion of nesting beaches can occur as a result of coastal development when native dune vegetation, which enhances beach stability and ac ts as an integral buffer zone between land and sea, is degraded or destroyed. This in turn often leaves insufficient nesting opportunities above the high tide line, and nests may be washed out. In contrast, the plan ting or invasion of less stabilizing, non-native plants can lead to increased erosion and degradation of suitable nesting habitat (Schmelz and Mezich 1988). In addition, ta ller plants and structures can increase shading

and alter the natural sex rati os of hatchlings (Mrosovsky et al . 1995). Non-native vegetation also may form impenetrable root mats that can preven t proper nest cavity excavation, as well as trap hatchlings attempting to emerge from nests.

Beach nourishment also affects the incubation e nvironment and nest success. Although the placement of sand on beaches may provide a greater quantity of nesting habitat, the quality of that habitat may be less suitable than pre-existing natural beaches. Constructed beaches tend to differ from natural beaches in several important ways. They are typically wider, flatter, more compact, and the sediments are more mois t than those on natu ral beaches (Nelson et al . 1987,

Ackerman et al . 1991, Ernest and Martin 1999). On seve rely eroded sections of beach, where little or no suitable nesting habitat previously existed, sand placement can result in increased nesting (Ernest and Martin 1999). However, on mo st beaches, nesting success typically declines for the first year or two followi ng construction, even though more ne sting habitat is available for turtles (Trindell et al . 1998, Ernest and Martin 1999, Herren 1999). Reduced nesting success on

constructed beaches has been at tributed to increased sand compaction, escarpment formation, and changes in beach profile (Nelson et al . 1987, Crain et al . 1995, Lutcavage et al . 1997,

Steinitz et al . 1998, Ernest and Martin 1999, Rumbold et al . 2001). Compaction can inhibit nest

construction or increase the amount of time it ta kes for turtles to construct nests, while escarpments often cause female turt les to return to the ocean wit hout nesting or to deposit their nests seaward of the escarpment where they are more susceptible to frequent and prolonged tidal inundation. In short, sub-optimal nesting habita t may cause decreased ne sting success, place an increased energy burden on nesting females, resu lt in abnormal nest c onstruction (Carthy 1996), and reduce the survivorship of eggs and hatchlin gs. In addition, sand used to nourish beaches may have a different composition than the original beach; thus introducing lighter or darker sand, consequently affecting the relative ne st temperatures (Ackerman 1997, Milton et al . 1997).

Crain et al . (1995) provides a review of the potential e ffects of beach nourishment on sea turtles.

Groins and jetties, shore-perpendicular structures designed to trap sand th at would otherwise be transported by longshore currents, interfere with nesting turtle access to the beach, result in a change in beach profile and width (e.g., downdrif t erosion, loss of sandy berm s, and escarpment formation) (Komar 1983, Pilkey et al . 1984, National Research Council 1987), trap hatchlings, and concentrate predatory fishes, resulting in hatchling predation.102 The construction of beachfront armoring (i.e., rigi d structures placed parallel to the shoreline on the upper beach to prevent both landward retreat of the shoreline and inundation or loss of upland property by flooding and wave action) incl udes bulkheads, seawalls, soil retaining walls, rock revetments, sandbags, and geotextile tubes. These structures can greatly impact nesting opportunities and hatching success of loggerhead tu rtles. Mosier (1998) reported that fewer loggerheads made nesting attempts on beaches fronted by seawalls than on adjacent beaches where armoring structures were absent. Mosi er (1998) and Mosier and Witherington (2002) found that when turtles did emerge in the presence of armoring structures, more returned to the water without nesting than those on non-armore d beaches. Additionally, Mosier (1998) found

that turtles on armored sections of beach tende d to wander greater distances than those that emerged on adjacent natural beaches. Armoring stru ctures can effectively eliminate a turtle’s access to upper regions of the beach/dune system. Consequently, nests on armored beaches were generally found at lower elevations than thos e on non-walled beaches. Nests laid at lower elevations are subject to a grea ter risk of repeated tidal i nundation and erosion, which can potentially alter thermal regimes, and thus sex ratios (Mrosovsky and Provancha 1992, Mrosovsky 1994, Ackerman 1997). The demand for both nourishment and the placement of hardened structures on the beach as management op tions for beach erosion are likely to increase in the future in the face of pr ojected sea level rise and more intense storm activity associated with global climate change.

Developments, such as roadways, high rise hotels, and condominium s, also contribute to habitat degradation by increasing noise and light pollution. Studies ha ve shown that light pollution disorients hatchlings, causing them to move in land away from the ocean (Witherington 1997). Hatchlings unable to find the ocean, or delayed in reaching it, incur high mortality from dehydration, exhaustion, or pred ation (Ehrhart and Witherington 1987, Witherington and Martin 1996). Hatchlings lured into li ghted parking lots or toward streetlights are often crushed by passing vehicles (Witherington and Martin 1996). In addition, li ght pollution deters nesting females from emerging onto the b each to nest (Witherington 1992).

Burgeoning numbers of visitors to beaches may cause sand compaction and nest trampling (Kudo et al . 2003). In addition, the placement of recreational beach equipment (e.g., lounge chairs, cabanas, umbrellas, catamarans) degrades the suitability of beaches as nesting habitat by hampering or deterring nesting turtles fr om accessing the upper beach (Sobel 2002; Margaritoulis et al . 2007; Florida Fish and Wildlife Conservation Commission, unpublished data), thus limiting the potential area for nesting. Debris on nesting beaches is also a threat to hatchlings and nesting females. Hatchlings ofte n must navigate through a variety of obstacles before reaching the ocean. These include natural and human-made debris. Debris on the beach may interfere with a hatchling’ s progress toward the ocean. Research has shown that travel times of hatchlings from the nest to the wate r may be extended when tr aversing areas of heavy

foot traffic or vehi cular ruts (Hosier et al . 1981); the same is true of debris on the beach. Hatchlings may be upended and spend both time a nd energy in righting themselves. Some beach debris may have the potential to trap hatchlings and prevent th em from successfully reaching the ocean. In addition, debris over the tops of nest s may impede or prevent hatchling emergence. Nesting females also may be impeded by beach de bris in their movement up or down the beach, and this may influence selection of nest sites. In rarer circumstances, large debris items may entrap nesting females.103 Operating public vehicles on nesting beaches for recreational purposes or beach access also degrades nesting habitat. The ruts left by vehicles in the sand pr event 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). Hatchlings impeded by vehicle ruts are at gr eater risk of death from predation, fatigue, desiccation, and bei ng crushed by additional vehicle traffic. Light pollution from vehicle lights on the beach afte r dark may deter females from nesting and disorient hatchlings. Sand compaction due to ve hicles on the beach may hinder nest construction and hatchling emergence from nests. Driving di rectly above incubating egg clutches can cause sand compaction, which may decrease hatching and emerging success and directly kill preemergent hatchlings (Mann 1977). Additiona lly, vehicle traffic on nesting beaches may contribute to erosion, especially during high tides or on narrow beaches where driving is concentrated on the high beach and foredune.

Neritic/Oceanic Zone Habitat degradation and destruction in the neri tic and oceanic zones occur less conspicuously than they do in the terrestrial zone. Direct impacts to bottom habitats occur from activities including bottom trawl and dredge fishing, channel dredging, a nd sand extraction. Indirect effects can result from both point and non-point source pollution (e.g., upland runoff, direct sewage discharge) associated with coastal deve lopment. The impacts of climate change may also result in trophic level al terations, and therefore may af fect loggerhead prey abundance and/or distribution. There are currently no da ta available to quantify impacts from these potential trophic level alterations.

Other fishing methods such as drift and set gillnet s, longlines, and pots/traps affect both neritic and oceanic zones by not only incidentally cap turing loggerheads but also depleting fish populations, and thus altering ecosystem dynamics. In many cases, loggerhead foraging areas coincide with fishing zones. Selective and usually intense harvest of species in fisheries will result in changes in neritic and oceanic ecosy stems (e.g., predator-prey interactions, trophic dynamics and food webs; see Bjorndal 2003). Se ney and Musick (2007) reported changes in loggerhead diet resulting from a lterations of prey availability in the Chesapeake Bay, Virginia, USA, an important foraging habitat. Perhaps th e most destructive fishing methods in neritic ecosystems are bottom trawling (Auster 1998, Brailovskaya 1998, Engel and Kvitek 1998, Kaiser 1998, Pilskaln et al . 1998, Schwinghamer et al . 1998, Watling and Norse 1998) and dredging (Murray 2004, 2005, 2007). Benthic ha bitat alteration by mobile fishing gear, especially trawls and dredges, constitutes a globally significant physical disturbance to the marine environment and has significant effect s on marine biodivers ity (Watling and Norse 1998). The ecological effects of trawling and dr edging on the marine environment have been

likened to the terrestrial ecol ogical effects of clearcutting fo rests (Watling and Norse 1998). Mobile fishing gear has been shown to resu lt in short and long-term changes in benthic community composition, including species groups on which loggerheads forage (Gordon et al

. 1998). The National Research Coun cil (1994) found that ha bitat alteration by fi shing activities is perhaps the least understood of th e important environmental effect s of fishing. Comprehensive data are lacking to fully unders tand and quantify these impacts.

Marine pollution can affect logg erhead habitats in both the neritic and oceanic zones. These impacts include contamination from herbicides, pest icides, oil spills, and other chemicals, as well104 as impacts on water quality (e.g., increases in water column sediments) resulting from structural degradation from excessive boat anchoring, dredging, and marine explosives (Francour et al

. 1999, Lee Long et al . 2000, Margaritoulis et al . 2003, Waycott et al . 2005).

5.1.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

Deliberate hunting of loggerheads for their meat, shells, and eggs is reduced from previous exploitation levels, but still exists. The level of harvest varies significantly by region. Despite

national laws in various countries , in many areas the po aching of eggs and hun ting of adults and juveniles is still a problem. Details on thes e threats are described by DPS in Section 5.2.

5.1.3. Disease or Predation

The population level effects of diseases in logg erheads are not known. At least two bacterial

diseases have been described in wild loggerhead populations, incl uding bacterial encephalitis and

ulcerative stomatitis/obstructive rhinitis/pneumonia (George 1997). There are few reports of fungal infections in wild loggerhead populations (Homer et al . 2000). Both bacterial and fungal infections are common in cap tive sea turtles (H erbst and Jacobson 1995, George 1997).

Viral diseases have not been documented in free-ranging loggerheads, with the possible exception of sea turtle fibropapillomatosis, which may have a vira l etiology (Herbst and Jacobson 1995, George 1997). Fibropapillomatosis is a disease that is characterized by the presence of internal and/or external tumors (fibro papillomas). External tumors can interfere with swimming, vision, and feeding. Although fibropa pillomatosis reaches epidemic proportions in some wild green turtle populations, the prevalence of this disease in most loggerhead populations is thought to be small. Mortality levels associated with the di sease are still unknown.

A variety of endoparasites, in cluding trematodes, tapeworms, and nematodes have been described in loggerheads (Herbs t and Jacobson 1995). Heavy infe stations of endoparasites may

cause or contribute to debilitati on or mortality in sea turtles.

Depredation of sea turtle eggs and hatchlings by native and introduced species occurs on a large number of nesting beaches, as described by DPS in Section 5.2. Planting of non-native plants

can have a detrimental effect in the form of roots invading eggs (e.g., tamarisk tree roots invading eggs in Zakynthos, Greece) (Margaritoulis et al . 2007).

Harmful algal blooms, such as a red tide, also impact loggerheads. Karenia brevis , a type of microalgae known as a dinoflagellate that produces a toxin, has been known to cause red tides (Florida Marine Research Institute 2003). Othe r types of microorganisms cause different kinds of harmful algal blooms in other parts of the world as well (Florida Marine Research Institute 2003, Gilbert et al . 2005).

5.1.4. Inadequacy of Existing Regulatory Mechanisms

Loggerheads are highly migratory, which makes them a shared resource among many nations. Therefore, conservation efforts for loggerhead po pulations in one country may be jeopardized by105 activities in another. Many countries lack regulat ions or have inadequate regulations in place to address the impacts of a wide range of anthr opogenic activities that di rectly injure and kill loggerheads, disrupt necessary behaviors, and alte r terrestrial and marine habitats used by the species. In particular, improved re gulations of fisheries that inci dentally capture loggerheads are needed to reduce mortality. Improved fishery obse rver coverage is also needed to provide more basic information on loggerhead bycatch. G overnment regulations and community programs need to be initiated or strengthened to address the impacts of turtle hunting and egg poaching. Overall, increased efforts are needed to assist many foreign countries with the enactment and enforcement of national regulati ons to protect loggerheads.

The conservation and recovery of sea turtles is facilitated by a number of regulatory instruments at international, regional, national, and local levels. As a result of these designations and agreements, many of the intentional impacts directed at sea turtles have been lessened: harvest of eggs and adults has been slowed at severa l nesting areas through nesting beach conservation efforts and an increasing number of community-based initiatives ar e in place to slow the take of turtles in foraging areas. Moreover, there is now a more internationally concerted effort to reduce sea turtle interactions and mortality in artisanal and industrial fishing practices.

Despite these advances, human impacts con tinue throughout the world. The lack of comprehensive and effective monitoring and bycatch reduction efforts in many pelagic and nearshore fisheries operations still allows substantial direct and indirect mortality, and the uncontrolled development of coastal and marine habitats threatens to destroy the supporting ecosystems of long-lived sea turtles. Although se veral international agreements provide legal protection for sea turtles, additional multi-lateral e fforts are needed to ensure they are sufficiently implemented and/or strengthened, and key non-si gnatory parties need to be encouraged to

accede.

Considering the worldwide distribution of loggerhead turtles, the majority of legal instruments that target or impact sea turtles cover loggerh ead turtles. A summary of the main regulatory instruments from throughout the world that relate to loggerhead turtle management is provided below. Cross-regional regulatory instruments will be highlighted here. If an agreement or act is specific to a DPS’s geographic ar ea, that action will be discussed under the appropriate DPS in Section 5.2. The pros and cons of many of these were evaluated by Hykle (2002), and a summary of his findings is given when appropriate.

United States Magnuson-Stevens Fish ery Conservation and Management Act The recently-amended U.S. Magnuson-Stevens Fi shery Conservation and Management Act (MSA), implemented by NMFS, mandates environmen tally responsible fish ing practices within U.S. fisheries and has provisions that extend to fishing activities in waters beyond U.S. jurisdiction. Section 610 calls on the U.S. Secretary of Comm erce to identify nations with fishing vessels that are engaged or have been engaged in fishing activities in waters beyond any national jurisdiction or in areas beyond the Exclusive Economic Zone of the United States. This

section authorizes NMFS to conduc t bilateral consultati ons with such nations to protect living

marine resources. If a nation continues to condu ct fishing activities that result in bycatch of protected living marine resources, they can be ce rtified to the United States Congress. A result of this certification could be suspension in th e trade of fisheries products. Finally, the Act106 specifically encourages NMFS to conduct international cooperation and assistance with foreign nations that are identified so that bycatch of protected living marine re sources can be reduced.

FAO Technical Consultation on Sea Turtle-Fishery Interactions

While not a true internationa l instrument for conservati on, the Food and Agriculture Organization of the United Nations' (FAO) t echnical consultation on sea turtle-fishery interactions was groundbreaking in th at it solidified the commitment of this international body to reduce sea turtle bycatch in marine fisheries operations. Recommendations from the technical consultation were endorsed by the FAO Comm ittee on Fisheries (COFI) and called for the immediate implementation by member nations an d Regional Fishery Mana gement Organizations (RFMOs) of guidelines to reduce sea turtle morta lity in fishing operations, developed as part of the technical consultation. Compliance with these guidelines is voluntary.

Convention on the Conservation of Mi gratory Species of Wild Animals This Convention, also known as the Bonn Conventi on or CMS, is an international treaty that focuses on the conservation of migratory specie s and their habitats. As of March 2008, the Convention had 108 member states, including parties from Africa, Central and South America, Asia, Europe, and Oceania. While the Conventi on 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 Brazil, Canada, China, Indonesia, Japan, Mexico, Oman, and the United States. Additional information

is available at http://www.cms.int

.

Convention on Biological Diversity (CBD) The primary objectives of this international tr eaty are: (1) the cons ervation of biological diversity, (2) the sustainable use of its components, and (3) the fair and eq uitable sharing of the benefits arising out of the uti lization of genetic resources. Th is Convention has been in force since 1993 and had 191 Parties as of June 2008. While the Convention provides a framework within which broad conservation objectives may be pursued, it does not sp ecifically address sea turtle conservation (Hykle 2002). Add itional information is available at http://www.cbd.int

.

Convention on International Trade in Endangere d 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 implemen ted in 1975 and currently includes 173 Parties. Although CITES has been effective at minimizing the international trade of sea turtle products, it does not limit legal harvest with in countries, nor does it regulate intra-country commerce of sea turtle products (Hykle 2002). Additi onal information is available at http://www.cites.org

.

United Nations Convention on the Law of the Sea (UNCLOS) To date, 155 countries, including most mainland countries lining the west ern Pacific, and the European Community have joined in the conven tion. The United States has signed the treaty, 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 protec ting freedom of scientific research on the high

seas. Additional information is available at http://www.un.org/Depts/los/index.htm

.107 United Nations Resolution 44/225 on Large-Scale Pelagic Driftnet Fishing In 1989, the United Nations called, in an unanim ous resolution, for the elimination of all high seas driftnets by 1992. Additiona l information is available at http://www.intfish.ne t/treaties/ga44_225.htm

.

The International Convention for the Pr evention of Pollution from Ships (MARPOL)

The MARPOL Convention is a combination of tw o treaties adopted in 1973 and 1978 to prevent pollution of the marine environment by ships fr om operational or accidental causes. The 1973 treaty covered pollution by oil, chemicals, harmful substances in packaged form, sewage and garbage. The 1978 MARPOL Protocol was ad opted at a Conference on Tanker Safety and Pollution Prevention which included standards for tanker design and operation. The 1978 Protocol incorporated the 1973 C onvention as it had not yet been in force and is known as the

International Convention for the Prevention of Ma rine Pollution from Ships, 1973, as modified by the Protocol of 1978 relating thereto (MAR POL 73/78). The 1978 Convention went into force in 1983 (Annexes I and II). The Conventi on includes regulations ai med at preventing and minimizing accidental and routine operations pollu tion from ships. Amendments passed since have updated the convention. To date there are six Annexes with Annexes I and II being mandatory for State Parties a nd the others being voluntary.

Annex I Regulations for the Prevention of Pollution by Oil Annex II Regulations for the C ontrol of Pollution by Noxious Liquid Substances in Bulk Annex III Prevention of Pollution by Harmful Substances Carried by Sea in Packaged Form Annex IV Prevention of Pollution by Sewage from Ships Annex V Prevention of Pollution by Garbage from Ships

Annex VI Prevention of Ai r Pollution from Ships

The Convention Concerning the Protection of th e World Cultural and Natural Heritage (World Heritage Convention) The World Heritage Convention was signed in 1 972 and, as of November 2007, 185 states were parties to the Convention. The instrument requires parties to ta ke effective and active measures to protect and conserve habitat of threatened species of animal s and plants of scientific or aesthetic value. The World Heritage Conventi on currently includes 31 ma rine sites, including important marine turtle habitat such as the Belize Barrier Reef Reserve System, Belize.

Additional information is available at http://whc.unesco.or g/en/conventiontext

.

Ramsar Convention on Wetlands The Convention on Wetlands, signed in Ramsar, Iran, in 1971, is an intergovernmental treaty, which provides the framework for national act ion and international cooperation for the

conservation and wise use of wetlands and their re sources. Currently, there are 158 parties to the convention, with 1,752 wetland sites, including important marine tur tle habitat such as the Turtle Beaches and Coral Reefs of Tongaland, South Afri ca. Additional information is available at http://www.ramsar.org

.108 5.1.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear The most significant manmade factor affecting th e conservation and recovery of the loggerhead is incidental capture in commer cial and artisanal fisheries. Incidental capture (bycatch) of loggerheads occurs in various fi sheries throughout the range of th e species. Fishing gear types that may affect loggerheads in clude pelagic and demersal long lines, drift and set nets (e.g., gillnets, trammel nets), bottom and mid-water tr awling, fishing dredges, pound nets and weirs, haul and purse seines, pots and traps, and hook and line gear. While significant progress has been made to reduce bycatch in some fisheries in certain parts of the l oggerhead’s range, serious bycatch problems remain unaddressed. Specific information on bycatch ra tes or the number of takes is presented by DPS in Section 5.2.

Other Manmade and Natural Impacts In addition to fishery bycatch, climate change is another factor that has the potential to greatly impact loggerhead turtles. Impacts from global climate change induced by human activities are likely to become more apparent in future ye ars (Intergovernmental Panel on Climate Change (IPCC) 2007a). The global mean temperature has risen 0.76oC over the last 150 years, and the linear trend over the last 50 years is nearly twice that for the last 100 years (IPCC 2007a). There is a high confidence, based on substantial new evidence, that observed changes in marine systems are associated with rising water temperat ures, as well as related changes in ice cover, salinity, oxygen levels, and circulation. Ocean acidification resulting from massive amounts of carbon dioxide and pollutants rel eased into the air can have major adverse impacts on the calcium balance in the oceans and, thus, on sea turtles. Changes to the marine ecosystem due to climate change include shifts in ranges and changes in algal, plankton, and fish abundance (IPCC 2007b), which could affect loggerhead prey distribution and abundance.

One of the most certain consequences of climate change is rising sea leve ls (Titus and Narayanan 1995), which will result in increased erosion rates along nesting beach es. This could particularly impact areas with low-lying beaches where sa nd depth is a limiting f actor, as the sea will inundate nesting sites an d decrease available nesting habitat (Daniels et al . 1993, Fish et al

. 2005, Baker et al . 2006). The loss of habitat as a result of climate ch ange could be accelerated

due to a combination of other environmental and oceanographic changes such as an increase in the frequency of storms and/or changes in prevailing currents, both of which could lead to increased beach loss via erosion (Antonelis et al . 2006, Baker et al . 2006). On some undeveloped beaches, shoreline migration will have limited effects on the suitability of nesting habitat. Bruun (1962) hypothesized that during a sea level rise a typical beach profile will maintain its configuration but will be transl ated landward and upward. However, along developed coastlines, and especially in areas where erosion control structures have been constructed to limit shoreline movement, rising s ea levels will cause severe effects on nesting females and their eggs. Erosion control structures can result in the perman ent loss of dry nesting beach or deter nesting females from reaching su itable nesting sites (N ational Research Council 1990). Nesting females may deposit eggs seaward of the erosion control structures potentially subjecting them to repeated tidal inundation.109 Climate change also may affect loggerhead sex ra tios. Loggerhead turt les exhibit temperaturedependent sex determination. Rapidly increas ing global temperatures may result in warmer incubation temperatures and highly female-bia sed sex ratios (e.g., Glen and Mrosovsky 2004,

Hawkes et al . 2009).

Loggerhead sea turtles may be killed or injure d from collisions with boat hulls and propellers. These interactions are becoming more common. Capture and mortality of loggerheads during periodic dredging of sediments fr om navigational channels also have been documented. Power plants have also been found to impinge a nd/or entrain loggerhead s on/in their intake

screens/structures.

Marine pollution, including mari ne debris, oil spills, and bioaccu mulative chemicals, is one of the main anthropogenic threats to sea turtles (Tomás et al . 2002). Because of their habitat and feeding behavior, loggerheads appear to be one of two sea turtle sp ecies that ingest more debris in all of its life stages (Lutcavage et al . 1997). Direct or indirect di sposal of anthropogenic waste introduces potentially lethal materials into loggerhead foraging habitats or into convergence zones, affecting oceanic juveniles. Loggerhead s will ingest plastic pieces, styrofoam pieces, and other marine debris. Ingestion occu rs when debris is mistaken for or associated with prey items. Effects may be lethal or non-le thal, resulting in varying side effects that may increase the probability of death (Balazs 1985, Carr 1987, McCa uley and Bjorndal 1999, Witherington 2002). For example, nutrient dilution occurs when non-nut ritive debris displaces food in the gut, which then may decrease somatic growth and reproduc tion (McCauley and Bjorndal 1999). Sea turtles at all life stages appear to be highly sensitive to oil spills, perhap s due to certain aspects of their biology and behavior, including a lack of avoidance behavior, indiscriminate feeding in convergence zones, and large pre-dive inhalations (Milton et al . 2003). Milton et al . (2003) state

that the oil effects on turtles in clude increased egg mortality and developmental defects, direct mortality due to oiling in hatchl ings, juveniles and adults, and im pacts to the skin, blood, salt glands, and digestive and immune systems. In addition, loggerheads can become entangled in marine debris, such as discarded fishi ng gear or other entangling materials.

All loggerhead life history stages are vulnerable to the harmful effects of oil through direct contact, degradation of food res ources, and loss of habitat (Vargo et al . 1986, Minerals Management Service 2000). Vargo et al . (1986) reported that sea turt les would be at substantial risk if they encountered an oil sp ill or large amounts of tar in the environment. In a review of

available information on debris i ngestion, Balazs (1985) reported that tar balls were the second most prevalent type of debris i ngested by sea turtles. Physiologi cal experiments showed that sea turtles exposed to petroleum products may suffer inflammatory dermatitis, ventilatory disturbance, salt gland dysfuncti on or failure, red blood cell distur bances, immune response, and digestive disorders (Vargo et al . 1986, Lutz and Lutcavage 1989, Lutcavage et al . 1995). Operational discharge of produced waters, drill m uds, and drill cuttings are routinely discharged in marine waters as a result of petroleum pr oduction activities (Mineral s Managem ent Service 2000). Loggerheads may bioaccumulate heavy me tals found in drill muds resulting in

debilitation or death. The e ffects of large-scale oil spills on loggerheads could be locally catastrophic. The impacts of offshore light ed oil production platforms on loggerheads are unknown. Lighted platforms may a ttract hatchlings, making them mo re susceptible to predation110 (de Silva 1982). Neritic juveni les and adults may be attracted by high prey concentrations around the structures, making them more suscep tible to ingestion of petroleum products.

Natural factors that have the potential to aff ect loggerhead recovery include the effects of aperiodic hurricanes, seasonal typhoons, and cat astrophic environmental events such as tsunamis. In general, these events are episodic and, although they may affect loggerhead hatchling production, the results are generally localized and they rarely result in whole-scale losses over multiple nesting seasons. The negative effects of these types of events on low-lying and/or developed shorelines may be longe r-lasting and a greate r threat overall.

Loggerheads are also susceptibl e to cold stunning, a phenome non in which turtles become incapacitated as a result of rapidly dropping water temperatur es (Witherington and Ehrhart 1989,

Morreale et al . 1992). As temperatures fall below 8-10 o C, turtles may lose their ability to swim and dive, often floating to the surface. It appears to be the rate of cooling that precipitates cold stunning rather than the water temperature itse lf (Milton and Lutz 2003) . Sea turtles that overwinter in inshore waters are most susceptibl e to cold stunning, because temperature changes are most rapid in shallow wate r (Witherington and Ehrhart 1989).



### 5.2. Threats Specific to DPS

5.2.1. North Pacific Ocean DPS

5.2.1.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone As described in Section 5.1.1., destruction and modification of loggerhead nesting habitat in the North Pacific result from coastal development and construction, placement of erosion control

structures and other barriers to nesting, beachfront lighting, vehicu lar and pedestrian traffic, sand extraction, beach erosion, beach sand placement, beach pollution, removal of native vegetation, and planting of non-native vegetation (NM FS and FWS 1998). Beaches in Japan where loggerheads nest are extensively eroded due to dredging and dams constructed upstream, and are obstructed by seawalls as well. Unfortunately, no quantitative studies ha ve been conducted to determine the impact to the loggerhead nesting populations (Kamezaki et al . 2003). However, it is clear that loggerhead nesting habitat has been impacted by er osion and extensive beach use by tourists, both of which have contributed to unusually high mortality of eggs and pre-emergent hatchlings at many Japanese rookeries (Matsuzawa 2006).

Maehama Beach and Inakahama Beach on Yakushima in Kagoshima Prefecture account for approximately 30% of loggerhead nesting in Japan (Kamezaki et al . 2003), making Yakushima an important area for nesting beach protection. However, the beaches suffer from beach erosion and light pollution, especially from passing cars , as well as from tourists encroaching on the nesting beaches (Matsuzawa 2006). Burgeoning nu mbers of visitors to beaches may cause sand compaction and nest trampling. Egg and pre-em ergent hatchling mortality in Yakushima has been shown to be higher in areas where public access is not restricted and is mostly attributed to human foot traffic (“ stomping”) on nests (Kudo et al . 2003). Fences have been constructed111 around areas where the highest densiti es of nests are laid; however, there are still lower survival rates of eggs and pre-emergent hatchlings due to excessive foot traffi c (Ohmuta 2006). Recently, the Ministry of Environment has supported th e local non-governmental organization conducting turtle surveys and conservation on Yakushima in establishing guidelines for surveys and minimizing impacts by humans encroaching on th e nesting beaches. As of the 2009 nesting season, humans accessing Inakahama, Maehama, and Yotsuse beaches at night must comply with the established rules (Y. Matsuzawa, Sea Turtle Association of Japan, personal communication, 2009).

Loggerhead nesting habitat also has been lost at important rookeries in Miyazaki due in part to port construction that involved deve lopment of a groin of one kilometer from the coast into the sea, a yacht harbor with breakwaters and artifi cial beach, and an airport, causing erosion of beaches on both sides of the construction zone. This once excellent breeding habitat for loggerheads is now seriously thr eatened by erosion (Takeshita 2006).

Minabe-Senri beach, Wakayama Prefecture is a “submajor” nesting beach (in Kamezaki et al

. 2003), but is one of the most important rookeries on the main island of Japan (Honshu). Based on unpublished data, Matsuzawa (2006) reported ha tching success of unwashed-out clutches at Minabe-Senri beach to be 24% in 1996, 50% in 1997, 53% in 1998, 48% in 1999, 62% in 2000, 41% in 2001, and 34% in 2002.

Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, channel dredging, sa nd extraction, marine pollution, and climate change. These threats also occur in the No rth Pacific Ocean. Fishing methods not only incidentally capture loggerheads (Section 5.1.5. ), but also deplete invertebrate and fish populations and thus alter ecosystem dynamics . In many cases loggerhead foraging areas coincide with fishing zones. For example, usi ng aerial surveys and satellite telemetry, juvenile foraging hotspots have recently been identified o ff the coast of Baja California, Mexico; these hotspots overlap with intensive small-scale fisheries (Peckham and Nichols 2006; Peckham et al

. 2007, 2008). Comprehensive data currently are unavailable to fully understand how intense harvesting of fish resources changes neritic and oceanic ecosystems. Climate change also may result in future trophic change s, thus impacting loggerhead prey abundance and/or distribution.

5.2.1.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

In Japan, the use of loggerhead meat for f ood is not popular except historically in local communities such as Kochi and Wakayama prefectu res. In addition, egg collection was common in the coastal areas during times of hunger and later by those w ho valued loggerhead eggs as revitalizers or aphrodisiacs and acquired them on the black market (in Kamezaki et al . 2003, Takeshita 2006). Currently, due in large part to research and conservation efforts throughout the country, egg harvesting no longer represents a problem in Japan (Kamezaki et al . 2003, Ohmuta 2006, Takeshita 2006). Laws were enacted in 19 73 to prohibit egg collection on Yakushima, and in 1988, the laws were extended to the entire Kagoshima Prefecture, where two of the most important loggerhead nesting beaches are protected (Matsuzawa 2006).112 Despite national laws, in many ot her countries where loggerhead s are found migrating through or foraging, the hunting of adults and juveniles is st ill a problem, as seen in Baja California Sur, Mexico (Koch et al . 2006). Sea turtles have been protected in Mexico since 1990, when a federal law decreed the prohibition of the “extraction, capture and pursuit of all species of sea

turtle in federal waters or from beaches with in national territory ... [and a requirement that] ... any species of sea turtle inci dentally captured during the opera tions of any commercial fisheryshall be returned to the sea, independently of its physical state, dead or alive” (in GarciaMartinez and Nichols 2000). Despite the ban, studies have shown that sea tu rtles continue to be caught, both indirectly in fisherie s and by a directed ha rvest of juveniles a nd subadults. Turtles are principally hunted using nets , longlines, and harpoons. While some are killed immediately, others are kept alive in pens and transported to market. The market for sea turtles consists of two types: the local market (consumed locally) and th e export market (sold to restaurants in Mexico cities such as Tijuana, Ensenada, and Mexicali, and U.S. cities such as San Diego and Tucson). Consumption is highest during holidays such as Easter and Christmas (Wildcoast/Grupo Tortuguero de las Californias 2003).

Based on a combination of analys es of stranding data, beach a nd sea surveys, tag-recapture studies, and extensive intervie ws, all carried out between J une 1994 and January 1999, Nichols (2003) conservatively estimated the annual take of sea turtles by various fisheries and through direct harvest in the Baja California, Mexico, region. Sea turtle mortality data collected between

1994 and 1999 indicate that over 90% of sea turtles re corded dead were either green turtles (30% of total) or loggerheads (61% of total), and si gns of human consumption were evident in over half of the specimens. These studies resulted in an estimated 1,950 loggerheads killed annually, affecting primarily immature size classes. The primary causes for mortality were the incidental take in a variety of fishing g ears and direct harvest for consum ption and [illegal] trade (Nichols 2003).

From April 2000 to July 2003 throughout the Bahi a Magdalena region (including local beaches and towns), researchers found 1,945 sea turtle carca sses, 44.1% of which were loggerheads. Of

the sea turtle carcasses found, slaughter for human consumption was the primary cause of death for all species (63% for loggerheads). Over 90% of all turtles found were juveniles or subadults (Koch et al . 2006). As the population of gr een turtles has declined in Baja California Sur waters, poachers have switched to loggerh eads (H. Peckham, Pro Penins ula, personal communication, 2006).

5.2.1.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the North Pacific Ocean. As in other nesting locations, egg predation also exists in Japan, particularly by raccoon dogs ( Nyctereutes procyonoides ) and weasels ( Mustela

itatsi ); however, quantitative data do not exist to evaluate the im pact on loggerhead populations (Kamezaki et al . 2003). Loggerheads in the North Paci fic Ocean also may be impacted by harmful algal blooms.113 5.2.1.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the No rth Pacific. The following additional instruments also apply to loggerheads found in this area.

United States Magnuson-Stevens Fish ery Conservation and Management Act The recently-amended U.S. Magnuson-Stevens Fi shery Conservation and Management Act (MSA), implemented by NMFS, mandates environmen tally responsible fish ing practices within

U.S. fisheries. Section 301 of the MSA esta blishes National Standard s to be addressed in management plans. Any regulations promul gated to implement such plans, including conservation and management measures, shall, to the extent practicable, (A) minimize bycatch and (B) to the extent bycatch cannot be avoide d, minimize the mortality of such bycatch. Section 301 by itself does not require specific meas ures. However, mandatory bycatch reduction measures can be incorporated into management plan s for specific fisheries, as has happened with the U.S. pelagic longline fisheries in the Atlan tic and Pacific oceans. Section 316 requires the establishment of a bycatch reduction engineerin g program to develop "t echnological devices and other conservation engineering changes designed to minimize by catch, seabird interactions, bycatch mortality, and post-release mort ality in Federally managed fisheries."

Indian Ocean – South-East Asian Marine Tu rtle Memorandum of Understanding (IOSEA) This Memorandum of Understanding (MOU) puts in place a framework through which States of the Indian Ocean and South-East Asian region, as well as other concerned States, can work together to conserve and replenish depleted marine turtle populations for which they share responsibility. This collaboration is achieve d through the collective implementation of an associated Conservation and Management Plan. Currently, there are 27 signatory states. The United States became a signatory in 2001. Numerous accomplishments have been made under the auspices of this MOU. For detailed information, visit the IOSEA website at http://www.ioseaturtles.org

.

Memorandum of Understanding on ASEAN Sea Turtle Conservation and Protection The objectives of this MOU, initiated by the Association of South East Asian Nations (ASEAN), are to promote the protection, c onservation, replenishing, and rec overy of sea turtles and their

habitats based on the best available scientific ev idence, taking into account the environmental,

socio-economic and cultural characteristics of the Pa rties. It currently has 10 signatory states in the South East Asian Region: Brunei Darussalam, Cambodia, Indonesia, Laos, Malaysia, Myanmar, Philippines, Singapore, Thailand, and Viet nam. Additional information is available at http://www.aseansec.org/6185.htm

.

Inter-American Convention for the Protecti on and Conservation of Sea Turtles (IAC) This Convention is one of only a handful of inte rnational treaties dedica ted exclusively to sea turtles, setting standards for the conservation of these endangered animals and their habitats with a large emphasis on bycatch reduction. It is the only binding multi-national agreement for sea turtles and is open to all countries in North, Central, and South America, and the Caribbean. It currently has 12 Contracting Par ties, with the United States being a signatory in 1999.

Additional information is available at http://www.iacseaturtle.org

.114 Convention for the Protection of the Marine En vironment and Coastal Area of the South-East Pacific (Lima Convention) This Convention’s signatories in clude all countries along the Paci fic Rim of South America from Panama to Chile. Among other resource management components, this Convention established a protocol for the conservation and management of protected marine reso urces. Stemming from

this Convention is the Commision Permanente de l Pacifico Sur (CPPS) that has developed a

Marine Turtle Action Plan for the Southeast Pa cific that outlines a st rategy for protecting and recovering marine turtles in this region.

Convention for the Protection of Natural Re sources and Environment of the South Pacific Region (Noumea Convention) In force since 1990, this Convention includes 19 Parties from throughout the Southwest Pacific Ocean: Australia, Cook Islands, Fe derated States of Micronesia, Fiji, France, Kiribati, Marshall Islands, Nauru, New Zealand, Niue, Palau, Pa pua New Guinea, Samoa, Solomon Islands, Tonga, Tuvalu, United Kingdom, United States, and Vanuatu. This Convention is relevant only for the Exclusive Economic Zones of Party nations, and does not have jurisdiction in international

waters. Relevant to marine turtles are the asso ciated Protocol for the Prevention of Pollution of the South Pacific Region by Du mping (reduction of marine debr is), and the Action Plan for Managing the Natural Resources and Environment of the South Pacific Region, which occurs under the auspices of the South Pacific Regiona l Environment Programme (SPREP). Additional

information is available at http://sedac.ciesin.org/e ntri/texts/natural.resour ces.south.pacific.1986.htm l

.

5.2.1.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear Incidental capture in artisanal a nd commercial fisheries is a significant threat to the survival of loggerheads in the North Pacific. Sea turtles may be caught in pelagic and demersal longlines, drift and set gillnets, bottom a nd mid-water trawling, fishing dredges, pound nets and weirs, haul and purse seines, pots and traps, and hook and line gear.

Based on turtle sightings and capture rates reported in a survey of fisheries research and training vessels and extrapolated to tota l longline fleet effort by the Ja panese fleet in 1978, Nishemura and Nakahigashi (1990) estimated that 21,200 turtles, including gr eens, leatherbacks, loggerheads, olive ridleys, and hawksbills, were captured annually by Japane se tuna longliners in the western Pacific and South China Sea, with a reported mortality of approximately 12,300 turtles per year. Using commercial tuna longline logbooks, research vessel data, and questionnaires, Nishemura and Nakahigashi (199 0) estimated that for every 10,000 hooks in the western Pacific and South China Sea, one turtle is captured, with a mortality rate of 42%. Although species-specific informati on on the bycatch is not availabl e, vessels reported that 36% of the sightings of turtles in locations that ov erlap with these commercia l fishing grounds were loggerheads.

Caution should be used in interpreting the results of Nishemura a nd Nakahigashi (1990), including estimates of sea turtle take rate (per number of hooks) and resultant mortality rate, and115 estimates of annual take by the fishery, for the following reasons: (1) the data collected were based on observations by training and research vessels, logbooks, and a questionnaire (i.e., hypothetical), and do not represent actual, substan tiated logged or observed catch of sea turtles by the fishery; (2) the authors assumed that tur tles were distributed homogeneously; and (3) the authors used only one year (1978) to estimate total effort and dist ribution of the Japanese tuna longline fleet. Although the data and analyses provided by Ni shemura and Nakahigashi (1990) are conjectural, longliners fishing in the Pacific ha ve significantly impacted and, with the current level of effort, probably will continue to have significant impacts on sea turtle populations.

Foreign high-seas driftnet fishing in the North Pa cific Ocean for squid, tuna, and billfish ended with a United Nations moratorium in December 1992. Except for observer data collected in 1990-1991, there is virtually no information on the in cidental take of sea turtle species by the driftnet fisheries prior to the moratorium. Th e high-seas squid driftnet fishery in the North Pacific was observed in Japan, Korea, and Taiwan , while the large-mesh fisheries targeting tuna

and billfish were observed in the Japanese fleet (1990-1991) and the Taiwanese fleet (1990). A

combination of observer data and fleet effort sta tistics indicate that 2,986 loggerhead turtles were entangled by the combined fleets of Japan, Korea, and Taiwan during June 1990 through May 1991, when all fleets were monitored. Of th ese incidental entanglements, an estimated 805 loggerheads were killed (27% mo rtality rate) (Wetherall 1997). Data on size composition of the turtles caught in the high-seas driftnet fisheries also were collected by observers. The majority of loggerheads measured by observers were im mature (Wetherall 1997). The cessation of highseas driftnet fishing in 1992 should have reduced the incidental take of marine turtles. However,

nations involved in driftnet fish ing may have shifted to other gear types (e.g., pelagic or demersal longlines, coastal gillnets); this shift in gear types could have resulted in either similar or increased turtle bycatch and associated mortality.

These rough mortality estimates for a single fish ing season provide only a narrow glimpse of the impacts of the driftnet fishery on sea turtles, a nd a full assessment of imp acts would consider the turtle mortality generated by the driftnet fleets over their entire range. Unfortunately, comprehensive data are lacking, but the obser ver data do indicate the possible magnitude of turtle mortality given the best information available. Wetherall et al . (1993) speculate that the actual mortality of sea turtles may have been be tween 2,500 and 9,000 per year, with most of the mortalities being loggerheads taken in the Japa nese and Taiwanese large-mesh fisheries.

While a comprehensive, quantitative assessment of the impacts of the North Pacific driftnet fishery on turtles is impossible without a better understanding of turtle population abundance, genetic identities, exploitation history, and popul ation dynamics, it is likely that the mortality inflicted by the driftnet fisheries in 1990 a nd in prior years was significant (Wetherall et al

. 1993), and the effects may still be evident in sea turtle populations today. The high mortality of juveniles, subadults, and reproductiv e adults in the high-seas driftn et fishery has probably altered the current age structure (espec ially if certain age groups were more vulnerable to driftnet fisheries) and therefore diminished or limited th e reproductive potential of affected sea turtle populations.

Extensive ongoing studies regarding loggerhead mo rtality and bycatch have been administered off the coast of Baja California Sur, Mexico. The location and timing of loggerhead strandings116 documented in 2003-2005 along a 43-kilometer beach (Playa San Lázaro) indicated bycatch in

local small-scale fisheries. In order to corrobor ate this, in 2005, resear chers observed two smallscale fleets operating closest to an area identified as a high-use area for loggerheads. One fleet, based out of Puerto López Mateos, fished primarily for halibut using bottom set gillnets, soaking from 20 to 48 hours. This fleet consisted of up to 75 boats in 2005, and, on a given day, 9 to 40 vessels fished the deep area (32-45 meter dept hs). During a two-month period, 11 loggerheads were observed taken in 73 gillnet day-trips, with eight of those loggerheads landed dead (observed mortality rate – 73%). The other flee t, based in Santa Rosa, fished primarily for demersal sharks using bottom-set longlines baited with tuna or mackerel and left to soak for 20 to 48 hours. In 2005, the fleet numbered only fi ve to six vessels. During the seven daylong bottom-set longline trips observed, 26 loggerhead s were taken, with 24 of them landed dead (observed mortality rate – 92%). Based on these observations, researchers estimated that in 2005 at least 299 loggerheads died in the bottom-set gillnet fishery a nd at least 680 loggerheads died in the bottom-set longline fishery. This a nnual bycatch estimate of approximately 1,000 loggerheads is considered a minimum and is al so supported by shoreline mortality surveys and informal interviews (Peckham et al . 2007).

These results suggest that incident al capture at Baja California Su r is one of the most significant sources of mortality identified for the North Pacific loggerhead population and underscores the importance of reducing bycatch in small-scale fisheries.

In the U.S. Pacific, longline fisheries targeting swordfish and tuna and drift gillnet fisheries targeting swordfish have been id entified as the primary fisherie s of concern for loggerheads. Bycatch of loggerhead turtles in these fisheries has been signifi cantly reduced as a result of timearea closures, required gear modifications, and hard caps imposed on turtle bycatch, with 100% observer coverage in certain areas.

The California/Oregon (CA/OR) drif t gillnet fishery targets swordf ish and thresher shark off the west coast of the United States. The fishery has been observed by NMFS since July 1990, and currently averages 20%. From July 1990 to January 2000, the CA/OR drift gillnet fishery was observed to incidentally capture 17 loggerheads (12 released alive, 1 injured, and 4 killed). Based on a worst-case scenario, NMFS estimated that a maximum of 33 loggerheads in a given year could be incidentally taken by the CA/OR drif t gillnet fleet. Sea turt le mortality rates for hard-shelled species were estimated to be 32% (NMFS 2000).

In 2000, analyses conducted under the mandates of the ESA showed that th e CA/OR drift gillnet fishery was taking excessive numbers of sea turt les, such that the fishery “jeopardized the continued existence of” loggerheads and leatherb acks. In this case, the consulting agency (NMFS) was required to provide a reasonable and prudent altern ative to the action (i.e., the fishery). In order to reduce th e likelihood of interactions with loggerhead sea turtles, NMFS has regulations in place to close areas to drift gillnet fishing off sout hern California during forecasted or occurring El Niño events from June 1 th rough August 31, when loggerheads are likely to

move into the area from the Pacific coast of Ba ja California following a preferred prey species, pelagic red crabs.117 Prior to 2000, the Hawaii-based longline fisher y targeted highly migratory species north of Hawaii using gear largely used by fleets around the world. From 1994-1999, the fishery was estimated to take between 369 and 501 loggerheads per year, with between 64 and 88 mortalities per year (NMFS 2000). Currently, the Hawaii- based shallow longline fishery targeting swordfish is strictly regulated such that an annual take of 17 loggerheads is authorized for the fishery, beginning in 2004, when the fishery was re -opened after being closed for several years. In 2004 and 2005, the fishing year was completed w ithout reaching the turtle take levels (1 and 10 loggerheads were captured, respectively, with fleets operating with 100% observer coverage). However, in 2006, 17 loggerheads were taken, forc ing the fishery to be shut down early. In 2007, 15 loggerheads were taken by the fishery. Most loggerheads were released alive (NMFSPacific Islands Regional Office, Observ er Database Public Website, 2008).

Recent investigations off the coast of Japan, part icularly focused off the main islands of Honshu, Shikoku, and Kyushu, have revealed a major th reat to the more mature stage classes of loggerheads (approximately 70-80 cm SCL) due to pound net fisheries set offshore of the nesting beaches and in the coastal foraging areas. While pound nets constitute the third largest fishery in

terms of metric tons of fish ca ught in Japan, they account for th e majority of loggerhead bycatch by Japanese fisheries. Open-type pound nets studied in an area off Shikoku were shown to take loggerheads as the most prevalent sea turtle sp ecies caught but had lowe r mortality rates (less than 15%), primarily because turtles could re ach the surface to breath e. Middle layer and bottom-type pound nets in particular have high rate s of mortality (nearly 100%), because the nets are submerged and sea turtles are unable to reach the surface. Estimates of loggerhead mortality in one area studied between April 2006 a nd September 2007 were on the order of 100 individuals. While the fishing industry has an in terest in changing its g ear to open-type, it is very expensive, and the support from the Japanese government is limited (T. Ishihara, Sea Turtle Association of Japan, personal communication, 2007). Nonetheless, the BRT recognizes that coastal pound net fisheries off Japan may pose a si gnificant threat to the North Pacific population of loggerheads.

Quantifying the magnitude of the th reat of fisheries in the North Pacific Ocean on loggerhead sea turtles is very difficult given the low level of obs erver coverage or inves tigations into bycatch conducted by countries that have large fishing fleets. Efforts have been made to quantify the effect of pelagic longline fish ing on loggerheads, and annual estim ates of bycatch were on the order of over 10,000 sea turtles, with as many as 2,600 indivi dual loggerheads killed annually through immediate or delayed mortality as a re sult of interacting w ith the gear (Lewison et al

. 2004).

Other Manmade and Natural Impacts Similar to other areas of the world (see Section 5.1. 5.), climate change and sea level rise have the potential to impact loggerheads in the Nort h Pacific Ocean. For example, Matsuzawa et al. (2002) found heat-related mortalit y of pre-emergent hatchlings in Minabe Senri Beach and concluded that this population is vulnerable to even small temperature increases resulting from global warming because sand temperatures al ready exceed the optimal thermal range for incubation. Recently, Chaloupka et al . (2008) used generalized additive regression modeling and autoregressive-prewhitened cross-correlatio n analysis to consider whether changes in regional ocean temperatures affect long-te rm nesting population dynamics for Pacific118 loggerheads from primary nesting assemblages in Japan and Australia. Researchers chose four nesting sites with a generally long time series to model, two in Japan (Kamouda rookery, declining population, and Yakushima rookery, gene rally increasing in the last 20 years), and two nesting sites in Australia (Woongarra rookery, generally declining through early 1990s and beginning to recover, and Wreck Island rookery, which is generall y declining). Analysis of 51 years of mean annual sea surface time series around two core foraging areas off Japan and eastern Australia, showed a genera l warming of the oceans in these regions. In general, nesting abundance for all four rookeries was inversely rela ted to sea surface temperatures; that is, higher sea surface temperatures during the previous year in the core foraging area resulted in lower summer season nesting at all rookeries. Given th at cooler ocean temperatures are generally associated with increased productivity and that fe male sea turtles generally require at least one year to acquire sufficient fat stor es for vitellogenesis to occur in the foraging grounds, as well as the necessary energy required for migration, any lag in productivity due to warmer temperatures has physiological basis. Over the long-term, warm ing ocean temperatures could therefore lead to lower productivity and prey abundance, and thus reduced nesting and recruitment by Pacific loggerheads (Chaloupka et al . 2008).

As described in Section 5.1.5., other anthropogenic impacts, such as boat strikes and ingestion or entanglement in marine debris, as well as entrainment in coastal power plants, also apply to loggerheads in the North Pacific.

Natural environmental events, such as cyclones and hurricanes, may affe ct loggerheads in the North Pacific Ocean. Typhoons have also been shown to cause severe beach erosion and negatively affect hatching success at many loggerh ead nesting beaches in Japan, especially in areas already prone to erosion. For example, during the 2004 season, the Japanese archipelago suffered a record number of typhoons and many ne sts were drowned or washed out. Extreme sand temperatures at nesting beache s also create highly skewed female sex ratios of hatchlings or threaten the health of hatchlings. Without human intervention to protect clutches against some of these natural threats, many of these nests woul d be lost (Matsuzawa 2006).

5.2.2. South Pacific Ocean DPS

The following information on threats pertains to all loggerhead turtles th at may be found within

the boundaries of the South Pacific Ocean DPS. Because foraging loggerheads from this DPS

also are found within the boundaries of the Southeast-Indo Pacific Ocean DPS (specifically in the Gulf of Carpentaria west of 142.5oE), the narrative for that region should also be consulted.

5.2.2.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone As described in Section 5.1.1., destruction and modification of loggerhead nesting habitat in the South Pacific result from coastal development and construction, placement of erosion control structures and other barriers to nesting, beachfr ont lighting, vehicular traffic, beach erosion, beach pollution, removal of native vegetation, and planting of non-native vegetation (NMFS and FWS 1998, Limpus 2009).119 Removal or destruction of native dune vegetation, which enhances beach stability and acts as an integral buffer zone between land and sea, results in erosion of nesting habitat. Preliminary studies on nesting beaches in New Caledonia incl ude local oral histories that attribute the decrease in loggerhead nesting to the remova l of vegetation for construction purposes and subsequent beach erosion (Limpus et al . 2006).

Beach armoring presents a barrier to nesting in th e South Pacific. On the primary nesting beach in New Caledonia, a rock wall was constructed to prevent coastal erosion, and sea turtle nesting attempts have been unsuccessful. Local reside nts are seeking authori zation to extend the wall further down the beach (Limpus et al . 2006).

Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, channel dredging, sa nd extraction, marine pollution, and climate change. These threats also occur in the South Pacific Ocean. Climate change, for instance, may result in future trophic change s, thus impacting loggerhead prey abundance and/or distribution.

5.2.2.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

Legislation in Australia outlaws the harvesting of loggerheads by indigenous peoples (Limpus et
al . 2006). Despite national laws, in many areas th e poaching of eggs and hunting of adults and juveniles is still a problem, and Limpus (2009) suggests that th e harvest rate of loggerheads by indigenous hunters, both within Australia and in neighboring countries, is on the order of 40 turtles per year. Preliminary studies suggest that local harvesting in New Caledonia constitutes about 5% of the nesting population (Limpus et al . 2006). Loggerheads also are consumed after being captured incidentally in high-seas fisherie s of the southeastern P acific (Alfaro-Shigueto et
al . 2006), and occasionally may be the product of illegal trade throughout the region.

5.2.2.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the South Pacific. While the prevalence of fibropa pillomatosis in most loggerhead populations is thought to be small, an exception is in Moreton Bay, Australia, where 4.4% of the 320 loggerheads captured exhi bited the disease during 1990-1992 (Limpus et al

. 1994). A subsequent study also found a high pr evalence of fibropapillo matosis in the area (Quackenbush et al . 2000).

Predation on nests and hatchlings by terrestria l vertebrates is a major problem at loggerhead

rookeries in the South Pacific. At mainland r ookeries in eastern Australia, for example, the introduced fox ( Vulpes vulpes ) has been the most significan t predator on loggerhead eggs (Limpus 1985, 2009). Although this has been mini mized in recent years (to <5%; Limpus 2009), researchers believe the earlier egg loss will greatl y impact recruitment to this nesting population in the early 21 st century (Limpus and Reimer 1994). Predation on hatchlings by crabs and

diurnal birds is also a threat (Limpus 2009). In New Ca ledonia, feral dogs pose a predation threat to nesting loggerheads, and thus far no management has been implemented (Limpus et al

. 2006).120 5.2.2.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the South Pacific Ocean. The following additional instruments also apply to loggerheads found in this area.

Indian Ocean – South-East Asian Marine Tu rtle Memorandum of Understanding (IOSEA) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Memorandum of Understanding on ASEAN Sea Turtle Conservation and Protection For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Inter-American Convention for the Protecti on and Conservation of Sea Turtles (IAC) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Convention for the Protection of the Marine En vironment and Coastal Area of the South-East Pacific (Lima Convention) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Convention for the Protection of Natural Re sources and Environment of the South Pacific Region (Noumea Convention) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

5.2.2.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear Incidental capture in artisanal a nd commercial fisheries is a significant threat to the survival of loggerheads throughout the South Paci fic. The primary gear types i nvolved in these interactions include longlines, driftnets, set nets, and trawl fisheries. These are employed by both artisanal and industrial fleets, and target a wide variet y of species including t unas, sharks, sardines, swordfish, and mahi mahi (see below).

In the southwestern Pacific, bottom trawling gear has been a contributing f actor to the decline in the eastern Australian logge rhead population (Limpus and Reimer 1994). The northern Australian prawn fishery (NPF) is made up of both a banana prawn fish ery and a tiger prawn fishery, and extends from Cape York, Qu eensland (142oE) to Cape Londonberry, Western Australia (127oE). The fishery is one of the mo st valuable in all of Australia and in 2000 comprised 121 vessels fishing approx imately 16,000 fishing days (Robins et al . 2002a). In 2000, the use of turtle excluder devices (TEDs) in th e NPF was made mandatory, due in part to several factors: (1) objectives of the Dr aft Australian Recovery Plan for Marine Turtles, (2) requirement

of the Australian Environment Protection and Biodiversity Conservation Act for Commonwealth fisheries to become ecologically sustainable, and (3) the 1996 U.S. import embargo on wildcaught prawns taken in a fishery without adequate turt le bycatch management practices (Robins

et al . 2002a). Data primarily were collected by volunt eer fishers who were tr ained extensively in the collection of scientific data on sea turtles caught as bycatch in their fishery. Prior to the use of TEDs in this fishery, the NPF annually took between 5,000 and 6,000 sea turtles as bycatch,121 with a mortality rate of an estimated 40% due to drowning, injuries, or being returned to the water comatose (Poiner and Harris 1996). Since the mandatory use of TEDs has been in effect, the annual bycatch of sea turtle s in the NPF has dropped to less than 200 sea turtles per year, with a mortality rate of approximately 22% (bas ed on recent years). This lower mortality rate also may be based on better sea turtle handli ng techniques adopted by the fleet. In general, loggerheads were the third most common sea turtle taken in this fishery.

Loggerheads are also taken by longline fisheries operating out of Australia (Limpus 2009). For example, Robins et al . (2002b) estimate that approximately 400 turtles are killed annually in Australian pelagic longline fishery operations. Of this annual estimat e, leatherbacks accounted for over 60% of this total, while unidentified hardshelled turtles accounted for the remaining species. Therefore, the eff ect of this longline fisher y on loggerheads is unknown.

Loggerheads also have been the most common turt le species captured in shark control programs in Australia (Kidston et al . 1992, Limpus 2009). From 1998-2002, a total of 232 loggerheads were captured with 195 taken on drum lines and 37 taken in nets, bot h with a low level of direct mortality (Limpus 2009 ).

In the southeastern Pacific, significant bycatch ha s been reported in artisanal gillnet and longline shark and mahi mahi fisheries operating out of Peru (Kelez et al . 2003, Alfaro-Shigueto et al

. 2006) and to a lesser extent, Chile (Donoso and Dutton 2006). The fish ing industry in Peru is the second largest economic activity in the country, and, over the past few years, the longline fishery has rapidly increased. Currently, nearly 600 long line vessels fish in the winter and over 1,300 vessels fish in the summer. During an obser ver program in 2003/2004, 588 sets were observed during 60 trips, and 154 sea turtles were taken as bycatch. Loggerheads were the species most often caught (73.4%). Of the loggerheads taken, 68% were entangled an d 32% were hooked. Of the two fisheries, sea turtle bycatch was highest during the mahi mahi season, with 0.597 turtles/1,000 hooks, while the shark fishery ca ught 0.356 turtles/1,000 hooks (Alfaro-Shigueto et
al . 2008b). A separate study by Kelez et al . (2003) reported that approximately 30% of all turtles bycaught in Peru were loggerheads. In many cases, loggerheads are kept on board for human consumption; therefore, th e mortality rate in this artisa nal longline fishery is likely high because sea turtles are retained for future consumption or sale.

Data on loggerhead bycatch in Chile are limited to the industrial swor dfish fleet. Since 1990, fleet size has ranged from 7 to 23 vessels with a mean of approximately 14 vessels per year. These vessels fish up to and over 1,000 nautical miles along the Chilean coast with mechanized sets numbering approximately 1,200 hooks (M. D onoso, ONG Pacifico Laud - Chile, personal communication, 2007). Loggerhead bycatch is presen t in Chilean fleets; however, the catch rate is substantially lower than that reported fo r Peru (P. Dutton, NMFS, and M. Donoso, ONG Pacifico Laud - Chile, unpublished data).

Other Manmade and Natural Impacts

Other threats such as debris in gestion, boat strikes, and port dr edging also impact loggerheads in the South Pacific, although these th reats have been minimized in re cent years due to a variety of legislative actions (Limpus 2009). Loggerhead mort ality resulting from dredging of channels in Queensland is a persistent, albeit minor problem. From 1999-2002, the average annual reported122 mortality was 1.7 turtles per year (range = 1- 3) from port dredging operations (Limpus 2009). Climate change and sea level rise have the poten tial to impact loggerhead s in the South Pacific Ocean, yet the impact of these th reats has not been quantified.

Natural environmental events, such as cyclones or hurricanes, may affect loggerheads in the South Pacific Ocean. These types of events may disrupt loggerhead nesting activity, albeit on a temporary scale. As described in mo re detail in Section 5.2.1.5., Chaloupka et al . (2008) demonstrated that nesting abundance of loggerhea ds in Australia was inversely related to sea surface temperatures, and suggested that a long te rm warming trend in the South Pacific may be adversely impacting the recovery potential of this population.

5.2.3. North Indian Ocean DPS

5.2.3.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone As described in Section 5.1.1., destruction and modification of loggerhead nesting habitat in the North Indian Ocean result from coastal development and construction, beachfront lighting, vehicular and pedestrian traffic, beach pollution, removal of native vegetation, and planting of non-native vegetation (NMFS and FWS 1998).

The primary loggerhead nesting beaches of this DPS are at Masirah Isla nd, Oman, and are still relatively undeveloped but now facing increasing development pressures. Newly paved roads closely paralleling most of the Masirah Isla nd coast are bringing newl y constructed highway lights (E. Possardt, FWS, personal observation, 2008) and greater access to nesting beaches by the public. Light pollution from the military installa tion at Masirah Island also is evident at the most densely nested northern end of the island an d is a likely cause of hatchling misorientation and nesting female disturbance (E. Possardt, FWS, personal observation, 2008). Beach driving occurs on most of the major beaches outside the military installation. This vehicular traffic

creates ruts that obstruct hatchl ing movements (Mann 1977, Hosier et al . 1981, Cox et al . 1994, Baldwin 1992), tramples nests, and destroys vegetation and dune formation processes, which exacerbates light pollution effects. Free rang ing camels, sheep, and goats overgraze beach vegetation, which impedes natural dune format ion (E. Possardt, FWS, personal observation, 2008). Development of a new hotel on a major l oggerhead nesting beach at Masirah Island is near completion and, although not yet approved, ther e are plans for a major re sort at an important loggerhead nesting beach on one of the Halaniya t Islands. Armoring structures common to many developed beaches throughout the world are not yet evident on the major loggerhead nesting

beaches of this DPS.

Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, channel dredging, sa nd extraction, marine pollution, and climate change. These threats also o ccur in the North Indian Ocean.123 Fishing methods not only incidentally capture loggerheads (Section 5.1.5. ), but also deplete invertebrate and fish populations and thus alter ecosystem dynam ics. In many cases loggerhead foraging areas coincide with fishi ng zones. There has been an apparent growth in artisanal and

commercial fisheries in waters surrounding Masirah Isla nd (Baldwin 1992). Climate change also may result in future trophic changes, thus impacting loggerhead prey abundance and/or

distribution.

5.2.3.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

The use of loggerhead meat for food in Oman is not legal or popular. However, routine egg collection on Masirah Island does occur (Baldw in 1992). The extent of egg collection as estimated by Masirah rangers and local reside nts is approximately 2,000 clutches per year (<10%).

5.2.3.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the North Indian Ocean. Natural egg predation on Oman loggerhead nesting beaches undoubtedly occurs, but is not well documented or believed to be significant. Predation on hatchlings by Arabian red fox ( Vulpes vulpes arabica ), ghost crabs ( Ocypode

saratan ), night herons ( Nycticorax nycticorax ), and gulls ( Larus spp.) likely occurs. While quantitative data do not exist to evaluate thes e impacts on the North Indian Ocean loggerhead population, they are not li kely to be significant.

5.2.3.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the No rth Indian Ocean. Th ere are two additional instruments that apply to loggerheads found in this area.

Indian Ocean – South-East Asian Marine Tu rtle Memorandum of Understanding (IOSEA) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Nairobi Convention for the Protection, M anagement and Developmen t of the Marine and Coastal Environment of th e Eastern African Region

The Nairobi Convention was si gned in 1985 and came into force in 1996. The Convention covers 10 States, including five island States in the Western Indian Ocean. The Contracting Parties are Comoros, France (La Reunion), Kenya, Madagascar, Mauritius, Mozambique, Seychelles, Somalia, Tanzania and the Republic of South Africa. This instrument “provides a mechanism for regional cooperati on, coordination and collaborative actions , and enables the Contracting Parties to harness resources and expe rtise from a wide range of stakeholders and interest groups towards solving interlinked problems of the coastal and marine environment.”

Additional information is available at http://www.unep.org/NairobiConvention.124 5.2.3.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear The magnitude of the threat of incidental capt ure of sea turtles in ar tisanal and commercial

fisheries in the North Indian Ocean is difficult to assess. A bycatch survey administered off the coast of Sri Lanka between September 1999 an d November 2000 reported 5,241 total turtle entanglements, of which 1,310 were loggerheads, between Kalpitiya and Kirinda (Kapurusinghe and Saman 2001, Kapurusinghe and Cooray 2002). Sea turtle bycatch has been reported in driftnet and set gillnets, longlines, trawls, and hook and line gear (Kapurusinghe and Saman 2001, Kapurusinghe and Cooray 2002, Lewison et al . 2004).

Quantifying the magnitude of the th reat of fisheries on loggerheads in the North Indian Ocean is difficult given the low level of observer coverage or investigations into bycatch conducted by countries that have large fishing fleets. Efforts have been made to quantify the effects of pelagic longline fishing on loggerh eads globally (Lewison et al . 2004). While ther e were no turtle bycatch data available from the North Indian Ocean to use in their assessment, extrapolations that considered bycatch data for the Pacific and Atlantic basins gave a conservative estimate of 6,000 loggerheads captured in the Indian Ocean in the year 2000. Interviews with rangers at Masirah Island reveal that shark gillnets captu re many loggerheads off nesting beaches during the nesting season. As many as 60 boats are involv ed in this fishery with up to 6 km of gillnets being fished daily from June through October along the Masirah Island coast. Rangers reported one example of 17 loggerheads in one net (E. Possardt, FWS, personal communication, 2008).

Other Manmade and Natural Impacts As described in Section 5.1.5., other anthropogenic impacts, such as boat strikes and ingestion or entanglement in marine debris, as well as entrainment in coastal power plants, likely apply to loggerheads in the North Indian Ocean. Similar to other areas of the world (see Section 5.1.5.), climate change and sea level rise have the potenti al to impact loggerheads in the North Indian

Ocean. This includes beach erosion and loss from rising sea levels, skewed hatchling sex ratios from rising beach incubation temperatures, and abrupt disruption of ocean currents used for natural dispersal during the complex life cycle. Climate change impacts could have profound long term impacts on nesting populations in the North Indian Ocean, but it is not possible to quantify the potential impacts at this point in time.

Natural environmental events, such as cyclones, tsunamis, and hurricanes, affect loggerheads in the North Indian Ocean. For example, during the 2007 season, Oman suffered a rare typhoon. In general, however, severe storm events are episodic and, although they may affect loggerhead hatchling production, the results are generally localized and they rarely result in whole-scale losses over multiple nesting seasons.

5.2.4. Southeast Indo-Pacific Ocean DPS

The following information on threats pertains to all loggerhead turtles th at may be found within

the boundaries of the Southeast Indo-Pacific Ocea n DPS. Because foraging loggerheads from this DPS also are found within the boundaries of the South Pacific Ocean DPS, the narrative for that region should also be consulted.125 5.2.4.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone The primary loggerhead nesting beaches for this DPS occur in Australia on Dirk Hartog Island and Murion Islands (Baldwin et al . 2003), which are undeveloped. Dirk Hartog Island is soon to become part of the National Park System.

Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, channel dredging, sa nd extraction, marine pollution, and climate change. These threats also occur in the Southe ast Indo-Pacific Ocean. Fishing methods not only incidentally capture loggerheads (Section 5.1.5. ), but also deplete invertebrate and fish populations and thus alter ecosystem dynamics . In many cases, loggerhead foraging areas coincide with fishing zones. Climate change al so may result in future trophic changes, thus impacting loggerhead prey abundance and/or distribution.

5.2.4.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

Legislation in Australia outlaws the harvesting of loggerheads by indigenous peoples (Limpus et
al . 2006). Dirk Hartog Island and Mu rion Islands are largely uninha bited, and poaching of eggs and turtles is likely negligible.

5.2.4.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the Southeast Indo-Pacifi c Ocean. On the North West Cape and the beaches of the Ningaloo coast of mainland Austra lia, a long established feral European red fox ( Vulpes vulpes ) population preyed heavily on eggs and is thought to be respons ible for the lower numbers of nesting turtles on the mainland beaches (Baldwin et al . 2003). The fox populations

have been eradicated on Dirk Hart og Island and Murion Islands (Baldwin et al . 2003).

5.2.4.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in th e Southeast Indo-Pacific Ocean. The following additional instrument also applies to loggerheads found in this area.

Indian Ocean – South-East Asian Marine Tu rtle Memorandum of Understanding (IOSEA) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

5.2.4.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear The extent of the threat of incidental capture of sea turtles in artisanal and commercial fisheries

in the Southeast Indo-Pacific Ocean is unknown. Sea turtles are caught in pelagic and demersal126 longlines, gillnets, trawls, seines, and pots and traps (Environment Australia 2003). There is evidence of significant historic byc atch from prawn fisheries, wh ich may have depleted nesting populations long before nesting surveys were initiated in the 1990s (Baldwin et al . 2003).

Quantifying the magnitude of the threat of fish eries on loggerheads in the Southeast Indo-Pacific Ocean is very difficult given the low level of obs erver coverage or investigations into bycatch conducted by countries that have large fishing fleets. Efforts have been made to quantify the effects of pelagic longline fishi ng on loggerheads globally (Lewison et al . 2004). While there were no turtle bycatch data available from th e Southeast Indo-Pacific Ocean to use in their assessment, extrapolations that considered bycatch data for the Pacific and Atlantic basins gave a conservative estimate of 6,000 loggerheads captured in the Indian Ocean in the year 2000. Loggerheads are known to be taken by Japanese longline fisheries oper ating off of Western Australia (Limpus 2009). The effect of the longl ine fishery on loggerheads in the Indian Ocean is largely unknown (Lewison et al . 2004).

The northern Australian prawn fishery (NPF) is made up of both a banana prawn fishery and a tiger prawn fishery, and exte nds from Cape York, Queensland (142oE) to Cape Londonberry, Western Australia (127oE). The fishery is one of the most valuable in all of Australia and in 2000 comprised 121 vessels fishing approxi mately 16,000 fishing days (Robins et al . 2002a). In 2000, the use of turtle excluder devices in the N PF was made mandatory, due in part to several factors: (1) objectives of the Dr aft Australian Recovery Plan for Marine Turtles, (2) requirement

of the Australian Environment Protection and Biodiversity Conservation Act for Commonwealth fisheries to become ecologically sustainable, and (3) the 1996 U.S. import embargo on wildcaught prawns taken in a fishery without adequate turt le bycatch management practices (Robins

et al . 2002a). Data primarily were collected by volunt eer fishers who were tr ained extensively in the collection of scientific data on sea turtles caught as bycatch in their fishery. Prior to the use of TEDs in this fishery, the NPF annually took between 5,000 and 6,000 sea turtles as bycatch, with a mortality rate of an estimated 40%, due to drowning, injuries, or being returned to the water comatose (Poiner and Harris 1996). Since the mandatory use of TEDs has been in effect, the annual bycatch of sea turtle s in the NPF has dropped to less than 200 sea turtles per year, with a mortality rate of approximately 22% (bas ed on recent years). This lower mortality rate also may be based on better sea turtle handli ng techniques adopted by the fleet. In general, loggerheads were the third most common sea turtle taken in this fishery.

Loggerheads also have been the most common turt le species captured in shark control programs in Pacific Australia (Kidston et al . 1992, Limpus 2009); however , the Western Australian demersal longline fishery for sharks has no reco rded interaction with loggerheads. From 1998- 2002, a total of 232 loggerheads were captured w ith 195 taken on drum lines and 37 taken in nets, both with a low level of direct mortality (Limpus 2009).

Other Manmade and Natural Impacts As described in Section 5.1.5., other anthropogenic impacts, such as boat strikes and ingestion or entanglement in marine debris, likely apply to lo ggerheads in the Southeast Indo-Pacific Ocean. Similar to other areas of the world (see Section 5.1. 5.), climate change and sea level rise have the potential to impact loggerheads in the South east Indo-Pacific Ocean. This includes beach erosion and loss from rising sea levels, skewed hatchling sex ratios from rising beach incubation127 temperatures, and abrupt disruption of ocean cu rrents used for natural dispersal during the complex life cycle. Climate change impacts could have profound long term impacts on nesting populations in the Southeast Indo-Pacific Ocean, but it is not possible to quantify the potential impacts at this point in time.

Natural environmental events, such as cyclones and hurricanes, may affe ct loggerheads in the Southeast Indo-Pacific Ocean. In general, howev er, severe storm events are episodic and, although they may affect loggerhea d hatchling production, the result s are generally localized and they rarely result in whole-scale losses over multiple nesting seasons.

5.2.5 Southwest Indian Ocean DPS

5.2.5.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone All nesting beaches within South Africa are within protected areas (Baldwin et al . 2003). In Mozambique, nesting beaches in the Maputo Spec ial Reserve (approximately 60 km of nesting beach) and in the Paradise Islands ar e within protected areas (Baldwin et al . 2003, Costa et al

. 2007). There are no protected areas for loggerheads in Madagascar (Baldwin et al . 2003).

Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, channel dredging, sa nd extraction, marine pollution, and climate change. These threats also occur in the Sout hwest Indian Ocean. Fishing methods not only incidentally capture loggerheads (Section 5.1.5. ), but also deplete invertebrate and fish populations and thus alter ecosystem dynamics . In many cases, loggerhead foraging areas coincide with fishing zones. Climate change al so may result in future trophic changes, thus impacting loggerhead prey abundance and/or distribution.

5.2.5.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

In the Southwest Indian Ocean, on the east coast of Africa, subsistence hunting by local people is a continued threat to loggerheads (Baldwin et al . 2003). Illegal hunting of marine turtles and egg harvesting remains a threat in Mozambique as well (Louro et al . 2006).

5.2.5.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the Southwest Indi an Ocean. Side striped jackals ( Canis adustus) and honey badgers ( Melivora capensis ) are known to depred ate nests (Baldwin et al . 2003).128 5.2.5.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the Southwest Indian Ocean. The following additional instrument also applies to loggerheads found in this area.

Indian Ocean – South-East Asian Marine Tu rtle Memorandum of Understanding (IOSEA) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Nairobi Convention for the Protection, M anagement and Developmen t of the Marine and Coastal Environment of th e Eastern African Region For information about this instrument, see sec tion 5.2.3.4. for the North Indian Ocean DPS.

5.2.5.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear The full extent of the threat of incidental capt ure of sea turtles in ar tisanal and commercial

fisheries in the Southwest Indian Ocean is unkno wn. Sea turtles are caught in demersal and pelagic longlines, trawls, gillnets, and seines (Petersen 2005; Louro et al . 2006; Petersen et al. 2006, 2009; Costa et al . 2007; Fennessy and Isaksen 2007). Th ere is evidence of significant historic bycatch from prawn fisheries, which ma y have depleted nesting populations long before nesting surveys were initia ted in the 1990s (Baldwin et al . 2003).

Quantifying the magnitude of the threat of fi sheries on loggerheads in the Southwest Indian Ocean is very difficult given the low level of obs erver coverage or investigations into bycatch conducted by countries that have large fishing fleets. Efforts have been made to quantify the effects of pelagic longline fishi ng on loggerheads globally (Lewison et al . 2004). While there were no turtle bycatch data available from the Southwest Indian Ocean to use in their assessment, extrapolations that considered bycatch data for the Pacific and Atlantic basins gave a conservative estimate of 6,000 loggerheads captured in the Indian Ocean in the year 2000. The effect of the longline fisher y on loggerheads in the Indian Ocean is largely unknown (Lewison et
al . 2004).

Other Manmade and Natural Impacts As described in Section 5.1.5., other anthropogenic impacts, such as boat strikes and ingestion or entanglement in marine debris, likely apply to loggerheads in the Southwest Indian Ocean. Similar to other areas of the world (see Section 5.1. 5.), climate change and sea level rise have the potential to impact loggerheads in the Southwest Indian Ocean. This includes beach erosion and loss from rising sea levels, skewed hatchl ing sex ratios from rising beach incubation temperatures, and abrupt disruption of ocean cu rrents used for natural dispersal during the complex life cycle. Climate change impacts could have profound long term impacts on nesting populations in the Southwest Indian Ocean, but it is not possible to quantify the potential impacts at this point in time.

Natural environmental events, such as cyclones, tsunamis and hurricanes, may affect loggerheads in the Southwest Indian Ocean. In general, however, severe storm events are episodic and,129 although they may affect loggerhea d hatchling production, the result s are generally localized and they rarely result in whole-scale losses over multiple nesting seasons.

5.2.6. Northwest Atlantic Ocean DPS

The following information on threats pertains to all loggerhead turtles that may be found in the

Northwest Atlantic. Because it is likely that oceanic juveniles from this DPS also are found in

the Northeast Atlantic Ocean and in the Medite rranean Sea, the narratives for those regions

should also be consulted.

5.2.6.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone As described in Section 5.1.1., destruction and modification of loggerhead nesting habitat in the

Northwest Atlantic results from coastal development and cons truction, placement of erosion

control structures and other barriers to nesti ng, placement of nearshore shoreline stabilization structures, beachfront lighting, vehicular and pe destrian traffic, beach erosion, beach sand placement, removal of native vegetation, and pl anting of non-native vegetation (NMFS and FWS 2008).

In the southeastern U.S., numerous erosion control structures that create barriers to nesting have been constructed. The proportion of coastline th at is armored is approx imately 18% (239 km) in Florida (Clark 1992, Schroeder and Mosier 2000, Witherington et al . 2006), 9% (14 km) in Georgia (M. Dodd, GDNR, personal communication, 2000), 12% (29 km) in South Carolina (D. Griffin, SCDNR, personal comm unication, 2009), and 3% (9 km) in North Carolina (M. Godfrey, North Carolina Wild life Resources Commission, 2009). These assessments of armoring extent do not include structures that are a barrier to sea turtle ne sting but that do not fit the definition of armoring, such as dune cr ossovers, cabanas, sand fences, and recreational equipment.

In the Northwest Atlantic, jetties have been plac ed at many ocean inlets to keep transported sand from closing the inlet channel. Witherington et al . (2005) found a significant negative relationship between loggerhead nesting density a nd distance from the nearest of 17 ocean inlets

on the Atlantic coast of Florida. The effect of inlets in lowering nesting density was observed both updrift and downdrift of the in lets, leading researchers to propose that beach instability

from both erosion and accretion ma y discourage loggerhead nesting.

Numerous beaches in the southeastern U.S. ar e eroding due to both natural (e.g., storms, sea level changes, waves, shorel ine geology) and anthropogenic (e .g., construction of armoring structures, groins, and jetties; coastal development; inlet dredging) factors. Such shoreline erosion leads to a loss of nesting habitat for se a turtles. As a result, beach nourishment is a frequent activity, and many beaches are on a periodic nourishment schedule. 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 southeaste rn U.S., nesting success typically declines for the first year or130 two following construction, even though more nesti ng habitat is available for turtles (Trindell et
al . 1998, Ernest and Martin 1999, Herren 1999). Reduced nesting success on constructed

beaches has been attributed to increased sand co mpaction, escarpment formation, and changes in beach profile (Nelson et al . 1987, Crain et al . 1995, Lutcavage et al . 1997, Steinitz et al . 1998, Ernest and Martin 1999, Rumbold et al . 2001).

Stormwater and other water source runoff from b eachfront parking lots, building rooftops, roads, decks, and draining swimming pools adjacent to the beach is frequently discharged directly onto

Northwest Atlantic beaches a nd dunes either by sheet flow , through stormwater collection system outfalls, or through small diameter pipes. These outfalls create localized erosion channels, prevent natural dune esta blishment, and wash out sea turtle nests (Florida Fish and Wildlife Conservation Commission, unpublished data). Contaminants contained in stormwater, such as oils, grease, antifreeze, gasoline, metals, pesticides, chlorine, and nutrients, are also discharged onto the beach.

In the U.S., vehicular driving is allowed on certa in beaches in northeast Florida (Nassau, Duval, St. Johns, and Volusia counties), northwest Fl orida (Walton and Gulf counties), Georgia (Cumberland, Little Cumberland, and Sapelo Islands), North Carolina (Fort Fisher State Recreation Area, Carolina Beach, Freeman Pa rk, Onslow Beach, Emerald Isle, Indian Beach/Salter Path, Pine Knoll Shores, Atlantic Beach, Cape Lookout National Seashore, Cape Hatteras National Seashore, Nag’s Head, Kill Devil Hills, Town of Duck, and Currituck Banks), Virginia (Chincoteague NWR and Wallops Island), and Texas (the majority of beaches except for a highly developed section of South Padre Island and Padr e Island National Seashore, San Jose Island, Matagorda Island, and Matagorda Peninsula where dr iving is not allowed or is limited to agency personnel, land owners, and/or researchers). Beach driving has been found to reduce the quality of loggerhead nesting habitat in several ways. In the southeastern U.S., vehicle ruts on the beach have been found to pr event 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 has been found to hinder nest construction and hatchling emergence from nests (Mann 1977). Vehi cle lights and vehicle

movement on the beach after dark results in reduced habitat suita bility, which can deter females from nesting and disorient hatchlings. Additionally, vehicle traffic on nesting beaches contributes to erosion, especi ally during high tides or on narrow beaches where driving is concentrated on the high beach and foredune.

As described in Section 5.1.1., coastal deve lopment contributes to habitat degradation by increasing light pollution. Reports of hatchling disorientation even ts in Florida alone describe several hundred nests each year and are likely to involve tens of thousands of hatchlings (Nelson

et al . 2002). However, this number calculated fro m disorientation reports is likely a vast underestimate. Independent of these reports, Witherington et al . (1996) surveyed hatchling orientation at nests located at 23 representativ e beaches in six counties around Florida in 1993 and 1994 and found that, by county, approximately 10 to 30% of nests showed evidence of hatchlings disoriented by lighting. From this survey and from measures of hatchling production (Florida Fish and Wildlife Conservation Commission, unpublished data), the number of hatchlings disoriented by lighting in Florida is calculated in the range of hundreds of thousands

per year.131 Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, channel dredging, sand extraction, oil explora tion and development, marine pollution, and climate change. These threat s also occur in the Northwest Atlantic Ocean. Fishing methods not only incidentally capture loggerheads (Section 5.1.5. ), but also deplete invertebrate and fish populati ons and thus alter ecosystem dynamics. Although anthropogenic disruptions of natural ecological interactions have been difficult to discern, a few studies have been focused on the effects of these disruptions on loggerheads. For instance, Youngkin (2001) analyzed gut contents from hundreds of loggerheads stranded in Georgia over a 20-year period. His findings point to the probability of major eff ects on loggerhead diet from activities such as shrimp trawling and dredging. Lutcavage a nd Musick (1985) found that horseshoe crabs strongly dominated the diet of loggerheads in Chesapeake Bay in 1980-1981. Subsequently, fishermen began to harvest horseshoe crabs, prim arily for use as bait in the eel and whelk pot fisheries, using several gear types. Atlantic coast horseshoe crab landi ngs increased by an order of magnitude (0.5 to 6.0 million pounds) betwee n 1980 and 1997, and in 1998 the Atlantic States Marine Fisheries Commission implemented a horseshoe crab fishery management plan to curtail catches (Atlantic States Marine Fisheries Co mmission 1998). The decline in horseshoe crab availability has apparently cause d a diet shift in juvenile l oggerheads, from predominantly horseshoe crabs in the early to mid-1980s to bl ue crabs in the late 1980s and early 1990s, to mostly finfish in the late 1990s and early 2000s (Seney 2003, Seney and Musick 2007). These data suggest that turtles are foraging in gr eater numbers in or around fishing gears and on discarded bycatch (Seney 2003).

Periodic dredging of sediments from navigational channels is carried out at large ports to provide for the passage of large commercial and military vessels. In addition, sand mining (dredging) for beach renourishment and construction projects occu rs in the Northwest Atlantic along the U.S., Mexico, Central American, Colombia, and Venezuel a coasts. Although directed studies have not been conducted, dredging activities, which occur regularly in the Northw est Atlantic, have the potential to destroy or degrade benthic habitats used by logg erheads. Channelization of inshore

and nearshore habitat and the subs equent disposal of dredged mate rial in the marine environment can destroy or disrupt resting or foraging grounds (including grass beds and coral reefs) and may affect nesting distribution by alte ring physical features in the marine environment (Hopkins and Murphy 1980). Oil exploration and developm ent on live bottom areas may disrupt foraging grounds by smothering benthic organisms with se diments and drilling muds (Coston-Clements and Hoss 1983). The effects of benthic habita t alteration on loggerhead prey abundance and distribution, and the effects of these potential changes on loggerh ead populations, have not been

determined but are of concern. Climate chan ge also may result in trophic changes, thus impacting loggerhead prey abundance and/or distribution.

5.2.6.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

Deliberate hunting of loggerheads for their meat, shells, and eggs is reduced from previous exploitation levels, but still exists. In the Cari bbean, 13 of 29 (45%) countries/territories allow the harvest of loggerheads (NMFS and FWS 2008, see Appendix 3). The loggerhead harvest in the Caribbean is generally restricted to the non-ne sting season with the exception of St. Kitts and Nevis, where turtle harvest is allowed annually from March 1 through September 30, and the132 Turks and Caicos Islands, where turtle harvest is allowed year-round. Mo st countries/territories that allow harvest have regulations that favor th e harvest of large juveniles and adults, the most reproductively valuable members of the populati on. Exceptions include the Cayman Islands, which mandates maximum size limits, and Haiti a nd Trinidad and Tobago, which have no size restrictions. All North, Cent ral, and South American countries in the Northwest Atlantic legislate complete protection of loggerheads from harvest in their territorial waters with the exception of Guyana. Despite national laws, in many countries the poaching of eggs and hunting of adults and juveniles still occurs at varying levels (NM FS and FWS 2008, see Appendix 3).

5.2.6.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the Northwes t Atlantic. Viral diseases ha ve not been documented in freeranging loggerheads, with the possible exception of sea turtle fibropapi llomatosis, which may have a viral etiology (Herbst and Jacobson 1995, George 1997). Althoug h fibropapillomatosis reaches epidemic proportions in some wild gr een turtle populations, the prevalence of this disease in most loggerhead populat ions is thought to be small. An exception is Florida Bay where approximately 9.5% of the loggerheads captured exhibit fibropapilloma-like external lesions (B. Schroeder, NMFS, pe rsonal communication, 2006). Mortality levels and populationlevel effects associated with the disease are st ill unknown. Heavy infestations of endoparasites may cause or contribute to debilitation or mortality in sea turtle s. Trematode eggs and adult trematodes were recorded in a variety of tissues in cluding the spinal cord and brain of debilitated loggerheads during an epizootic in South Florid a, USA, during late 2000 and early 2001. These endoparasites were implicated as a po ssible cause of the epizootic (Jacobson et al . 2006). Although many health problems have been descri bed in wild populations through the necropsy of stranded turtles, the significance of diseases on the ecology of wild l oggerhead populations is not known (Herbst and Jacobson 1995).

Predation of eggs and hatchli ngs by native and introduced species occurs on almost all nesting beaches throughout the Northwest A tlantic. The most common pred ators at the primary nesting beaches in the southeastern United States are ghost crabs ( Ocypode quadrata ), raccoons ( Procyon lotor ), feral hogs ( Sus scrofa ), foxes ( Urocyon cinereoargenteus and Vulpes vulpes ), coyotes ( Canis latrans ), armadillos ( Dasypus novemcinctus ), and red fire ants ( Solenopsis

invicta ) (Stancyk 1982, Dodd 1988). In the absence of well managed nest protection programs, predators may take significant numbers of eggs; however, nest protection programs are in place at most of the major nesting beaches in the Northwest Atlantic.

Non-native vegetation has invaded many coasta l areas and often outcompetes native plant species. Exotic vegetation may form impenetrable root mats that can invade and desiccate eggs, as well as trap hatchlings. The Australian pine ( Casuarina equisetifolia ) is particularly harmful

to sea turtles. Dense stands have taken ove r many coastal areas th roughout central and south Florida. Australian pines cause excessive shading of the beach that would not otherwise occur. Studies in Florida suggest that nests laid in shaded areas are subject ed to lower incubation temperatures, which may alter the natural hatchl ing sex ratio (Marcus an d Maley 1987, Schmelz and Mezich 1988, Hanson et al . 1998). Fallen Australian pines limit access to suitable nest sites and can entrap nesting females (Austin 1978, R eardon and Mansfield 1997). The shallow root133 network of these pines can interf ere with nest construction (Sch melz and Mezich 1988). Davis and Whiting (1977) reported that nesting activity declined in Everglades National Park where dense stands of Australian pine took over nati ve dune vegetation on a remote nesting beach. Beach vitex ( Vitex rotundifolia ) was introduced to the horticulture trade in the mid-1980s and is

often sold as a “dune stabilizer.” The plant is native to countries in the western Pacific and was introduced into the southeastern U.S. in the mid-1980s. Its presence on North Carolina and South Carolina beaches have a negative effect on sea turtle nesting as its dense mats interfere with sea turtle nesting and hatchling emergence from nests (Bra bson 2006). This exotic plant is crowding out the native species, such as sea oats and bitter panicu m, and can colonize large areas in just a few years. Sisal, or century plant, ( Agave americana ) is native to arid regions of Mexico. The plant was widely grown in sandy so ils around Florida in order to provide fiber for cordage. It has escaped cultivation in Flor ida and has been purposely planted on dunes. Although the effects of sisal on sea turtle nesting are uncertain, thickets with impenetrable sharp spines are occasionally found on developed beaches.

Harmful algal blooms, such as a red tide, also im pact loggerheads in the Northwest Atlantic. In Florida, USA, the species that causes most red tides is Karenia brevis , a dinoflagellate that produces a toxin (Florida Marine Research Institute 2003). During four red tide events along the west coast of Florida, sea turtle stranding tre nds indicated that these events were acting as a mortality factor (Redlow et al . 2003). Sea turtles that washed ashore alive during these red tide events displayed symptoms that were consistent with acute brevitoxicosis (e.g., uncoordinated and lethargic but otherwise robus t and healthy in appearance) a nd completely recovered within days of being removed from the area of the red tid e. The population level effects of these events are not yet known.

5.2.6.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the No rthwest Atlantic Ocean. The following additional instruments also apply to loggerheads found in this area.

United States Magnuson-Stevens Fish ery Conservation and Management Act For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Inter-American Convention for the Protecti on and Conservation of Sea Turtles (IAC) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Convention for the Protection and Development of the Marine Environment of the Wider Caribbean Region Also called the Cartagena Conventio n, this instrument has been in place since 1986 and currently has 21 signatory states. Under this Convention, the component that may relate to loggerhead

turtles is the Protocol Concerni ng Specially Protected Areas and Wildlife (SPAW) that has been in place since 2000. The goals of this protocol are to encourage Parties "to take all appropriate measures to protect and preserve rare or fragil e ecosystems, as well as the habitat of depleted, threatened or endangered species, in the Convention area." All six sea turtle species in the Wider

Caribbean are listed in Annex II of the protocol, which prohib its (a) the taking, possession or134 killing (including, to the extent possible, the incidental taking, posse ssion or killing) or commercial trade in such species, their eggs, parts or products, and (b) to the extent possible, the disturbance of such species, part icularly during breeding, incubati on, estivation, migration, and other periods of biological stress. Hykle (2002) believes that in view of the limited participation

of Caribbean States in the af orementioned Convention on the Cons ervation of Migratory Species of Wild Animals, the provisions of the SPAW Pr otocol provide the legal support for domestic conservation measures that might otherwise not have been afforded. Additional information is

available at http://www.cep.unep.org/cartagena-convention

.

Convention for the Protection of the Marine Environment of the North-East Atlantic

This Convention is included under the Northwest A tlantic Ocean DPS because it is likely that oceanic juveniles from this DPS also are found in the Northeast Atlantic Ocean. For information about this instrument, see section 5.2.7.4. for the Northeast Atlantic Ocean DPS.

5.2.6.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear Bycatch of loggerheads in commercial and recreati onal fisheries in the No rthwest Atlantic is a significant threat facing the species in this region. A variety of fishing gears that incidentally capture loggerhead turtles are employed includ ing gillnets, trawls, hook and line, longlines, seines, dredges, pound nets, and various types of pots/traps. Among these, gillnets, longlines, and trawl gear contribute to the vast majority of bycatch mortality of loggerheads annually throughout their range in the Atlantic Ocean and Gulf of Mexico (Epperly et al . 1995; NMFS 2002a, 2004, 2007a, 2008a; Lewison et al . 2003, 2004; Richards 2007; NMFS, unpublished data). Considerable effort has been expende d since the 1980s to document and address fishery bycatch, especially in the United States and Mexi co. Observer programs have been implemented in some fisheries to collect turtle bycatch data , and efforts to reduce bycatch and mortality of loggerheads in certain fishing operations have been undertaken and implemented or partially implemented. These efforts include developing gear solutions to prevent or reduce captures or to allow turtles to escape without harm (e.g., TEDs, circle hooks and bait combinations), implementing time and area closures to prevent inte ractions from occurring (e.g., prohibitions on gillnet fishing along the mid-Atla ntic coast during the critical time of northward migration of loggerheads, implementation of careful release protocols (e.g., requirements for careful release of turtles captured in longline fisheries), prohibit ions of gillnetting in some U.S. state waters), and/or modifying gear (e.g., re quirements to reduce mesh size in the leaders of pound nets in certain U.S. coastal waters to prevent entanglement).

The primary bycatch reduction focus in the Nort hwest Atlantic, since the 1978 ESA listing of the loggerhead, has been on bycatch re duction in shrimp trawls. The development of turtle excluder devices (TEDs) in the 1970s and the refinement of these devices over the past three decades has been a primary focus of loggerhead bycatch reduc tion efforts. The U.S. has required the use of TEDs throughout the year since th e mid-1990s, with modifications required and implemented as necessary (NMFS 1987, 52 FR 24244; NMFS 1992, 57 FR 57348). Most notably, in 2003, NMFS implemented new requirements for TEDs in the shrimp trawl fishery to ensure that large loggerheads could escape thr ough TED openings (NMFS 2003, 68 FR 8456). Significant effort has been expended to transfer this technology to other shrimping fleets in the Northwest Atlantic;135 however, not all nations where loggerheads occur require the device be used. Enforcement of TED regulations is difficult and compliance is not believed to be complete. Because TEDs are not 100% effective, a significant number of loggerh eads are estimated to s till be killed annually in shrimp trawls throughout the Northwest Atlant ic. In the U.S. Southeast food shrimp trawl fishery, NMFS estimated the annual mortality of loggerheads in the Gulf of Mexico and Southeast U.S. Atlantic Ocean as 3,948 indi viduals (95% CI, 1,221-8,498) (NMFS 2002a). Shrimping effort in the Southeast U.S. has re portedly declined, a revised estimate of annual loggerhead mortality for the Gulf of Mexico segment of the Sout heast food shrimp trawl fishery is 647 individuals (NMFS, unpublished data).

Other trawl fisheries operating in Northwest Atlan tic waters that are known to capture sea turtles include, but are not limited to, su mmer flounder, calico scallop, sea scallop, blue crab, whelk, cannonball jellyfish, horseshoe crab, and mid-Atlan tic directed finfish trawl fisheries and the Sargassum fishery. In the U.S., the su mmer flounder fishery is the on ly trawl fishery (other than the shrimp fishery) with federally mandated TED use (in certain areas). Loggerhead annual bycatch estimates in 2004 and 2005 in U.S. mid-Atla ntic scallop trawl gear ranged from 81 to 191 turtles, depending on the estimation methodol ogy used (Murray 2007). Estimated average annual bycatch of loggerheads in other mid-A tlantic federally managed bottom otter trawl fisheries during 1996-2004 was 616 turtle s (Murray 2006). The harvest of Sargassum by

trawlers can result in incident al capture of post-hatchlings a nd habitat destruction (Schwartz 1988, Witherington 2002); this fish ery is not currently active.

Dredge fishing gear is the predominant gear us ed to harvest sea scallops off the mid- and northeastern U.S. Atlantic coast. Sea scallop dredges are composed of a heavy steel frame and cutting bar located on the bottom part of the frame and a bag, made of metal rings and mesh twine, attached to the frame. The gear is fished along the bottom and weighs from 500-1,000 pounds (National Research Council 2002). Turtles can be struck and injured or killed by the dredge frame and/or captured in the bag where they may drown or be further injured or killed when the catch and heavy gear are dumped on the vessel deck. Total estimated bycatch of loggerhead turtles in the U.S. sea scallop dre dge fishery operating in the mid-Atlantic region (New York to North Carolina) from June thr ough November is on the order of several hundred turtles per year (Murra y 2004, 2005, 2007). The impact of the sea scallop dredge fishery on loggerheads in U.S. waters of the Northw est Atlantic remains a serious concern.

Incidental take of oceanic-stage loggerheads in pelagic longline fisheries has recently received significant attention (Balazs and Pooley 1994; Bolten et al . 1994, 2000; Aguilar et al . 1995;

Laurent et al . 1998; Long and Schroeder 2004; Watson et al . 2005). Large-scale commercial longline fisheries operate throughout the pelagic range of the No rthwest Atlantic loggerhead, including the western Mediterranean. The largest size classes in the oceanic stage are the size classes impacted by the swordfish longline fisher y in the Azores (Bolten 2003) and on the Grand Banks (Watson et al . 2005), and this is likely the case for other nation’s fleets operating in the region, including but not limited to, the European Union, United States, Japan, and Taiwan. The demographic consequences relative to population recovery of the increas ed mortality of these size classes have been discussed (Crouse et al . 1987; see also Heppell et al . 2003 and Chaloupka 2003). Estim ates derived from data recorded by the international observer program (IOP) suggest that thousands of mos tly immature loggerheads have b een captured in the Canadian136 pelagic longline fishery in the western North A tlantic since 1999 (Brazn er and McMillan 2008). NMFS (2004a) estimates that 635 loggerheads (143 le thal) will be taken annually in the U.S. pelagic longline fishery.

Incidental capture of ne ritic-stage loggerheads in demersal lo ngline fishing gear has also been documented. Richards (2007) estimated total an nual bycatch of loggerheads in the Southeast

U.S. Atlantic and U.S. Gulf of Mexico commer cial directed shark bottom longline fishery from 2003-2005 as follows: 2003: 302 – 1,620 (CV 0.45); 2004: 95 – 591 (CV 0.49); 2005: 139-778 (CV 0.46). NMFS (2009) estimated the total number of captures of hardshe ll turtles in the U.S. Gulf of Mexico reef fish fishery (demersal longline fishery) from July 2006 – December 2008 as 861 turtles (95% CI 383 – 1934). These estimates are not comprehensive across this gear type (i.e., pelagic and demersal longline) throughout the Northwest Atlantic Ocean. Cumulatively, the bycatch and mortality of Nort hwest Atlantic loggerheads in long line fisheries is significant.

Gillnet fisheries may be the most ubiquitous of fisheries operating in the neritic range of the Northwest Atlantic loggerhead. Comprehensive es timates of bycatch in gillnet fisheries do not yet exist and, while this precludes a quantit ative analysis of thei r impacts on loggerhead populations, the cumulative mortality of loggerheads in gillnet fisheries is likely high. In the U.S. mid-Atlantic, the averag e annual estimated bycatch of loggerheads from 1995 – 2006 was 350 turtles (CV= 0.20., 95% CI ove r the 12-year period: 234 to 504) (Murray in press). In the U.S., some states (e.g., South Carolina, Georgia, Florida, Louisiana, and Texas) have prohibited gillnets in their waters, but there remain active gi llnet fisheries in other U.S. states, in U.S. federal waters, Mexican waters, Central and South America, and the Northeast Atlantic.

Pound nets are fixed gear composed of a seri es of poles driven in to the bottom upon which netting is suspended. Pound nets basically opera te like a trap with th e pound constructed of a series of funnels leading to a bag that is open at the top, and a long leader of netting that extends from shallow to deeper water where the pound is loca ted. In some configurations, the leader is suspended from the surface by a series of stringers or vertical lines. Se a turtles incidentally captured in the open top pound, which is composed of small mesh webbing, are usually safe from injury and may be released easily wh en the fishermen pull the nets (Mansfield et al . 2002). However, sea turtle mortalities have been docum ented in the leader of certain pound nets. Large mesh leaders (greater than 12-inch stretched mesh ) may act as a gillnet, entangling sea turtles by the head or foreflippers (Bellmund et al . 1987) or may act as a barrie r against which turtles may be impinged (NMFS, unpublished data). Nets wi th small mesh leaders (less than 8 inches stretched mesh) usually do not present a mortality threat to loggerheads, but some mortalities have been reported (Morreale and Standora 1998; Epperly et al . 2000, 2007; Mansfield et al

. 2002). In 2002, the U.S. prohibited, in certain areas within the Chesapeake Bay and at certain times, pound net leaders having mesh greater than or equal to 12 inches and leaders with stringers (NMFS 2002b, 67 FR 41196). Subsequent regulati ons have further restricted the use of certain pound net leaders in certain geographic areas and established pound net leader gear modifications (NMFS 2004b, 69 FR 24997; NMFS 2006, 71 FR 36024).

Pots/traps are commonly used to target crabs, lobsters, whelk, and reef fishes. These traps vary in size and configuration, but al l are attached to a surface float by means of a vertical line leading to the trap. Entanglement and mortality of l oggerheads has been documented in various pot/trap137 fisheries in the U.S. Atlantic and Gulf of Mexico. Data from the U.S. Sea Turtle Stranding and Salvage Network indicate that 82 loggerheads (d ead and rescued alive) were documented by the stranding network in various pot /trap gear from 1996-2005, of th ese approximately 30-40% were adults and the remainder juveniles (NMFS, unpublis hed data). Without intervention it is likely that the majority of the live, entangled turtles would die. Additionally, documented strandings represent only a portion of total interactions and mortality. Recently, a small number of loggerhead entanglements also have been recorded in whelk pot br idles in the U.S. Mid-Atlantic (Meredith Fagan, Virginia Institute of Ma rine Science, personal communication, 2008). However, no dedicated observer programs exist to provide estimates of take and mortality from pot/trap fisheries; therefore, comprehensive esti mates of loggerhead interactions with pot/trap gear are not available, but the gear is widely used throughout the range of the DPS, and poses a continuing threat.

Other Manmade and Natural Impacts Similar to other areas of the world (see Section 5.1. 5.), climate change and sea level rise have the potential to impact loggerheads in the Northwest Atlantic Ocean.

Propeller and collision injuries from boats and ships are becoming more common in sea turtles. In the U.S. Atlantic, from 1997 to 2005, 14.9% of all stranded loggerheads were documented as having sustained some type of propeller or co llision injuries (NMFS, unpublished data). The incidence of propeller wounds observed in sea turtles stranded in the U.S. has risen from approximately 10% in the late 1980s to a re cord high of 20.5% in 2004 (NMFS, unpublished data). In the U.S., propeller wounds are greatest in southeast Florida; during some years, as many as 60% of the loggerhead strandings found in these areas had propeller wounds (Florida Fish and Wildlife Conservation Commission, unpub lished data). As the number of vessels increases, in concert with increased coastal development, especially in nearshore waters, propeller and vessel collision injuries are also expected to rise.

Several activities associated with offshore oi l and gas production, including oil spills, water quality (operational discharge), seismic surveys, explosive platform removal, platform lighting, and noise from drillships and production activitie s, are known to impact loggerheads (National Research Council 1996; MMS 2000; Gregg Gitschlag, NMFS, personal communication, 2007;

Viada et al . 2008). Currently, there are 3,443 federally regulated offshore platforms in the Gulf of Mexico dedicated to natural gas and oil production. Additional state-regulated platforms are located in state waters (Texas and Louisiana). There are currently no active leases off the

Atlantic coast.

Oil spills also threaten loggerheads in the Northwes t Atlantic. Two oil spills that occurred near loggerhead nesting beaches in Fl orida were observed to affect eggs, hatchlings, and nesting females. Approximately 350,000 gallons of fuel oil spilled in Tampa Bay in August 1993 and was carried onto nesting beaches in Pinellas County. Observed mortalities included 31 hatchlings and 176 oil covered nests; an addi tional 2,177 eggs and hatchlings were either exposed to oil or disturbe d by response activities (FDEP et al . 1997). Another spill near the beaches of Broward County in August 2000 involved approximately 15,000 gallons of oil and tar (NOAA and FDEP 2002). Models estimated that approximately 1,500 to 2,000 hatchlings and 0 to 1 adults were injured or killed. Annually abou t 1% of all sea turtle strandings along the U.S.138 east coast have been associated with oil, but high er rates of 3 to 6% have been observed in South Florida and Texas (Teas 1994, Rabalais a nd Rabalais 1980, Plotkin and Amos 1990).

In addition to the destruction or degradation of habitat described in Sect ion 5.2.6.1., periodic dredging of sediments from navigational channels can also result in incidental mortality of sea turtles. Direct injury or mo rtality of loggerheads by dredges has been well documented in the

southeastern and mid-Atlantic U. S. (National Research Council 1990). Solutions, including modification of dredges and time/area closures, have been successfully implemented to reduce mortalities and injuries in the U.S. (NMFS 1991, 1995, 1997; Nelson and Shafer 1996).

The entrainment and entrapment of loggerheads in saltwater cooling inta ke systems of coastal power plants has been documented in New Jers ey, North Carolina, Florida, and Texas (Eggers 1989; National Research Council 1990; Caroli na Power and Light Company 2003; FPL and Quantum Resources, Inc. 2005; Pr ogress Energy Florida, Inc. 2003) . Average annual incidental capture rates for most coastal pl ants from which captures have b een reported amount to several turtles per plant per year. One notable exception is the St. Lucie Nuclear Power Plant located on Hutchinson Island, Florida. During the first 15 years of operation ( 1977-1991), an average of 128 loggerheads per year was captured in the intake canal with a mortality rate of 6.4%. During 1991-2005, loggerhead captures more than doubled (average of 308 per year), while mortality rates decreased to 0.3% per year ( FPL and Quantum Resources, Inc. 2005).

Although not a major source of mortality, cold stunning of loggerheads has been reported at several locations in U.S., including Cape Cod Bay, Massachusetts (Still et al . 2000); Long Island Sound, New York (Meylan and Sadove 1986, Morreale et al . 1992); the Indian River system, Florida (Mendonça and Ehrhart 1982, Witheringt on and Ehrhart 1989); and Texas inshore waters (Hildebrand 1982, Shaver 1990).

Another natural factor that has the potential to a ffect recovery of loggerh ead turtles is aperiodic hurricanes. In general, these events are ep isodic and, although they may affect loggerhead hatchling production, the results are generally localized and they rarely result in whole-scale losses over multiple nesting seasons. The negative effects of hurricanes on low-lying and/or developed shorelines may be longer-la sting and a greate r threat overall.

Similar to other areas of the world (see Section 5.1. 5.), climate change and sea level rise have the potential to impact loggerheads in the Northwest Atlantic. This includes beach erosion and loss from rising sea levels, repeated inundation of nests, skewed ha tchling sex ratios from rising beach incubation temperatures, and abrupt disrupti on of ocean currents used for natural dispersal during the complex life cycle.

5.2.7. Northeast Atlantic Ocean DPS

The following information on threats pertains to all loggerhead turtles that may be found in the Northeast Atlantic. Because it is likely that ocea nic juveniles from this DPS also are found in the

Northwest Atlantic Ocean and in the Mediterranean Sea, the narr atives for those regions should

also be consulted.139 5.2.7.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone As described in Section 5.1.1., destruction and modification of loggerhead nesting habitat in the

Northeast Atlantic result from coastal devel opment and construction, placement of erosion

control structures and other barriers to nesti ng, beachfront lighting, ve hicular and pedestrian traffic, sand extraction, beach erosion, a nd beach pollution (NMFS and FWS 1998, Formia et al

. 2003).

In the Northeast Atlantic, the only loggerhead nesti ng of note occurs in the Cape Verde Islands. The Cape Verde government’s plans to develop Boa Vista Island, the location of the main nesting beaches, could increase the terrestrial threats to loggerheads (v an Bogaert 2006). Sand extraction on Santiago Island, Cape Verde, may be responsible for the apparent decrease in nesting there (Loureiro 2008). Sca ttered and infrequent nesting o ccurs in western Africa, where

much industrialization is locat ed on the coast and population gr owth rates fluctuate between 0.8% (Cape Verde) and 3. 8% (Côte D’Ivoire) (Abe et al . 2004, Tayaa et al . 2005). Land mines on some of the beaches of mainland Africa, within the reported historical range of nesting by loggerheads (e.g., the Western Sahara region) would be detrimental to nesters and are an impediment to scientific su rveys of the region (Tiwari et al . 2001). Tiwari et al . (2001) noted a high level of human use of ma ny of the beaches in Morocco – enough that any evidence of nesting activity would be quickly erased. Ga rbage litters many developed beaches (Formia et al

. 2003). Erosion is a problem along the long stretc hes of high energy ocean shoreline of Africa and is further exacerbated by sand mining and harbor building (Formia et al . 2003); crumbling buildings claimed by the sea may presen t obstructions to nesting females.

Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, marine pollution and climate change. These threats also occur in the

Northeast Atlantic Ocean. Ec osystem alterations have occurr ed due to the tremendous human pressure on the environment in the region. Turt les, including loggerhea ds, usually are included in ecosystem models of the region (see Palomare s and Pauly 2004). In the Canary Current Large Marine Ecosystem (LME), the area is charac terized by the Global International Waters Assessment (GIWA) as severely impacted in th e area of modification or loss of ecosystems or ecotones and health impacts, but these impacts are decreasing ( http://www.lme.noaa.gov ). The Celtic-Biscay Shelf LME is affected by alterations to the seabed, agriculture, and sewage (Valdés and Lavin 2002). The Gulf of Guinea has b een characterized as severely impacted in the area of solid wastes by the GIWA; this a nd other pollution indi cators are increasing ( http://www.lme.noaa.gov ). Marine pollution, such as oil and debris, has been shown to negatively impact loggerheads and represen t a degradation of the habitat (Orós et al . 2005, 2009; Calabuig Miranda and Liria Loza 2007). Climate change also may result in future trophic changes, thus impacting loggerhead prey abundance and/or distribution.

Additionally, fishing is a major sour ce of ecosystem alteration of the neritic and oceanic habitats of loggerhead turtles in the region. Fishing effo rt off the western African coast is increasing and record low biomass has been recorded for expl oited resources, representing a 13X decline in140 biomass since 1960 (see Palomares and Pauly 2004) . Throughout the North Atlantic, fishery landings fell by 90% during the 20 th century, foreboding a trophic cascade and a change in foodweb competition (Pauly et al . 1998, Christensen et al . 2003). For a descrip tion of the exploited marine resources in the region, see Lamboeuf (1997). The Celtic-Biscay Shelf LME, the Iberian Coastal Ecosystem LME, the Canary Current LME, and the Guinea Current LME all are severely overfished, and effort now is turning to a focus on pelagic fisheries, wher eas historically there were demersal fisheries. The impacts continue to increase in the Guinea Current LME despite efforts throughout the region to reduce fishing pressure ( http://www.lme.noaa.gov ).

The threats to bottom habitat for loggerheads pr eviously described in this document include modification of the habitat through bottom trawling. Trawling occurs off the European coast and

the area off Northwest Africa is one of the most intensively trawled areas in the world (Zeeberg

et al . 2006). Trawling has been bann ed in the Azores, Madeira, a nd Canary Islands to protect cold-water corals (Lutter 2005). Although illegal, trawling also occu rs in the Cape Verde Islands (López-Jurado et al . 2003). The use of destru ctive fishing practices, such as explosives and toxic chemicals, has been reported in the Canary Current area, causing serious damage to both the resources and the habitat (Tayaa et al . 2005).

5.2.7.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

Deliberate hunting of loggerheads for their meat, shells, and eggs still exists and remains the most serious threat facing nesting turtles in the Northeast Atlantic. Historical records indicate turtles were harvested throughout Macaronesia (s ee López-Jurado 2007). Intensive exploitation has been cited for the extirpation of the loggerh ead nesting colony in the Canary Islands (LópezJurado 2007), and heavy human predation on nesti ng and foraging animals occurred on Santiago Island, Cape Verde, the first in the Archipelago to be settled (Loureiro 2008), as well as on Sal and Sao Vicente islands (López-Jurado 2007). Nes ting loggerheads and eggs are still harvested at Boa Vista, Cape Verde (Cabrera et al . 2000, López-Jurado et al . 2003). In 2007, over 1,100 (36%) of the nesting turtles were hunted, which is about 15% of the estimated adult female population (Marco et al . in press). In 2008, the military protected one of the major nesting

beaches on Boa Vista where in 2007 55% of the mortality had occurred ; with the additional protection, only 17% of the turtles on that beach were slaughtered (Roder et al . in press). On Sal Island, 11.5% of the emergences on unprotected b eaches ended with mortality, whereas mortality was 3% of the emergences on protected beaches (Cozens et al . in press). The slaughter of nesting turtles is a problem wherever turtles ne st in the Cape Verde Islands and may approach 100% in some places (C. Roder, Turtle Foundation, Münsing, Germany, personal communication, 2009; Cozens in press). The meat and eggs are consumed locally as well as traded among the archipelago (C. Roder, Turtle Foundation, Münsing, Germany, personal communication, 2009). Hatchlings ar e collected on Sal Island, but this activity appears to be rare on other islands of the archipelago (J. Cozens, SOS Tartarugas Cabo Verde, Santa Maria, Sal Island, Cape Verde, personal communication, 2009) . Additionally, free dive rs target turtles for consumption of meat, often selectiv ely taking large males (López-Jurado et al . 2003). Turtles are harvested along the African coas t and, in some areas, are considered a significant source of food and income due to the poverty of many residents along the African coast (Formia et al

. 2003). Loggerhead carapaces are sold in markets in Morocco and Western Sahara (Fretey 2001,

Tiwari et al . 2001, Benhardouze et al . 2004).141 5.2.7.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the Northeas t Atlantic Ocean. Spontaneous diseases documented in the Northeast Atlantic include pneumonia, hepatitis, meningitis, septicemic processes, and neoplasia (Orós et al . 2005). Pneumonia could result from the aspiration of water from forced submergence in fishing gear. The authors al so reported nephritis, esophagitis, nematode infestation, and eye lesions. Fibr opapillomatosis does not appear to be an issue in the Northeast

Atlantic.

Nest depredation by ghost crabs ( Ocypode cursor ) occurs in Cape Verde (López-Jurado et al

. 2000). The ghost crabs feed on both eggs and hatchlings. Arvy et al . (2000) reported predation of loggerhead eggs in two nests in Mauritania by golden jackals ( Canis aureus ); a loggerhead turtle creating a third nest also had been killed , with meat and eggs eaten, but the predator was

not identified.

Loggerheads in the Northeast Atlantic also ma y be impacted by harmful algal blooms, which have been reported infrequently in the Canary Islands and th e Iberian Coastal LME (Ramos et al

. 2005, Akin-Oriola et al . 2006, Amorim and Dale 2006, Moita et al . 2006; http://www.lme.noaa.gov ).

5.2.7.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the Northeast Atlantic Ocean, a lthough not all nations in the region are signatory to them all (e.g., West ern Sahara is not signa tory to the CMS/Bonn Convention). The following additional instrume nts and regulation also apply to loggerheads

found in this area.

Convention for the Protection of the Marine Environment of the North-East Atlantic Also called the OSPAR Convention, this 1992 in strument 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 mi ssion of the OSPAR Convention “...is to conserve marine ecosystems and safeguard human health in the North-East Atlantic by preventing and eliminating pollution; by protecti ng the marine environment from the adverse effects of human activities; and by contributing to the sustainable use of the seas.” Loggerheads

are included on the OSPAR List of Threatened a nd/or Declining Species and Habitats, which is used by the OSPAR Commission fo r setting priorities for work on the conservation and protection of marine biodiversity. Additional information is available at http://www.ospar.org

.142 Convention on the Conservation of Eur opean Wildlife and Natural Habitats Also known as the Bern Convention, the goals of th is instrument are to co nserve wild flora and fauna and their natural habitats, especially those species and habita ts whose conservation requires the cooperation of severa l States, and to promote such cooperation. The Convention was enacted in 1982 and currently includes 48 Eur opean and African States and the European Union. Sea turtles are included on the “strictly protected” list. Acco rding to Hykle (2002), while the Convention's “innovative approach to holding States to account for their implementation of the Convention is laudable, and has certainly drawn attention to issues of species and habitat protection, its efficacy in relation to particular marine turtle cases that have been deliberated for many years is debatable.” Additional information is available at http://conventions.coe.int/Tre aty/EN/Treaties/Htm l/104.htm

.

African Convention on the Conservation of Natu re 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 a nd with due regard to th e best interests of the people”. It was followed by the 1972 Stockhol m Conference on the Human Environment and

led to the establishment of environmental minist ries in African nations and the creation of the United Nations Environment Programme (UNEP) headquartered in Nairobi. The Algiers Convention recently has undergone re vision (not yet in force) and its objectives are to enhance environmental protection, foster conservation an d 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

.

Convention on the Conservation of Migratory Species of Wild Animals, 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 co ast 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 loggerhead tu rtle – 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 projection of hatch lings 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 ac tivities. However, despite this agreement, killing of adult turtles and harvesting of eggs remains ramp ant in many areas along the Atlantic

African coast. Additional information is available at http://www.cms.int/species/afri ca_turtle/AFRICAt urtle_bkgd.htm

.

The Convention for the Co-operation in the Protection and Development of the Marine and Coastal Environment of the West and Ce ntral African Region (Abidjan Convention) The Abidjan Convention covers the marine environm ent, coastal zones, a nd related inland waters from Mauritania to Namibia. The Abidjan C onvention countries are An gola, Benin, Cameroon,143 Cape Verde, Congo, Cote d'Ivoire, Democratic Republic of Congo, Equatorial Guinea, Gabon, Gambia, Ghana, Guinea, Guinea-Bissau, Liberia, Mauritania, Namibia, Ni geria, Sao Tome and Principe, Senegal, Sierra Leone, and Togo. Th e Abidjan Convention is an agreement for the protection and management of the marine and coas tal areas that highlights sources of pollution, including pollution from ships, dumping, land-base d sources, exploration a nd exploitation of the sea-bed, and pollution from or through the atmo sphere. The Convention also identifies where co-operative environmental management efforts are needed. These areas of concern include coastal erosion, specially prot ected areas, combating pollution in cases of emergency, and environmental impact assessment. The Action Plan and the Abidjan Convention were adopted by the Governments in 1981; the Convention entere d into force in 1984. Western Sahara and Morocco are not signatories of the Abidjan Convention.

Accra Declaration of the Ministerial Committee of the Gulf of Guinea Large Marine Ecosystem (GOG-LME)-1998 Abuja Declaration of the Guinea Current Large Marine Ecosystem Project2006 In 1998, the environmental ministers of Cote d’Ivoire, Ghana, Togo, Benin, Nigeria, and Cameroon signed the Accra Declaration to streng then regional capacity to prevent and correct pollution in the LME and prevent and correct degradation of cri tical habitats. The ministers identified the living resources and management problems in the area. The countries decided on a detailed survey of industries, defined regional effluent standards, instituted community based mangrove restoration activities, and created a campaign for the re duction, recovery, recycling, and re-use of industrial wastes. In 2006, th e Guinea Current LME Proj ect expanded the project scope to 10 neighboring countries (Guinea-Bissau, Guinea, Sierra Leone, Liberia, Sao Tome and Principe, Equatorial Guinea, Gabon, Congo Brazzaville, Congo-Kinshasa, and Angola).

Council Regulation (EC) No

. 1239/98 of 8 June 1998 Amending Regulation (EC) No

. 894/97 Laying Down Certain Technical Measures for th e Conservation of Fishery Measures (Council of the European Union) This measure banned the use of driftnets by 1 January 2002 for European fleets. Fleets from other nations fishing in international waters can still use driftnets.

5.2.7.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear Loggerhead turtles strand throughout the No rtheast Atlantic (F retey 2001; Tiwari et al . 2001; Duguy et al . 2004, 2005; Witt et al . 2007), and there are indicati ons that the turtles become entangled in nets and monofilament and swallow hooks in the region (Orós et al . 2005, Calabuig Miranda and Liria Loza 2007). On the European coasts, most stranded loggerheads are small (mean < 30 cm SCL), but a few ar e greater than 60 cm SCL (Witt et al . 2007). Similarly, Tiwari

et al . (2001) and Benhardouze et al . (2004) indicated that the anim als they viewed in Morocco and Western Sahara were small juveniles and pr eliminary genetic analys es of stranded turtles indicate that they are of western Atlantic origin (M. Tiwari, NMFS, and A. Bolten, University of Florida, unpublished data), whereas Fretey ( 2001) reported that l oggerheads captured and stranded in Mauritania were both ju venile and adult-sized animals.144 Incidental capture of sea turtles in artisanal and commercial fisheries is a threat to the survival of loggerheads in the Northeast Atlantic. Sea turtles may be caught in a multitude of gears deployed in the region: pelagi c and demersal longlines, drift and set gillnets, bottom and midwater trawling, weirs, haul and purse seines, pots and traps, cast nets, and hook and line gear (see Pascoe and Gréboval 2003, Bayliff et al . 2005, Tayaa et al . 2005, Dossa et al . 2007). Fishing

effort off the western African coast has b een increasing (see Palomares and Pauly 2004). Impacts continue to increase in the Guinea Curr ent LME, but, in contrast, the impacts are reported to be decreasing in the Canary Current LME ( http://www.lme.noaa.gov ). Throughout the region, fish stocks are depl eted and management authorities are striving to reduce the fishing

pressure.

In the Northeast Atlantic, loggerheads, particul arly the largest size classes in the oceanic environment (most of which are small juvenile s), are captured in surface longline fisheries targeting swordfish ( Ziphias gladius ) and tuna ( Thunnus spp .) (Ferreira et al . 2001, Bolten 2003). Bottom longlines in Madeir a Island targeting black-scabbard ( Aphanopus carbo ) capture and kill small juvenile loggerhead turtles as the fishing depth does not allow hooked turtles to surface (Dellinger and Encarnaçâo 2000, Delgado et al . in press).

In United Kingdom and Irish waters, loggerhead bycatch is uncommon but has been noted in pelagic driftnet fisheries (Pie rpoint 2000, Rogan and Mackey 2007). Loggerheads have not been captured in pelagic trawls, demersal trawls, or gillnets in United Kingdom and Irish waters (Pierpoint 2000), but have been captu red in nets off France (Duguy 2004, 2005).

International fleets of trawl fisheries operate in Mauritania and have been documented to capture sea turtles, including loggerheads (Zeeberg et al . 2006). Despite being illegal, trawling occurs in the Cape Verde Islands and has the potential to capture and kill loggerhead turtles; one piece of abandoned trawl net washed shore with eigh t live and two dead loggerheads (López-Jurado et al

. 2003). Longlines, seines, and hook and line have been documented to capture loggerheads 35-73

cm SCL off the northwestern Mo roccan coast (Benhardouze 2004).

Other Manmade and Natural Impacts As described in Section 5.1.5., other anthropogenic impacts, such as boat strikes and ingestion or entanglement in marine debris, also apply to l oggerheads in the Northeast Atlantic. Propeller and boat strike injuries have been docu mented in the Northeast Atlantic (Orós et al . 2005, Calabuig Miranda and Liria Loza 2007). Exposure to crude oil is also of concern. Loggerhead strandings in the Canary Islands have s hown evidence of hydrocarbon exposure as well as ingestion of marine debris, such as plastic and monofilament (Orós et al . 2005, Calabuig

Miranda and Liria Loza 2007), and in the Azores and elsewh ere plastic debris is found both on the beaches and floating in the waters (Barrerios and Barcelos 2001, Tiwari et al . 2001). Pollution from heavy metals is a concern for the seas around the Iberian Peninsula (European Environmental Agency 1998) and in the Guinea Current LME (Abe et al . 2004). Bioaccumulation of metals in loggerheads has be en measured in the Canary Islands and along the French Atlantic Coast (Caurant et al . 1999, Torrent et al . 2004). However, the consequences of long term exposure to heavy metals are unknown (Torrent et al . 2004). The effects of pollution on the ecosystem are a ddressed in Section 5.2.7.1.145 Natural environmental events, such as climate ch ange, could impact affect loggerheads in the

Northeast Atlantic. Similar to other areas of th e world (see Section 5.1.5.), climate change and sea level rise have the potential to impact logge rheads in the Northeast Atlantic, and the changes may be further exacerbated by the burning of fossil fuels and defore station. These effects range from flooding of nesting beaches, shifts in ocean currents, ecosystem shifts in prey distribution and abundance, and a shift in the sex ratio of the population if rookeries do not migrate concurrently (e.g., northward in the case of global warming) or if nesting phenology does not change (see Doody et al . 2006). Tropical and sub-tropical storms occasionally strike the area and could have a negative impact on nesting, although such an impact would be of limited

duration.

5.2.8. Mediterranean Sea DPS

The following information on threats pertains to all loggerhead turtles that may be found in the

Mediterranean Sea. Because it is likely that oceanic juveniles from this DPS also are found in the Northeast Atlantic Ocean, the narrative for that region should also be consulted.

5.2.8.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone In the Mediterranean, some areas known to host nes ting activity in the past have been lost to turtles (e.g., Malta) or severely de graded (e.g., Israel) (Margaritoulis et al . 2003). As described in Section 5.1.1., destruction and modification of loggerhead nesting habitat in the Mediterranean result from coastal development and constructio n, placement of erosion control structures and other barriers to nesting, beachfront lighting, ve hicular and pedestrian tr affic, sand extraction, beach erosion, beach sand placement, beach pollution, removal of native vegetation, and planting of non-native vegetation (Bal dwin 1992, Margaritoulis et al . 2003). These activities may directly impact the nesting su ccess of loggerheads and survivabil ity of eggs and hatchlings. Nesting in the Mediterranean almost exclusively occurs in the Eastern basin, with the main concentrations found in Cyprus, Greece, Turkey, and Libya (Margaritoulis et al . 2003, Laurent et
al . 1999); therefore, the fo llowing threats to the ne sting habitat are concentrated in these areas.

The Mediterranean experiences a large influx of tourists during the summer months, coinciding with the nesting season. Margaritoulis et al . (2003) state that exte nsive urbanization of the coastline, largely a result of tourism and recr eation, is likely the most serious threat to loggerhead nesting areas. The la rge numbers of tourists that us e Mediterranean beaches result in an increase in umbrellas, chairs, garbage, and towe ls, as well as related hotels, restaurants, and stationary (e.g., street lights, hotels) and moving (e.g., cars) ligh ting, all which can impact sea turtle nesting success (D emetropoulos 2000). Further, the easte rn Mediterranean is exposed to high levels of pollution and marine debris, in pa rticular the nesting beaches of Cyprus, Turkey, and Egypt (Camiñas 2004).

Construction and infrastructure development also have the potential to alter nesting beaches and subsequently impact nesting success. The c onstruction of new buildings on or near nesting beaches has been a problem in Greece and Turkey (Camiñas 2004). The construction of a jetty146 and waterworks around Mersin, Turkey, has contribu ted significantly to the continuous loss of adjacent beach (Camiñas 2004).

Beach erosion and sand extraction also pose a problem for sea turtle nesting sites. The noted decline of the nesting population at Rethymno, Island of Crete, Gr eece, is partly attributed to beach erosion caused by construction on the high b each and at sea (e.g., groins) (Margaritoulis et
al . 2009). A 2001 survey of Lebanese nesting b eaches found severe erosion on beaches where previous nesting had been reported, and in some cases the beaches had disappeared completely (Venizelos et al . 2005). Definitive causes of this eros ion were found to be sand extraction, offshore sand dredging, and sediment removal from river beds for construction and military purposes. Beach erosion also may occur from natural changes, with the same deleterious effects to loggerhead nesting. On Patara, Turkey, b each erosion and subsequent inundation by waves and shifting sand dunes are responsi ble for about half of all logge rhead nest losses (Camiñas 2004). Erosion can further be exacerbated when native dune vegetation, which enhances beach stability and acts as an integral buffer zone betwee n land and sea, is degraded or destroyed. This in turn often leaves insufficient nesting opportun ities above the high tide line, and nests may be washed out. In contrast, the plan ting or invasion of less stabilizing, non-native plants can lead to increased erosion and degradation of suitable nesti ng habitat. Finally, sand extraction has been a serious problem on Mediterranean nesting beaches , especially in Turkey (Türkozan and Baran 1996), Cyprus (Godley et al . 1996, Demetropoulos and Hadjichristophorou 1989), and Israel (Levy 2003).

While the most obvious effect of nesting beac h destruction and modification may be to the existence of the actual nests, hatchlings are al so threatened by habitat alteration. In the Mediterranean, disorientation of ha tchlings due to artificial lighti ng has been recorded mainly in Greece (Rees 2005, Margaritoulis et al . 2007, Margaritoulis et al . 2009), Turkey (Türkozan and Baran 1996), and Lebanon (Newbury et al . 2002). Additionally, vehicle traffic on nesting beaches may disrupt the natural beach environmen t and contribute to erosion, especially during high tides or on narrow beaches where driving is concentrated on the high beach and foredune. On Zakynthos Island in Greece, Venizelos et al . (2006) reported that vehicles drove along the beach and sand dunes throughout th e tourist season on East Laganas and Kalamaki beaches, leaving deep ruts in the sand, disturbing sea turt les trying to nest, and im pacting hatchlings trying

to reach the sea.

Neritic/Oceanic Zone As described in Section 5.1.1., threats to habitat in the loggerh ead neritic and/or oceanic zones include fishing practices, channel dredging, sa nd extraction, marine pollution, and climate change. These threats also occur in the Mediterranean.

Trawling occurs throughout the Me diterranean, most notably in areas off Albania, Algeria, Croatia, Egypt, France, Greece, Italy, Libya, Mo rocco, Slovenia, Spain, Tunisia, and Turkey (Gerosa and Casale 1999, Camiñas 2004, Casale 2008). This fishing practice has the potential to destroy bottom habitat in these areas. Fishing me thods affect neritic zones by not only impacting bottom habitat and incidentally capturing loggerheads but also depleting fish populations, and thus altering ecosystem dynamics. For exampl e, depleted fish stocks in Zakynthos, Greece, likely contributed to predation of adult loggerheads by monk seals ( Monachus monachus )147 (Margaritoulis et al . 1996). Further, by depleting fish populations, the trophic dynamics will be altered, which may then in turn affect the abil ity of loggerheads to find prey resources. If loggerheads are not able to forage on the necessa ry prey resources, their long term survivability may be impacted. Climate change also may re sult in future trophic changes, thus impacting loggerhead prey abundance and/or distribution.

Marine pollution, including dir ect contamination and structural habitat degradation, can affect loggerhead neritic and oceanic habitat. As the Mediterranean is an enclosed sea, organic and inorganic wastes, toxic effluent s, and other pollutants rapidly affect the ecosystem (Camiñas 2004). The Mediterranean has been declared a “special area” by the MARPOL Convention, in which deliberate petroleum discharges from vess els 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 are as high as 1,200,000 metric tons (Alpers 1993). Direct oil spill events also occur as happened in Lebanon in 2006 when 10,000 to 15,000 tons of heavy fuel oil spilled into the eastern Mediterranean (United Na tions Environment Programme 2007).

Destruction and modification of logg erhead habitat also may occur as a result of other activities. For example, underwater explosives have been identified as a key threat to loggerhead habitat in internesting areas in the Mediterranean (Margaritoulis et al . 2003). Further, the Mediterranean is

a site of intense tourist activ ity, and corresponding boat anchori ng also may impact loggerhead

habitat in the neritic environment.

5.2.8.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

Mediterranean turtle populations were subj ect to severe explo itation until the mid-1960s (Margaritoulis et al . 2003). Deliberate hunting of loggerheads for their meat, shells, and eggs is reduced from previous exploitati on levels, but still exists. For example, Nada and Casale (2008) found that egg collection (for indivi dual consumption) still occurs in Egypt. In some areas of the Mediterranean, like on the Greek Island of Zakynthos, nesting beaches are protected (Panagopoulou et al . 2008), so egg harvest by humans in those areas is likely negligible.

Exploitation of juveniles and adults still occurs in some Mediterranean areas. In Tunisia, clandestine trade for local consumption is still re corded, despite prohibition of the sale of turtles in fish markets in 1989 (Laurent et al . 1996). In Egypt, turtles are so ld in fish markets despite prohibitive laws; of 71 turtles observed at fish markets in 1995 and 1996, 68% were loggerheads (Laurent et al . 1996). Nada (2001) reported 135 turtle s (of which 85% were loggerheads) slaughtered at the fish market of Alexandria in six months (December 1998-May 1999). Based on observed sea turtle slaught ers in 1995 and 1996, Laurent et al . (1996) estimated that several thousand sea turtles were probably killed each ye ar in Egypt. More recently, a study found that the open selling of sea turtles in Egypt generally has been curtailed due to enforcement efforts, but a high level of intentional ki lling for the black market or for direct personal consumption still exists (Nada and Casale 2008). Given the high numbers of turtle s caught in this area, several hundred turtles are currently estimated to be sl aughtered each year in Egypt (Nada and Casale 2008). This estimate likely includes both juven ile and adult loggerheads, as Egyptian fish markets have been documented selling different sized sea turtles. While the mean sea turtle size148 was 65.7 cm CCL (range 38-86.3 cm CCL; n=48), 37. 5% of observed loggerhead samples were greater than 70 cm CCL (Laurent et al . 1996).

5.2.8.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the Mediterranean. Endoparasi tes in loggerheads have been studied in the

western Mediterranean. While th e composition of the gastrointestinal community of sea turtles is expected to include digeneans, nematode s, and aspidogastreans, loggerheads in the Mediterranean were found to harbor only four digenean species typical of marine turtles (Aznar

et al . 1998). There have been no r ecords of fibropapillomatosis in the Mediterranean. While there is the potential for diseas e in this area, information on the prevalence of such disease is lacking.

In the Mediterranean Sea, loggerh ead hatchlings and eggs are s ubject to depredation by wild canids (i.e., foxes ( Vulpes vulpes ), golden jackals ( Canis aureus )), feral/domestic dogs, and ghost crabs ( Ocypode cursor ) (Margaritoulis et al . 2003). Predators have cau sed the loss of 48.4% of loggerhead clutches at Kyparissia Bay, Greece (Margaritoulis 1988), 70-80% at Dalyan Beach, Turkey (Erk'akan 1993), 36% (includes green turtle clutches) in Cyprus (Broderick and Godley 1996), and 44.8% in Libya (Laurent et al . 1995). A survey of the Syrian coast in 1999 found 100% nest predation, mostly due to stray dogs and humans (Venizelos et al . 2005). Loggerhead eggs are also depredated by inse ct larvae in Cyprus (McGowan et al . 2001), Turkey (Özdemir et
al . 2004), and Greece (Lazou and Rees 2006). Ghost crabs have been reported preying on loggerhead hatchlings in northern Cyprus and Egypt, suggesting 66% of emerging hatchlings succumb to this mortality source (Simms et al . 2002). Predation also has been influenced by anthropogenic sources. On Zakynthos, Greece, a landf ill site next to loggerhead nesting beaches has resulted in an artificially high level of seagulls ( Larus spp.), which results in increased predation pressure on hatchlings (Panagopoulou et al . 2008). Planting of non- native plants also

can have a detrimental effect on nests in the fo rm of roots invading eggs (e.g., tamarisk tree ( Tamarix spp

. ) roots invading eggs in Za kynthos, Greece; Margaritoulis et al . 2007).

Predation on adult and juvenile loggerheads has also been docum ented in the Mediterranean. Predation of nesting loggerheads by golden jack als has been recorded in Turkey (Peters et al

. 1994). During a 1995 survey of loggerhead nesti ng in Libya, two nesting females were found killed by carnivores, probably jackals (Laurent et al . 1997). Off the sea turtle nesting beach of Zakynthos, Greece, adult loggerheads were found being predated upon by Mediterranean monk seals ( Monachus monachus ). Of the eight predated turtles observed or reported, 62.5% were adult males (Margaritoulis et al . 1996). Further, stomach contents were examined from 24 Mediterranean white sharks ( Carcharodon carcharias ), and 17% contained remains of marine turtles, including two loggerheads, one gr een, and one unidentifiable turtle (Fergusson et al

. 2000). One of the loggerhead turtles ingested was a juvenile with a carapace length of approximately 60 cm (length not reported as either SCL or CCL). Fergusson et al . (2000) report

that white shark interactions with sea turtles are likely rare east of the I onian Sea, and while the impact of shark predation on tu rtle populations is unknown, it is probably small compared to other sources of mortality.149 The Mediterranean is a low-productivity body of wa ter, with high water cl arity as a result. However, harmful algal blooms do occur in this area (e.g., off Algeria in 2002), and the problem is particularly acute in enclosed ocean basins such as the Medite rranean. In the northern Adriatic Sea, fish kills have occurred as a result of noxious phytoplankton blooms and anoxic conditions (Mediterranean Sea LME). While fish may be mo re susceptible to thes e harmful algal blooms, loggerheads in the Mediterranean also may be impacted by such noxious or toxic phytoplankton

to some extent.

5.2.8.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the Mediterranean Sea. The following additional instruments and regulation also appl y to loggerheads found in this area.

Protocol Concerning Specially Protected Area s and Biological Diversity in the Mediterranean This Protocol is under the auspic es of the Barcelona Convention of 1976 for the Protection of the Mediterranean Sea against Pollution (amended in 1995). The Protocol has been in force since 1999 and includes general provisions to protect sea turtles and their habitats within the

Mediterranean Sea. The Protocol requires Parties to pr otect, preserve, and manage threatened or endangered species, establish protected areas, and coordinate bilateral or multilateral conservation efforts (Hykle 2002). In the framewor k of the Barcelona Convention, to which all Mediterranean countries are part ies, the Action Plan for the Conservation of Mediterranean

Marine Turtles has been in eff ect since 1989. The main objectiv es of the Action Plan are to enhance the population of marine turtles, protect their critical habitats (including nesting, feeding, wintering, and migrating areas), a nd improve scientific know ledge by research and monitoring. In particular, the Action Plan has focused on promulgati ng education and training (especially among fishermen), establishing and im proving rescue centers, promoting legislation guidelines, and improving research and monito ring of sea turtles (D emetropoulos 2007). Unfortunately, little research on or implementation of fishing gear bycatch techniques to reduce sea turtle incidental captures has occurred. Additional information is available at http://www.rac-spa.org

.

Convention on the Conservation of Eur opean Wildlife and Natural Habitats For information about this instrument, see section 5.2.7.4. for the Northeast Atlantic Ocean DPS.

Council Directive 92/43/EEC on the Conservation of Natural H abitats and of Wild Fauna and Flora (EC Habitats Directive) The EC Habitats Directive was adopted by the European Community in May 1992, as the means by which the Community meets its ob ligations as a signatory of the Bern Convention. It aims to protect approximately 200 habitats and 1,000 animal and plant species list ed in the Directive's Annexes, of which loggerhead sea turtles are included. The provi sions in the Directive require Member States to introduce a range of measures including the protection of species listed in the Annexes, undertake surveillance of habitats and species, and pr oduce a report every six years on the implementation of the Directive. The first co mplete set of country data was reported in 2007.

The Directive led to the establishment of a ne twork of Special Areas of Conservation that, together with the existing Special Protection Areas classified under th e separate EC Birds150 Directive, form a network of protected areas known as Natura 2000. Additional information is

available at http://ec.europa.eu/environm ent/natu re/legislation/habitatsdirective/index_en.htm

.

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) For information about this instrument, see section 5.2.7.4. for the Northeast Atlantic Ocean DPS.

Convention for the Protection of the Marine Environment of the North-East Atlantic

This Convention is included under the Mediterranean Sea DPS because it is likely that oceanic juveniles from this DPS also are found in the No rtheast Atlantic Ocean. For information about this instrument, see section 5.2.7.4. for the Northeast Atlantic Ocean DPS.

In addition to these international mechanisms, most Mediterranean countries have developed legislation to protect sea turtles and/or nesting habitats (Margaritoulis 2007). For example, in 1999, a National Marine Park was established on Zakynthos in western Greece, with the primary aim to provide protection to loggerhead nesti ng areas (Dimopoulos 2001). However, while improved over the last several year s, it appears that management a nd protection at this Park is lacking (Robinson et al . 2007, Panagopoulou et al. 2008). In Turkey, five nesting beaches were designated Specially Protected Area status, a nd in Cyprus, two nesting beaches have been afforded protection through the Fisheries Regula tion, with a maritime zone extending to the 20 meter isobath (Margaritoulis 2007). In Italy, a reserve to protect nesting on Lampedusa was established in 1984 (Margaritoulis et al . 2003). The overall effectivenes s of this country specific legislation is unknown at this time.

5.2.8.5. Other Natural or Manmade Factors Affecting Its Continued Existence

As noted in Section 5.1.5., othe r anthropogenic and natural factor s affecting loggerhead survival include incidental bycatch in fisheries, vessel collisions, marine pollution, climate change, and cyclonic storm events. Fishing practices alone ha ve been estimated to result in over 150,000 sea turtle captures per year, with approximately 50,000 mo rtalities (Casale 2008).

The only estimation of loggerhead survival probab ilities in the Mediterranean was calculated by using capture-mark-recapture t echniques from 1981-2003 (Casale et al . 2007c). Of the 3,254 loggerheads tagged, 134 were reca ptured at different sites thr oughout the Mediterranean. Most recaptured animals were juveniles (mean 54.4 cm CCL; range 25-88 cm CCL), but the study did not delineate between juvenile lif e stages. This research estimated a loggerhead annual survival probability of 0.73 (95% confidence interv als; 0.67-0.78), recogni zing that there are methodological limitations of the tec hnique used. Nonetheless, Casale et al . (2007a) stated that assuming a natural survivorship no higher than 0.95 and a tag loss rate of 0.1, a range of 0.1-0.2 appears reasonable for the additional huma n induced mortality (from all sources).

Incidental Bycatch in Fishing Gear Incidental capture of sea turtles in artisanal and commercial fisher ies is a significant threat to the survivability of loggerheads in the Mediterranean. Sea turtles may be caught in pelagic and demersal longlines, drift gillnets, set gillnets and trammel nets, bottom and mid-water trawls,151 seines, dredges, traps/pots, and hook and line gear. In a 2004 FAO Fisheries Report, Camiñas (2004) stated that the main fisher ies affecting sea turtles in the Mediterranean Sea (at that time) were Spanish and Italian longline, North Adriat ic Italian, Tunisian, and Turkish trawl, and

Moroccan and Italian driftnet. Available inform ation on sea turtle bycatch by gear type is discussed below. There is growing evidence th at artisanal/small vessel fisheries (set gillnet, bottom longline, and part of the pelagic longline fi shery) may be responsible for a comparable or higher number of captures with higher mortality ra tes than the commercial/large vessel fisheries (Casale 2008) as previously sugge sted by indirect clues (Casale et al . 2005a).

Mediterranean fish landings have increased steadily since the 1950s, but the FAO 10-year capture trend from 1990-1999 shows stab le landings (Mediterranean LME, http://www.lme.noaa.gov ). However, stable fish landings may result from stable fishing effort at the same catch rates, or higher fishing effort at lower catch rates. As fish stocks in the Mediterranean are being depleted (P. Casale , MTSG-IUCN Italy, personal communication, 2009), fishing effort in some areas may be increasi ng to catch the available fish. This trend has not yet been verified throughout th e Mediterranean, but fishing pr essures may be increasing even though landings appear stable.

Longline fisheries In the Mediterranean, pelagic longlin e fisheries targeting swordfish ( Ziphias gladius ) and albacore ( Thunnus alalunga ) may be the primary source of loggerhead bycatch. It appears that most of the incidental captures occur in the western and central portions of the area (Demetropoulos and Hadjichrist ophorou 1995). The most severe bycatch in the Mediterranean occurs around the Balearic Islands where 1, 950-35,000 juveniles are caught annually in the surface longline fishery (Mayol and Castel ló Mas 1983; Camiñas 1988, 1997; Aguilar et al

. 1995). Specifically, the following regions have repor ted annual estimates of total turtle bycatch from pelagic longlines: Spain – 17,000 to 35,000 turtles (Aguilar et al . 1995, Camiñas et al

. 2003); Italy (Ionian Sea) – 1, 084 to 4,447 turtles (Deflorio et al . 2005); Morocco – 3,000 turtles (Laurent 1990); Greece – 280 to 3,310 turtles (Panou et al . 1999, Kapantagakis and Lioudakis 2006); Italy (Lampedusa) – 2,100 turtles (Casale et al . 2007a); Malta – 1,500 to 2,500 turtles (Gramentz 1989); South Tunisia (Gulf of Gabès) – 486 turtles (Jribi et al . 2008); and Algeria – 300 turtles (Laurent 1990).

For the entire Mediterranean pelagic longline fish ery, an extrapolation resulted in a bycatch estimate of 60,000 to 80,000 loggerheads in 2000 (Lewison et al . 2004). Further, a more recent paper used the best available information to es timate that Spain, Morocco, and Italy have the highest level of sea turtle bycat ch, with over 10,000 turtle captures per year for each country, and Greece, Malta, Libya, and Tunisia each catch 1,0 00 to 3,000 turtles per year (Casale 2008). Available data suggest the annua l number of loggerhead sea turt le captures by all Mediterranean pelagic longline fisheries may be greater than 50,000 (Casale 2008). Note that these are not necessarily individual turtles, as the same sea turtle can be captured more than once.

Mortality estimates in the pelagic longline fishery at gear retrieval appear to be lower than in some other types of gear (e.g., set gillnet). A lthough limited to observation s of direct mortality at gear retrieval, Carreras et al . (2004) found mortality to be low (0-7.7%) in the longline fishery off the Balearic Islands, and Jribi et al . (2008) reported 0% direct mortality in the southern152 Tunisia surface longline fishery. These estimates are consistent with th ose found in other areas; direct mortality was estimated at 4.3% in Greece (n=23), 0% in Italy (n=214), and 2.6% in Spain (n=676) (Laurent et al . 2001). However, considering injure d turtles and those released with hooks, the potential for mortality is likely much higher. Based upon observations of hooked loggerhead turtles in captivity, Aguilar et al . (1995) estimated 20-30% of animals caught in longline gear may eventually die. More recently, Casale et al . (2008c) found, given variations in hook position affecting survivability, the mortality rate of turtles caught by pelagic longlines may be higher than 30%, which is greater th an previously thought (17-42%; Lewison et al . 2004). Considering direct and post-relea se mortality, Casale (2008) used a conservative approach to arrive at 40% for the average mortality from Me diterranean pelagic longlines. The result is an estimated 20,000 turtles killed per ye ar by pelagic longlines (Casale 2008).

These incidental capture levels may impact th e Mediterranean Sea DPS. Of loggerheads caught in the summer pelagic longline fishery from the western and eastern Med iterranean, more than half originated from the Mediterranean nesti ng population (as opposed to all animals in the western Mediterranean originating from the Atlantic as previously believed; Laurent et al . 1998). In general, most of the turtles captured in the Mediterranean surface longline fisheries are juvenile animals (Aguilar et al . 1995, Panou et al . 1999, Camiñas et al . 2003, Casale et al

. 2007a, Jribi et al . 2008), but some adult loggerhead bycatch is also reported. Considering data from many Mediterranean areas and research st udies, the average size of turtles caught by pelagic longlines was 48.9 cm CCL (range 20.5-79.2 cm CCL; n=1868) (Casale 2008). Specifically, in the Spanish surface longline fish ery, 13% of estimated carapace sizes (n=455) ranged from 75.36 to 107 cm CCL, consider ed to be adult animals (Camiñas et al . 2003), and in the Ionian Sea, 15% of a total 157 loggerhead tur tles captured in swordfish longlines were adult animals (estimated size ≥ 75 cm) (Panou et al . 1999).

Bottom longlines are also fished in the Mediterra nean, but specific capture rates for loggerheads are largely unknown for many areas. The countri es with the highest number of documented captures (in the thousands per year) are Tunisi a, Libya, Greece, Turkey, Egypt, Morocco, and Italy (Casale 2008). Available data suggest the annual number of loggerhead sea turtle captures (not necessarily individu al turtles) by all Mediterranean demersal longliners may be greater than 35,000 (Casale 2008).

Mortality from bottom longlines occurs if the hooked animal cannot reach the surface to breathe or if the ingested hooks and branch lines result in injuries and post-re lease mortality. For example, in the Gulf of Gabès off Tunisia, th e estimated total capture from bottom longlines was 733 loggerheads, with a direct and potential mort ality (assuming all comatose turtles would die) of 12.5 to 33%, respectively (Jribi et al . 2008). Given available information and using a conservative approach, mortality fr o m bottom longlines may be at least equal to pelagic longline mortality (40%; Casale 2008). Th e result is an estimated 14,000 turtles killed per year in Mediterranean bottom longlines (C asale 2008). It is likely that these animals represent mostly juvenile, but some adult, loggerheads. Casale (2 008) reported an average turtle size of 51.8 cm CCL (n=35) in bottom longlines based on av ailable data throughout the Mediterranean.

Artisanal longline fisheries also have the potential to take sea turtles. A survey of 54 small boat (4-10 meter length) artisanal fishermen in Cyprus and Turkey resulted in an estimated minimum153 bycatch of over 2,000 turtles per year, with an estimated 10% mo rtality rate (Godley et al

. 1998a). These small boats fished with a comb ination of longlines and trammel/gillnets. However, note that it is likely that a proportion (perhaps a large proportion) of the turtle bycatch estimated in this study are green turtles.

Set net (gillnet) fisheries As in other areas, sea turtles have the potential to interact with set nets (gillnets or trammel nets) in the Mediterranean. Mediterranean set nets refer to gillnets (a single layer of net) and trammel nets, which consist of three layers of net with di fferent mesh size. Casale (2008) estimated that the countries with the highest number of logge rhead captures (in the t housands per year) are Tunisia, Libya, Greece, Turkey, Cyprus, and Croa tia. Italy, Morocco, Egypt, and France likely have high capture rates as well. Available information suggests the annual number of loggerhead captures by Mediterranean set nets may be greater than 30,000 (Casale 2008).

Due to the nature of the gear and fishing practic es (e.g., relatively long soak times), incidental capture in gillnets is among the highest source of direct sea turtle mortality. An evaluation of turtles tagged then re captured in gillnets along the Ita lian coast found 14 of 19 loggerheads (73.7%) to be dead (Argano et al . 1992). Gillnets off France were observed to capture six loggerheads with a 50% mortalit y rate (Laurent 1991). Six loggerheads were recovered in gillnets off Croatia between 1993 and 1996; 83% were found dead (Lazar et al . 2000). Off the Balearic Islands, 196 sea turtles were estimated to be captured in lobster trammel nets in 2001, with a CPUE of 0.17 turtles per vessel (Carreras et al . 2004). Mortality estimates for this artisanal lobster trammel net fishery ranged from 78 to 100%. Given this mortality rate and the number of turtles reported in lobster trammel nets, Carreras et al . (2004) estimate that a few thousand loggerhead turtles are k illed annually by lobster trammel nets in the whole western Mediterranean. Considering data throughout the entire Mediterran ean, as well as a conservative approach, Casale (2008) considered mortality by set nets to be 60%, with a resulting estimate of 16,000 turtles killed per year. Most of these animals are likel y juveniles; Casale (2008)

evaluated available set net catc h data throughout the Mediterranean and found an average size of 45.4 cm CCL (n=74).

As noted above, artisanal set net fisheries also ma y capture numerous sea turtles, as observed off Cyprus and Turkey (Godley et al . 1998a).

Driftnet fisheries Historically, driftnet fishing in the Mediterranean ca ught large numbers of sea turtles. An estimated 16,000 turtles were cap tured annually in the Ionian S ea driftnet fishery in the 1980s (De Metrio and Megalofonou 1988). The United Na tions established a worldwide moratorium on driftnet fishing effective in 1992, but unregul ated driftnetting conti nued to occur in the Mediterranean. For instance, a bycatch estimat e of 236 loggerhead turtles was developed for the Spanish swordfish driftnet fishery in 1994 (Silvani et al . 1999). While the Spanish fleet curtailed activity in 1994, the Moroccan, Turkish, French, and Italian driftnet fleets continued to operate.

Tudela et al . (2005) presented bycatch rates for driftnet fi sheries in the Alboran Sea and off Italy.

The Moroccan Alboran Sea driftn et fleet bycatch rate ranged from 0.21 to 0.78 loggerheads per haul, whereas the Italian driftnet fleet had a lo wer bycatch rate of 0.046 to 0.057 loggerheads per haul (Di Natale 1995, Camiñas 1997, Silvani et al . 1999). The use of driftnets in the154 Mediterranean continues to be illegal: the General Fisheries Commission for the Mediterranean prohibited driftnet fishing in 1997; a total ban on driftn et 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 dr iftnets in 2003. Neve rtheless, there are an estimated 600 illegal driftnet vessels operating in the Mediterranean, including fleets based in Algeria, France, Italy, Morocco, and Turkey (Environmental Justice Foundation 2007). In particular, the Moroccan fleet, operating in the Al boran Sea and Straits of Gibraltar, comprises the bulk of Mediterranean driftnetting, and has been found responsible for high bycatch, including loggerhead tur tles (Environmental Justi ce Foundation 2007, Aksissou et al . in press). Driftnet fishing in the Mediterranean, and accomp anying threats to loggerh ead turtles, continues

to occur.

Trawl fisheries

Sea turtles are known to be incide ntally captured in traw ls in Albania, Algeria, Croatia, Egypt, France, Greece, Italy, Libya, Morocco, Slovenia, Spain, Tunisia, and Turkey (Gerosa and Casale 1999, Camiñas 2004, Casale 2008). Laurent et al . (1996) estimated that approximately 10,000 to 15,000 sea turtles (most of which are loggerheads ) are captured by bottom tr awling in the entire Mediterranean. More recently, Casale (2008) compiled available trawl bycatch data throughout the Mediterranean and reported that Italy and Tunisia have the highest level of sea turtle bycatch, potentially over 20,000 captures per year comb ined, and Croatia, Greece, Turkey, Egypt, and Libya each catch more than 2,000 turtles per year . Further, Spain and Albania may each capture a few hundred sea turtles per year (Casale 2008). Available data suggest the annual number of sea turtle captures by all Mediterranean trawlers may be greater than 40,000 (Casale 2008). Note that these are capture events and not necessarily i ndividual turtles.

Information from specific study areas of the Me diterranean provides ad ditional insight into individual bycatch rates and details of the captures. Casale et al . (2004) estimated that 4,273 loggerheads per year are caught by the Italian bottom tr awl fishery in the we stern part of the northern Adriatic Sea, but as catch rates in the ea stern part of the Adriatic Sea are higher than in the western, it is likely that actual total catch in the entire North Ad riatic Sea is much higher. In southern Italy, off Lampedusa, Casale et al . (2007a) estimated an annua l total catch of 1,016 for the vessels in the study port. However, the cat ch estimate increased to approximately 4,000 to 11,000 turtles if trawlers from other ports fishing in the study area were considered. In the Gulf of Gabès, Tunisia, 5,458 loggerh ead turtles per year were estimat ed to be captured by the trawl fishery (Jribi et al . 2007), while in Croatia, an annual in cidental catch of 2,500 turtles was previously estimated based on fishermen inte rviews (Lazar and Tvrtkovic 1995). In Greece, total catch (presented as 95% c onfidence intervals) in 2000 was estim ated at 0 to 418 turtles in the northern Aegean Sea and at 0 to 4 48 turtles in the Ionian Sea (Laurent et al . 2001).

Although juveniles are inci dentally captured in trawl gear in many areas of the Mediterranean (Casale et al . 2004, 2007a; Jribi et al . 2007), adult turtles are also found. In Egypt, 25% of loggerheads captured in botto m trawl gear (n=16) were ≥ 70 cm CCL, and in Tunisia, 26.2% (n=62) were of this larger size class (Laurent et al . 1996). Off Lampedusa Island, Italy, the average size of tur tles caught by bottom trawlers was 51.8 cm CCL (range 22-87 cm CCL; n=368), and approximately 10% of the animals measured greater than 75 cm CCL (Casale et al

. 2007a). For all areas of the Mediterranean, Casale (2008) reported that medium to large turtles are generally caught by bottom trawl gear (mean 53.9 cm CCL; range 22-87 cm CCL; n=648).155 While there is a notable interaction rate in the Mediterranean, it appears that the mortality associated with trawling is relatively low. Inci dents of mortality have ranged from 3.3% (n=60) in Tunisia (Jribi et al . 2007) and 3.3% (n=92) in France (Lau rent 1991) to 9.4% (n=32) in Italy (Casale et al . 2004). Casale et al . (2004) found that mortality w ould be higher if all comatose

turtles were assumed to die. It also should be noted that the mo rtality rate in trawls depends on the duration of the haul, with l onger haul durations resulting in higher mortality rates (Henwood and Stuntz 1987, Sasso a nd Epperly 2006). Jribi et al . (2007) stated that the low recorded mortality in the Gulf of Gabès is likely due to the short haul durations in this area. Based on available information from multiple areas of the Mediterranean, and assuming that comatose animals die if released in that condition, the overall average mortality rate for bottom trawlers was estimated to be 20% (Casale 2008). This re sults in at least 7,400 turtles killed per year by bottom trawlers in all of the Mediterranean, but the number is likely more than 10,000 (Casale 2008).

Bottom trawl fisheries are likely impacting the Me diterranean Sea DPS. Although in a previous assessment of bottom trawl fisheries sampled fr om Tunisia, Egypt, and Turkey, all loggerheads captured were from the Mediterranean nesti ng population (as opposed to Atlantic nesting animals; Laurent et al . 1998), a recent genetic study with a larger sample size found a minor but significant contribution from the Atla ntic populations in the turtles fo raging in the neritic habitats off Tunisia (Casale et al . 2008b).

Mid-water trawling may have less total impact on sea turtles found in the Mediterranean than some other gear types, but interactions still occur. Casale et al . (2004) found that while no turtles were caught on observed mid-water trawl tr ips in the North Adriatic Sea, vessel captains reported 13 sea turtles captured from April to September. Considering total fishing effort, these reports resulted in a minimum total catch estimat e of 161 turtles/year in the Italian mid-water trawl fishery. Off Turkey, 71 loggerheads were captured in mid-water trawls from 1995-1996, while 43 loggerheads were incidentally taken in bottom trawls (Oruç 2001). In this same study, of a total 320 turtles captured in mid-water trawls (loggerheads and greens combined), 95% were captured alive and apparently healthy. While the total catch numbers throughout the Mediterranean have not been estimated, mid- water trawl fisheries do present a threat to loggerhead sea turtles.

Other gear types Seine, dredge, trap/pot, and hook and line fisheries operate in Mediterranean waters and may affect loggerhead turtles, alt hough incidental captures in these gear types are largely unknown (Camiñas 2004). Artisanal fisheries using a variet y of gear types also have the potential for sea turtle takes, but the effects of most artisanal ge ar types on sea turtles ha ve not been estimated.

Other Manmade and Natural Impacts Other anthropogenic threat s, such as interactions with recreational and commercial vessels, marine pollution, and intentional killing, also impact loggerheads found in the Mediterranean. Propeller and collision injuries from boats and ships are becomi ng more common in sea turtles, although it is unclear as to whether the events are increasing or just the reporting of the injuries. Speedboat impacts are of particular concern in ar eas of intense tourist activity, such as Greece and Turkey. Losses of nesting females from ve ssel collisions have been documented in Zakynthos and Crete in Greece (Camiñas 2004). In the Gulf of Naples, 28.1% of loggerheads156 recovered from 1993-1996 had injuries attributed to boat strikes (Bentivegna and Paglialonga 1998). Along the Greece coastline from 1997-1999, boa t strikes were reported as a seasonal phenomenon in stranded turtles (Kopsida et al . 2002), but numbers we re not presented.

Direct or indirect disposal of anthropogenic debris introduces potentially lethal materials into loggerhead foraging habitats. Unattended or disc arded nets, floating plastics and bags, and tar balls are of particular concern (Camiñas 2004, Margaritoulis 2007). Monofilament netting appears to be the most dangerous waste produced by the fishing industry (Camiñas 2004). In the Mediterranean, 20 out of 99 loggerhead turtles examined from Maltese fisheries were found contaminated with plastic or metal litter and hydrocarbons, with crude oil being the most common pollutant (Gramentz 1988). Of 54 juven ile loggerhead turtles incidentally caught by fisheries in Spanish Mediterranean waters, 79.6% had debris in their digestive tracts (Tomás et
al . 2002). In this study, plastics were the most fr equent type of marine debris observed (75.9%), followed by tar (25.9%). However, an examinati on of stranded sea turtle s in Northern Cyprus and Turkey found that only 3 of 98 animals were affected by marine debris (Godley et al

. 1998b).

Pollutant waste in the marine environment may im pact loggerheads, likely more than other sea turtle species. Omnivorous loggerheads stra nded in Cyprus, Greece, and Scotland had the highest organochlorine contaminant concentrations, as compared to green and leatherback turtles (Mckenzie et al . 1999). In northern Cyprus, Godley et al . (1999) found heavy metal concentrations (mercury, cadmium, and lead) to be higher in loggerheads than green turtles. Even so, concentrations of contaminants from se a turtles in Mediterranean waters were found to be comparable to other areas, generally with le vels lower than concentrations shown to cause

deleterious effects in other species (Godley et al . 1999, Mckenzie et al . 1999). However, lead concentrations in some Mediterranean loggerhead hatchlings were at levels known to cause toxic effects in other vertebrate groups (Godley et al . 1999).

As in other areas of the world, intentional killing or injuring of sea turtles has been reported to occur in the Mediterranean. Of 524 strandings in Greece, it appeared that 23% had been intentionally killed or injured (Kopsida et al . 2002). While some turtle s incidentally captured are used for consumption (as noted in Section 5.2. 8.2 above), it has been reported that some

fishermen kill the sea turtles th ey catch for a variety of othe r reasons, including non-commercial use, hostility, prejudice, recovery of hooks, and ignorance (Laurent et al . 1996, Godley et al

. 1998a, Gerosa and Casale 1999, Casale 2008).

Natural environmental events also may affect loggerheads in the Mediterranean. Cyclonic storms that closely resemble tropical cyclones in satellite images occasionally form over the Mediterranean Sea (Emanuel 2005). While hur ricanes typically do not occur in the Mediterranean, researchers have suggested that climate change could trigger hurricane development in this area in the future (Gaertner et al . 2007). Any significant storm event that may develop could disrupt loggerhead nesting activ ity and hatchling produc tion, but the results are generally localized and rarely result in whole- scale losses over multiple nesting seasons.

Similar to other areas of the world (see Section 5.1. 5.), climate change and sea level rise have the potential to impact loggerheads in the Mediterranean. Over the long term, Mediterranean turtle157 populations could be threatened by the alteration of thermal sa nd characteristics (from global warming), resulting in the reduction or cessation of female hatchling production (Camiñas 2004). Further, a significant rise in s ea level would restrict loggerhead nesting habitat in the eastern

Mediterranean.

5.2.9. South Atlantic Ocean DPS

5.2.9.1. Present or Threatened Destruction, Mo dification, or Curtailment of Its Habitat or Range

Terrestrial Zone As described in Section 5.1.1., destruction and modification of loggerhead nesting habitat in the South Atlantic result from coastal development and construction, placement of erosion control

structures and other barriers to nesting, beachfront lighting, vehicu lar and pedestrian traffic, sand extraction, beach erosion, beach sand placement, beach pollution, removal of native vegetation, and planting of non-native vegetation (D'Amato and Marczwski 1993; Marcovaldi and Marcovaldi 1999; Naro-Maciel et al . 1999; Marcovaldi et al . 2002, 2005; Marcovaldi 2007).

The primary nesting areas for logge rheads in the South Atlantic ar e in the states of Sergipe, Bahia, Espírito Santo, and Ri o de Janeiro in Brazil (Marcova ldi and Marcovaldi 1999). These primary nesting areas are monitored by Projet o TAMAR, the national s ea turtle conservation program in Brazil. Since 1980, Projeto TAMAR has worked to establish legal protection for nesting beaches (Marcovaldi and Marcovaldi 1999) . As such, human activities, including sand extraction, beach nourishment, seaw all construction, beach driving, and artificial lighting, that can negatively impact sea turtle nesting habitat, as well as directly impact nesting turtles and their eggs and hatchlings duri ng the reproductive season, are re stricted by various state and federal laws (Marcovaldi a nd Marcovaldi 1999; Marcovaldi et al . 2002, 2005). Nevertheless, tourism development in coastal areas in Brazi l is high, and Projeto TAMAR works toward raising awareness of turtles a nd their conservation n eeds through educational and informational

activities at their Vis itor Centers that are di spersed throughout the nes ting areas (Marcovaldi et
al . 2005).

In terms of non-native vegetation, the majority of nesting beaches in northern Bahia, where loggerhead nesting density is highest in Brazi l (Marcovaldi and Chaloupka 2007), have coconut plantations dating back to the 17 th century backing them (Naro-Maciel et al . 1999). It is impossible to assess whether this structured ha bitat has resulted in long-term changes to the loggerhead nesting rooker y in northern Bahia.

Neritic/Oceanic Zone As described in Section 5.1.1., human activities that impact bottom habitat in the loggerhead neritic and oceanic zones include fishing practic es, channel dredging, sand extraction, marine pollution, and climate change. These threats also occur in the South Atla ntic Ocean (e.g., Ibe 1996, Silva et al . 1997). General human activities have a ltered ocean ecosystems, as identified by ecosystem models ( http://www.lme.noaa.gov ). On the western side of the South Atlantic, the Brazil Current Large Marine Ecosystem (L ME) region is characterized by the Global International Waters Assessment (GIWA) as suffer ing severe impacts in the areas of pollution,158 coastal habitat modification, and overe xploitation of fish stocks (Marques et al . 2004). The Patagonian Shelf LME is moderately affected by pollution, habitat modification, and overfishing (Mugetti et al . 2004). On the eastern side of the Sout h Atlantic, the Benguela Current LME has been characterized as moderately impacted in the area of overfishing, with future conditions expected to worsen by the GIWA (Prochazka et al . 2005). Climate change also may result in future trophic changes, thus impacting logge rhead prey abundance and/or distribution.

5.2.9.2. Overutilization for Commercial, Recreati onal, Scientific, or Educational Purposes

Deliberate hunting of loggerheads for their meat, shells, and eggs is reduced from previous exploitation levels, but still exists. Limited numb ers of eggs are taken for human consumption in Brazil, but the relative amount is considered minor when compared to historical rates of egg collection (Marcovaldi and Marcovaldi 1999, Marcovaldi et al . 2005, Almeida and Mendes 2007). Use of sea turtles including loggerheads for medicinal purposes occasionally occurs in

northeastern Brazil (Alves and Rosa 2006). Use of byca ught loggerheads for subsistence and medicinal purposes is likely to occur in southe rn Atlantic Africa, based on information from central West Africa (Fretey 2001, Fretey et al . 2007).

5.2.9.3. Disease or Predation

The potential effects of diseases and endoparasites, as described in Section 5.1.3., also exist for loggerheads found in the South A tlantic Ocean. There have been five confirmed cases of fibropapillomatosis in loggerheads in Brazil (Baptisto tte 2007). There is no indication that this disease poses a major threat for this speci es in the eastern South Atlantic (Formia et al . 2007).

Eggs and nests in Brazil experience depredati on, primarily by foxes (Marcovaldi and Laurent 1996). Nests laid by loggerheads in the southern Atlantic African coas tline, if any, likely experience similar predation pressures to those on nests of other species laid in the same area (e.g., jackals depredate green turtle nests in Angola; Weir et al . 2007).

Loggerheads in the South Atlantic also may be impacted by harmfu l algal blooms (Gilbert et al

. 2005).

5.2.9.4. Inadequacy of Existing Regulatory Mechanisms

Unless identified to a specific geographic area, the regulatory mechanisms described in Section 5.1.4. also apply to sea turtles found in the S outh Atlantic Ocean. The following additional instruments also apply to loggerheads found in this area.

Convention on the Conservation of Migratory Species of Wild Animals, Memorandum of Understanding Concerning Conservation Measures for Marine Turtles of the Atlantic Coast of Africa (Abidjan Memorandum) For information about this instrument, see section 5.2.7.4. for the Northeast Atlantic Ocean DPS.159 The Convention for the Co-operation in the Protection and Development of the Marine and Coastal Environment of the West and Ce ntral African Region (Abidjan Convention) For information about this instrument, see section 5.2.7.4. for the Northeast Atlantic Ocean DPS.

South-East Atlantic Fisher ies Organization (SEAFO) SEAFO manages fisheries activities in the Southeas t Atlantic high seas ar ea, excluding tunas and billfish. SEAFO adopted Resolution 01/06, “t o 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 interactions in SEAFO-managed fisheries. This Resolution is not legally binding. Additional information is

available at http://www.seafo.org

.

Inter-American Convention for the Protecti on and Conservation of Sea Turtles (IAC) For information about this instrument, see sec tion 5.2.1.4. for the North Pacific Ocean DPS.

Nairobi Convention for the Protection, M anagement and Developmen t of the Marine and Coastal Environment of th e Eastern African Region For information about this instrument, see sec tion 5.2.3.4. for the North Indian Ocean DPS.

5.2.9.5. Other Natural or Manmade Factors Affecting Its Continued Existence

Incidental Bycatch in Fishing Gear Incidental capture of sea turtles in artisanal and commercial fisher ies is a significant threat to the survivability of loggerheads in the South Atlantic. Sea turtle s may be caught in pelagic and demersal longlines, drift and set gillnets, bottom and mid-water trawling, fishing dredges, pound nets and weirs, haul and purse seines, pots a nd traps, and hook and line gear. In the western South Atlantic, there are various efforts aimed at mitigating bycatch of sea turtles in various fisheries. In Brazil, there is the National Action Plan to Reduce Incidental Capture of Sea

Turtles in Fisheries , coordinated by Projeto TAMAR (Marcovaldi et al . 2006). This action plan focuses on both artisanal and commercial fisheries, a nd collects data directly from fishers as well as on-board observers. Although lo ggerheads have been observed as bycatch in all fishing gear and methods identified above, Marcovaldi et al . (2006) have identified longlining as the major source of incidental capture of loggerhead turtles. Reports of loggerhead bycatch by pelagic longlines come mostly from th e southern portion of the Brazilian Exclusive Economic Zone,

between 20 o S and 35 o S. Bugoni et al . (2008) reported a loggerhead bycatch rate of 0.52 juvenile turtles/1000 hooks by surface longlines targeting dolphinfish. Pinedo et al . (2004) reported seasonal variation in bycatch of juvenile logge rheads (and other sea turtle species) by pelagic longlines in the same region of Brazil, with the highest rates (1.85 turtles/1000 hooks) in the austral spring. Kotas et al . (2004) reported the highest rate s of loggerhead bycatch (>10 turtles/1000 hooks) by pelagic longlines in the au stral summer/fall months. A study based on several years found that the highe st rate of loggerhead bycatch in pelagic longlines off of Uruguay and Brazil was in the late austral su mmer month of February: 2.72 turtles/1000 hooks (López-Medilaharsu et al . 2007). Sales et al . (2004) reported a logge rhead bycatch rate of 0.87/1000 hooks near the Rio Grande Elevação do Rio Grande, about 600 nm off the coast of southern Brazil. In Uruguayan waters, the prim ary fisheries with logge rhead bycatch are bottom trawlers and longlines (Domingo et al . 2006). Domingo et al . (2008) reported bycatch rates of160 loggerheads of 0.9-1.3/1000 hooks by longline deployed south of 30°S. In waters off Argentina, bottom trawlers also catch some loggerheads (Domingo et al . 2006).

In the eastern South Atlantic, s ea turtle bycatch in fisheries ha s been documented from Gabon to South Africa (Fretey 2001). Limited data are available on bycatch of loggerheads in coastal fisheries, although loggerheads are k nown (or strongly suspected) to occur in coastal waters from Gabon to South Africa (Fretey 2001, Bal et al . 2007, Weir et al . 2007). Coastal fisheries implicated in bycatch of loggerheads and other tu rtles include gillnets, beach seines, and trawlers (Bal et al . 2007).

In the high seas, longlines are used by fishing boats targeting tuna and swordfish in the eastern South Atlantic. A recent study by Honig et al . (2008) estimates 7,600-120,000 sea turtles are incidentally captured by commercial longlines fi shing in the Benguela Current LME; 60% of these are loggerheads. Petersen et al . 2007, 2009) report that the rate of loggerhead bycatch in South African longliners was around 0.02 turtle s/1000 hooks, largely in the Benguela Current LME. In the middle of the South Atlantic, l oggerhead bycatch by longlines was reported to be low, relative to other regions in the Atlantic (Mejuto et al 2008).

Other Manmade and Natural Impacts As described in Section 5.1.5., other anthropogenic impacts, such as boat strikes and ingestion or entanglement in marine debris, also apply to loggerheads in the S outh Atlantic. Bugoni et al

. (2001) have suggested the ingestion of plastic an d oil may contribute to loggerhead mortality on the southern coast of Brazil. Plastic marine debris in the eastern South Atlantic also may pose a problem for loggerheads and other sea turtles (R yan 1996). Similar to other areas of the world (see Section 5.1.5.), climate change and sea level rise have the pot ential to impact loggerheads in

the South Atlantic.

Oil reserve exploration and extraction activities also may pose a threat for sea turtles in the South Atlantic. Seismic surveys in Br azil and Angola have recorded s ea turtle occurrences near the seismic work (Gurjão et al . 2005, Weir et al . 2007). While no sea turtle takes were directly observed on these surveys, increased equipment and pr esence in the water that is associated with

these activities also increases the like lihood of sea turtle interactions (Weir et al . 2007).

Natural environmental events may affect loggerh eads in the South Atlantic. However, while a rare hurricane hit Brazil in March 2004, typica lly hurricanes do not occur in the South Atlantic (McTaggart-Cowan et al . 2006). This is generally due to hi gher windspeeds aloft, preventing the storms from gaining heig ht and therefore strength.161 

## SECTION 6—CONCLUSIONS—SYNTHESIS & INTEGRATION

6.1. North Pacific Ocean DPS

We used two approaches to compute extinction risk for the North Pacific Ocean DPS: (1) computation of susceptibility to quasi-extincti on (SQE), and (2) a stag e-based deterministic model to determine negative effect s of known threats to the DPS. Based on nest count data for nearly the past two decades, the North Pacific population of loggerheads is small. The SQE approach suggested that the North Pacific Ocean DPS appears to be declining, is at risk, and is thus likely to decline in the foreseeable future . These results are based on nesting beach census data from 33 nesting beaches in Japan over a 17- year period (in Snover 2008). We note that the SQE approach we used is based on past performa nce of the DPS (nesting da ta) and does not fully reflect ongoing and future threats to all life stages within the DPS. The stage-based deterministic modeling approach used a hypothe tical population that was at its maximum population growth rate and computed the population growth potential under known or suspected threats to different life stages of the North Pacific Ocean DPS. This approach provided a range of results: in the case of the lowest anthropogenic mortality rates (or the best case scenar io), the deterministic model suggests that the North P acific Ocean DPS would grow slightly, but in the worst-case scenario, the model indicates that the population w ould be likely to substa ntially decline in the foreseeable future. These results are largely driven by the mortality of juvenile and adult loggerheads from fishery bycatch that occurs throughout the North Pacific Ocean, including the coastal pound net fisheries off Japan, coastal fi sheries impacting juven ile foraging populations off Baja California, Mexico, and undescribed fish eries likely affecting loggerheads in the South China Sea and the North Pacific Ocean. Although national and internat ional governmental and non-governmental entities on both sides of the No rth Pacific are curren tly working toward reducing loggerhead bycatch, and some positive actions have been implemented, it is unlikely that this source of mortality can be sufficiently redu ced in the near future due to the challenges of mitigating illegal, unregulated, and unreported fish eries, the lack of comprehensive information on fishing distribution and effort, limitations on implementing demonstrated effective conservation measures, geopoli tical complexities, limitations on enforcement capacity, and availability of comprehensive bycatch reduction technologies. In addition to fishery bycatch, coastal development and coastal armoring on nesti ng beaches in Japan continues as a substantial threat. Coastal armoring, if left unaddressed, will become an even more substantial threat as sea level rises. Therefore, the BRT concluded that the North Pacific Ocean D PS is currently at risk

of extinction.



### 6.2. South Pacific Ocean DPS

We used two approaches to compute extinction risk for the South Pacific Ocean DPS: (1) computation of susceptibility to quasi-extincti on (SQE), and (2) a stag e-based deterministic model to determine negative effects of known thr eats to the DPS. The SQE approach suggested that, based on nest count data fo r the past three decades , the population is at risk and thus likely

to decline in the foreseeable future. These re sults are based on recently published nesting census data for loggerheads at index beaches in eas tern Australia (Limpus 2009). The stage-based deterministic modeling approach used a hypot hetical population that was at its maximum population growth rate and computed the populati on growth potential under known or suspected162 threats to different life stages of the South P acific Ocean DPS. This approach provided a wide range of results: in the case of the lowest anthropogenic mortality rates (or the best case scenario), the deterministic model suggests that the South Pacific Ocean DPS will grow slightly, but in the worst-case scenario, th e model indicates that the populat ion is likely to substantially

decline in the foreseeable future. These results are largely driven by mortality of juvenile and adult loggerheads from fishery bycatch that occurs throughout the S outh Pacific Ocean. Although national and inte rnational governmental and non-gove rnmental entities on both sides

of the South Pacific are curre ntly working toward reducing loggerhead bycatch, and some positive actions have been implemented, it is unlikely that this source of mortality can be sufficiently reduced in the near future due to the challenges of mitigating illegal, unregulated, and unreported fisheries, the cont inued expansion of artisanal fleet s in the southeastern Pacific, the lack of comprehensive information on fi shing distribution and effort, limitations on implementing demonstrated effective conser vation measures, geopolitical complexities, limitations on enforcement capacity, and avai lability of comprehensive bycatch reduction technologies. Therefore, the BRT concluded that the South Pacific Ocean DPS is currently at

risk of extinction.



### 6.3. North Indian Ocean DPS

Two approaches to computing extinction risk for the North Indian Ocean DPS were considered: (1) computation of susceptibility to quasi-extinction (SQE), and (2) a stage-based deterministic model to determine negative effect s of known threats to the DPS. The SQE approach is based on nesting data; however, an adequate time series of nesting data for this DPS was not available to the BRT. Therefore, we could not use this appro ach to evaluate extinction risk. The stage-based deterministic modeling approach used a hypot hetical population that was at its maximum population growth rate and computed the populati on growth potential under known or suspected threats to different life stages of the North I ndian Ocean DPS. The model indicated the North Indian Ocean DPS is likely to decline in the fo reseeable future. These results are driven by cumulative mortality from a variety of sources ac ross all life stages. Thr eats to nesting beaches are likely to increase, which would require addi tional and widespread nesting beach protection efforts. Little is currently being done to mon itor and reduce mortality from neritic and oceanic fisheries in the range of the North Indian Ocean DPS; this mortality is likely to continue and increase with expected additional fishing effort from commercial and artisanal fisheries. Reduction of mortality would be difficult due to a lack of comprehensive information on fishing distribution and effort, limitations on impl ementing demonstrated effective conservation measures, geopolitical complexities, limitations on enforcement capacity, and availability of comprehensive bycatch reduction technologies. Th erefore, the BRT concluded that the North Indian Ocean DPS is currently at risk of extinction.



### 6.4. Southeast Indo-Pacific Ocean DPS

Two approaches to computing extinction risk fo r the Southeast Indo-Pacific Ocean DPS were considered: (1) computation of susceptibility to quasi-extinction (SQE), and (2) a stage-based deterministic model to determine negative eff ects of known threats to the DPS. The SQE approach is based on nesting data; however, an adequate time series of nesting data for this DPS was not available to the BRT; therefore, we coul d not use this approach to evaluate extinction163 risk. The stage-based deterministic modeling approach used a hypothetical population that was at its maximum population growth rate and co mputed the population growth potential under known or suspected threats to different life stag es of the Southeast Indo-Pacific Ocean DPS. This approach provided a wide range of resu lts: in the case of the lowest anthropogenic mortality rates, the deterministic model suggest s that the South Pacifi c Ocean DPS will grow slightly, but in the worst-case scenario, the model indicates th at the population is likely to substantially decline in the foreseeable future. These results are largely driven by mortality of juvenile and adult loggerheads from fishery bycatch that occurs throughout the region, as can be

inferred from data from Aust ralia’s Pacific waters. Alt hough national and international governmental and non-governmental entities are currently working toward reducing loggerhead bycatch, and some positive actions have been impl emented, it is unlikely that this source of mortality can be sufficiently reduced in the near fu ture due to the challenges of mitigating illegal, unregulated, and unreported fisheries, the continue d expansion of artisanal fleets, the lack of comprehensive information on fishing distri bution and effort, limitations on implementing

demonstrated effective conservation measur es, geopolitical complexities, limitations on enforcement capacity, and availability of comprehensive bycatch reduction technologies. Therefore, the BRT concluded that the Southeast Indo-Pacific Ocean DPS is currently at risk of

extinction.



### 6.5. Southwest Indian Ocean DPS

Two approaches to computing extinction risk for the Southwest Indian Ocean DPS were considered: (1) computation of susceptibility to quasi-extinction (SQE), and (2) a stage-based deterministic model to determine negative effects of known threats to the DPS. These two approaches provided different, although not incompa tible, results for th e Southwest Indian Ocean DPS. The SQE approach, based on a 37-year time series of nesting female counts at Tongaland, South Africa (1963-1999), indicated this segment of th e population, while small, has increased, and the likelihood of quasi-extinction is negligible. We note that the SQE approach we used is based on past performance of the DPS (nesting data from 1963-1999) and does not fully reflect ongoing and future threats to all lif e stages within the DPS. The stage-based deterministic modeling approach used a hypot hetical population that was at its maximum population growth rate and computed the populati on growth potential under known or suspected threats to different life stages of the Southwes t Indian Ocean DPS. This approach provided a wide range of results: in the case of the lowest anthropogenic mortality rates, the deterministic model suggests that the South Pacific Ocean D PS will grow slightly, but in the worst-case scenario, the model indicates that the populati on is likely to substant ially decline in the foreseeable future. These results are largely dr iven by mortality of juvenile loggerheads from fishery bycatch that occurs throughout the Southwes t Indian Ocean. This mortality is likely to continue and may increase with expected additiona l fishing effort from co mmercial and artisanal

fisheries. Reduction of mortalit y would be difficult due to a lack of comprehensive information on fishing distribution and effort, limitations on implementing demonstrated effective conservation measures, geopoli tical complexities, limitations on enforcement capacity, and availability of comprehensive bycatch reduction technologies. Therefore, the BRT concluded that although the Southwest Indian Ocean DPS is likely not currently at immediate risk of extinction, the extinction risk is likely to increase in the foreseeable future.164 

### 6.6. Northwest Atlantic Ocean DPS

We used two approaches to compute extinction ri sk for the Northwest Atlantic Ocean DPS: (1) computation of susceptibility to quasi-extincti on (SQE), and (2) a stag e-based deterministic model to determine negative effect s of known threats to the DPS. The SQE approach suggested that, based on nest count data for the past two decades, the populati on is at risk and thus likely to

decline in the foreseeable future. These results are based on nesting data for loggerheads at index/standardized nesting survey beaches in the USA and the Yucatan Peninsula, Mexico (NMFS and FWS 2008). The stage-based determin istic modeling approach used a hypothetical population that was at its maxi mum population growth rate and computed the population growth potential under known or suspected threats to di fferent life stages of the Northwest Atlantic Ocean DPS. This approach indicated the Northw est Atlantic Ocean DPS is likely to decline in the foreseeable future, even under the scenario of the lowest anthropogenic mortality rates. These results are largely driven by mortality of juvenile and adult loggerheads from fishery bycatch that occurs throughout the North Atla ntic Ocean. Although national and international governmental and non-governmental entities on both sides of the North Atlantic are currently working toward reducing loggerhead bycatch, and some positive actions have been implemented, it is unlikely that this source of mortality can be sufficiently reduced across the range of the DPS

in the near future because of th e diversity and magnitude of the fisheries operating in the North Atlantic, the lack of comprehensive informati on on fishing distribution and effort, limitations on implementing demonstrated effective conser vation measures, geopolitical complexities, limitations on enforcement capacity, and avai lability of comprehensive bycatch reduction technologies. Therefore, the BRT concluded that the Northwest Atlantic Ocean DPS is currently

at risk of extinction.



### 6.7. Northeast Atlantic Ocean DPS

Two approaches to computing extinction risk for the Northeast Atlantic Ocean DPS were considered: (1) computation of susceptibility to quasi-extinction (SQE), and (2) a stage-based deterministic model to determine negative eff ects of known threats to the DPS. The SQE approach is based on nesting data. However, we had insufficient nest count data over an appropriate time series for this DPS and could not use this approach to evaluate extinction risk. The stage-based deterministic modeling approach used a hypothetical population that was at its maximum population growth rate and computed the population growth potential under known or suspected threats to different life stages of the Northeast Atlantic Ocean DPS. The model

indicated the Northeast Atlantic Ocean DPS is likely to decline in the foreseeable future, even under the scenario of the lowest anthropogenic mort ality rates. These results are largely driven by the ongoing directed lethal take of nesting fe males, low hatching and emergence success, and mortality of juveniles and adults from fisher y bycatch that occurs throughout the Northeast Atlantic Ocean. Currently, conservation efforts to protect nesting females are growing, and a reduction in this source of mortality is likely to continue in the near future. Although national and international governmental and non-governmental entities in the Northeast Atlantic are currently working toward reducing loggerhead bycatch, and some positive actions have been implemented, it is unlikely that this source of mortality can be sufficiently reduced across the range of the DPS in the near future because of th e lack of bycatch reductio n in high seas fisheries operating within the range of this DPS, lack of bycatch reduction in coastal fisheries in Africa,165 the lack of comprehensive information on fi shing distribution and effort, limitations on implementing demonstrated effective conser vation measures, geopolitical complexities, limitations on enforcement capacity, and avai lability of comprehensive bycatch reduction technologies. Therefore, the BRT concluded that the Northeast Atlantic Ocean DPS is at

immediate risk of extinction.



### 6.8. Mediterranean Sea DPS

Two approaches to computing extinction risk fo r the Mediterranean Sea DPS were considered: (1) computation of susceptibility to quasi-extinction (SQE), and (2) a stage-based deterministic model to determine negative effect s of known threats to the DPS. The SQE approach is based on nesting data; however, region-wide nesting data for this DPS were not available to the BRT. Therefore, we could not use this approach to evaluate extinction ri sk. The stage-based deterministic modeling approach used a hypot hetical population that was at its maximum population growth rate and computed the populati on growth potential under known or suspected threats to different life stages of the Medi terranean Sea DPS. The model indicated the Mediterranean Sea DPS is likely to decline in th e foreseeable future, even under the scenario of the lowest anthropogenic mortality rates. Thes e results are largely dr iven by mortality of juvenile and adult loggerheads from fishery byc atch that occurs thr oughout the Mediterranean Sea, as well as anthropogenic threats to ne sting beaches and eggs/hatchlings. Although conservation efforts to protect some nesting beaches are underway, more widespread and consistent protection is needed. Although na tional and international governmental and nongovernmental entities in the Mediterranean Sea are currently working toward reducing loggerhead bycatch, it is unlikely th at this source of mortality can be sufficiently reduced across the range of the DPS in the near future because of the lack of bycatch reduction in commercial and artisanal fisheries operating within the ra nge of this DPS, the lack of comprehensive information on fishing distribution and effo rt, limitations on implementing demonstrated

effective conservation measures , geopolitical complexities, lim itations on enforcement capacity, and availability of comprehensive bycatch reduction technologies. Therefore, the BRT

concluded that the Mediterranean Sea DPS is at immediate risk of extinction.



### 6.9. South Atlantic Ocean DPS

We used two approaches to compute extinction risk for the South Atla ntic Ocean DPS: (1) computation of susceptibility to quasi-extincti on (SQE), and (2) a stag e-based deterministic model to determine negative effects of known th reats to the DPS. These two approaches provided different, although not incompatible, resu lts for the South Atlantic Ocean DPS. The SQE approach suggested that, based on nest coun t data for the past two decades, the population is unlikely to decline in the foreseeable future. These results are consistent with Marcovaldi and Chaloupka’s (2007) nesting beach trend analyses. We note that the SQE approach is based on past performance of the DPS (n esting data) and does not fully re flect ongoing and future threats to all life stages within the DPS. The stag e-based deterministic modeling approach used a hypothetical population that was at its maximu m population growth rate and computed the population growth potential under know n or suspected threats to diffe rent life stages of the South Atlantic Ocean DPS. This approach indicated th e South Atlantic Ocean DPS is likely to decline in the foreseeable future, even under the scenario of the lowest anthropoge nic mortality rates.166 This result is largely driven by mortality of juve nile loggerheads from fish ery bycatch that occurs throughout the South Atlantic Oc ean. Although national and in ternational governmental and non-governmental entities on both sides of the No rth Atlantic are currently working toward reducing loggerhead bycatch in the South Atlantic, it is unlikely that this source of mortality can be sufficiently reduced across the range of the D PS in the near future because of the diversity and magnitude of the commercial and artisanal fish eries operating in the S outh Atlantic, the lack of comprehensive information on fishing distri bution and effort, limitations on implementing

demonstrated effective conservation measur es, geopolitical complexities, limitations on enforcement capacity, and availability of comprehensive bycatch reduction technologies. Therefore, the BRT concluded that although the S outh Atlantic Ocean DPS is not currently at immediate risk of extinction, the extinction risk is likely to increase substantially in the

foreseeable future.167 

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