NOAA Technical Memorandum NMFS-SEFSC-444

ASSESSMENT UPDATE FOR THE KEMP’S RIDLEY AND

LOGGERHEAD SEA TURTLE POPULATIONS IN THE WESTERN

NORTH ATLANTIC

A Report of the Turtle Expert Working Group

November 2000

U. S. Department of Commerce

National Oceanic and Atmospheric Administration

National Marine Fisheries Service

Southeast Fisheries Science Center

75 Virginia Beach Drive

Miami, FL 33149
NOAA Technical Memorandum NMFS-SEFSC-444

ASSESSMENT UPDATE FOR THE KEMP’S RIDLEY AND

LOGGERHEAD SEA TURTLE POPULATIONS IN THE WESTERN

NORTH ATLANTIC

A Report of the Turtle Expert Working Group

U. S. DEPARTMENT OF COMMERCE

Norman Y. Mineta, Secretary

NATIONAL OCEANIC AND ATMOSPHERIC ADMINISTRATION

D. James Baker, Administrator

NATIONAL MARINE FISHERIES SERVICE

Penelope H. Dalton, Assistant Administrator for Fisheries

November 2000

Unpublished reports are used for documentation and timely communication of preliminary

results, interim reports, or special-purpose information, and have not received complete

formal review, editorial control, or detailed editing.
ii

NOTICE

The National Marine Fisheries Service (NMFS) does not approve, recommend or endorse any

proprietary product or material mentioned in this publication. No reference shall be made

to NMFS, or to this publication furnished by NMFS, in any advertising or sales promotion

which would indicate or imply that NMFS approves, recommends or endorses any proprietary

product or proprietary material herein or which has as its purpose any intent to cause directly

or indirectly the advertised product to be used or purchased because of NMFS publication.

Correct citation of this report is:

Turtle Expert Working Group. 2000. Assessment Update for the Kemp’s Ridley and Loggerhead

Sea Turtle Populations in the Western North Atlantic. U.S. Dep. Commer. NOAA Tech.

Mem. NMFS-SEFSC-444, 115 pp.

Copies of this report can be obtained from:

National Marine Fisheries Service

Southeast Fisheries Science Center

75 Virginia Beach Drive

Miami, FL 33149

(305) 361-4285

or

National Technical Information Service

5285 Port Royal Road

Springfield, VA 22161

(703) 605-6000

(800) 553-6847 (rush orders)
NOAA Tech. Mem. NMFS-SEFSC-444 iii

Table of Contents

List of Tables .................................................... vi

List of Figures .................................................. viii

Preface ........................................................... x

Executive Summary ............................................... xii

Turtle Expert Working Group Members ............................ xiv

Part 1. Kemp’s Ridley Status Report ................................. 1

Summary .............................................................. 1

History of Nesting at Rancho Nuevo ....................................... 1

New Nesting Information ................................................ 2

México – Update of nest numbers 1966-1998 ........................... 2

Nesting records from South Texas .................................... 2

Life Stage Definitions and Geographic Locations ...................... 3

Post-hatchling stage ............................................... 3

Benthic immature stage (20-60 cm) ................................... 3

Adult stage ....................................................... 3

Status and Trends ...................................................... 4

Nesting data ...................................................... 4

Nests vs. hatchlings released 8, 10, and 12 years earlier ................... 4

Strandings ....................................................... 4

Strandings “hotspots” .............................................. 5

Changes in size distributions ........................................ 5

In-water surveys .................................................. 6

Conclusions from nesting and strandings data .......................... 8

Nest protection in México – Future strategies ........................... 9

Population size and trends from a deterministic age-based model ............... 9

Model modifications from the 1998 report ............................. 9

Model form

......................................................... 10

Parameterization ................................................. 11

Nesting data used .......................................... 11

Reproduction

................................................... 12

Sex ratio (all age classes)

................................................... 13

Age and growth (see Appendix 1) ............................. 13
iv NOAA Tech. Mem. NMFS-SEFSC-444

Benthic immature mortality from strandings catch curve estimates

(Zb) ................................................ 13

Model Results ............................................. 14

Model fits to observed nests ................................. 14

Sensitivity of model results to changes in small benthic immature Zb

................................................... 15

Model fits with an empirically-derived estimate of f ............. 15

Best-fit model projections ................................... 16

Literature cited ........................................................ 16

Part 2. Loggerhead Status Report ................................... 40

Summary ............................................................. 40

Stock Definition and Geographic Range ................................... 40

Life stage definitions .............................................. 42

Status and Trends ..................................................... 43

Nesting beaches .................................................. 43

Northern Subpopulation .................................... 43

South Florida Subpopulation ................................ 43

Florida Panhandle Subpopulation ............................ 44

Nest count correlations across states and regions ................ 44

Yucatán nesting subpopulation .............................. 44

Loggerhead nesting elsewhere in the Gulf of Mexico ............. 44

Strandings ...................................................... 45

Trends ................................................... 45

Strandings "hot spots" ..................................... 45

TED effect analyses ........................................ 46

In-water surveys ................................................. 46

Sex ratio ........................................................ 48

Size distributions ................................................. 48

Mean size of stranded loggerheads 1986-87 vs. 1996-97 .......... 48

Size distributions from the St. Lucie Power Plant ............... 49

Population size estimates .......................................... 49

Conclusions from nest numbers, strandings, and in-water surveys ......... 49

Literature Cited ....................................................... 51

Part 3. Anthropogenic Mortality Sources ............................. 71

Summary ............................................................. 71

Mortality Sources for Pelagic Stage ....................................... 71

Pelagic longline fisheries .......................................... 71

Marine debris ................................................... 72

Mortality Sources for Benthic Feeding Stages .............................. 72

Trawls ......................................................... 72
NOAA Tech. Mem. NMFS-SEFSC-444 v

Gill nets ........................................................ 73

Hook and line ................................................... 74

Power plants .................................................... 75

Marine debris ................................................... 75

Literature Cited ....................................................... 76

Part 4. Take Limits ............................................... 80

Introduction .......................................................... 80

Three Interim Strandings Limits (ISL) calculations: Methods and Limitations ... 81

1) Running average ............................................... 81

2) Slope method .................................................. 82

Slope calculation .......................................... 82

Setting the ISL ............................................ 82

Caveats of the slope method ................................. 82

3) Potential Biological Removal (PBR)

......................................................... 83

PBR calculations for sea turtles .............................. 83

Application to adult loggerheads

................................................... 84

Caveats of PBR ............................................ 84

Summary of results 1996 – 1998 .......................................... 84

Research recommendations for future evaluation of take limits ................ 85

Literature Cited ....................................................... 85

Part 5. Research Recommendations ................................. 92

Appendix 1. Von Bertalanffy Growth Models for Wild Kemp’s Ridley

Turtles: Analyses of the NMFS Miami Laboratory Tagging

Database ............................................. 94

Appendix 2. Sea Turtle Strandings ................................. 103
vi NOAA Tech. Mem. NMFS-SEFSC-444

List of Tables

Table 1. Nests counted at Rancho Nuevo, North Camp and South Camp 1966 - 1998.

Table 2. Small turtle strandings (20-30 cm SCL) and hatchlings released 1 - 4 years prior to

stranding year, 1986–1997.

Table 3. Kemp’s ridley strandings, 1986-1997.

Table 4. Kemp’s ridley strandings trends.

Table 5. Life stages defined in the age-based model for Kemp’s ridleys.

Table 6. Mortality rate parameters defined for the age-based model.

Table 7. Hypothetical changes in hatching rate expected with increasing numbers of nests, as

applied to density-dependent projections of the population model.

Table 8. Sum of squared residuals (observed - expected nests) for 3 models fit to various time

series.

Table 9. Parameters for model best-fits with small benthic immature z=0.5.

Table 10. Natal origin of foraging-ground loggerhead sea turtles. Table values are proportional

contributions.

Table 11. Loggerhead sea turtle nests in the U.S., 1989-1998.

Table 12. Mean annual percent change in loggerhead nest numbers in the U.S. 1989-1998, based

on slopes calculated from ln-transformed data. 

Table 13. Correlations (Pearson) of loggerhead nests from index beaches at Cape Island, Cape

Romain National Wildlife Refuge, S.C., and Hutchinson Island, Fla. with nest totals

(excluding each index beach) from each state and region, 1989-1998.

Table 14. Correlations (Pearson) of U. S. loggerhead nests from all states and regions, 1989-

1998.

Table 15. Loggerhead strandings by region.

Table 16. Loggerhead stranding trends. 
NOAA Tech. Mem. NMFS-SEFSC-444 vii

Table 17. Trends in loggerhead abundance at the St. Lucie Nuclear Power Plant, St. Lucie,

Florida (Quantum Resources, 1995, unpublished data), over various time intervals to

compare with nests and strandings.

Table 18. Observed size distributions for loggerheads.

Table 19. Pearson correlation coefficients between loggerhead nests and strandings in the

southeastern U. S. and Gulf of Mexico and loggerhead captures at the St. Lucie Power

Plant , 1989-1997.

Table 20. Total shrimping vessels and vessels operating in the Campeche and Tamaulipas fleets,

1982-1995.

Table 21. Hatchling production and strandings levels used to calculate slopes.

Table 22. PBR and ISL estimates for adult loggerheads, 1998.

Table 23. Summary of Incidental Stranding Limits (ISL) and total strandings for 1996-98.
viii NOAA Tech. Mem. NMFS-SEFSC-444

List of Figures

Figure 1. Nesting beaches of the Kemp’s ridley sea turtle. Primary nesting beach is Rancho

Nuevo.

Figure 2. Nests and hatchlings released from Rancho Nuevo, North Camp and South Camp,

1966-1999 (hatchlings for 1999 are approximate).

Figure 3. Nests counted at Rancho Nuevo, North Camp and South Camp 1989-1998.

Figure 4. Kemp's ridley nest projection if current trends continue.

Figure 5. Scatter plots of nests counted vs. hatchlings released 8, 10 or 12 years later.

Figure 6. Size frequency distributions of stranded benthic immature Kemp’s ridleys in 1986-

1987 and 1996-1997, pooled across all statistical zones (Appendix 2).

Figure 7. Best fit model estimates of nests for 10 years to maturity, Zb = 0.5, with and without

the post-1990 multiplier as a fitted parameter.

Figure 8. Model fits for 8, 10 or 12 years to maturity with post-1990 multiplier included.

Figure 9. Changes in best-fit parameter estimates for pelagic immature and large turtle

instantaneous mortality with changes in the small benthic immature mortality rate,

estimated from a catch curve of strandings.

Figure 10. Effect of small benthic immature mortality estimate on the post-1990 multiplier.

Figure 11. Model estimated population size in 1998, 10 years to maturity model with a range of

small benthic immature survival rates.

Figure 12. Model fits and residuals for pre-1990 small benthic immature Zb = 0.53, post-1990 Zb

= 0.34. A. Model with best-fit post-1990 multiplier for large benthic immature and

adult turtles f = 0.52. B. Post-1990 multiplier f fixed at 0.65 for all benthic

immatures and adults.

Figure 13. Model projections with Zb = 0.5, assuming mortality rates remain constant after 1990.

Figure 14. Model projection of nesting female numbers with and without a reduction in nest

survival with nest density (10 years to maturity, Zb = 0.5).

Figure 15. Loggerhead nesting areas, Western North Atlantic.
NOAA Tech. Mem. NMFS-SEFSC-444 ix

Figure 16. Loggerhead nests for the Northern Nesting Subpopulation, 1989-1998. A. Nests

counts by state. B. Long-term nesting records for Cape Island, S.C.

Figure 17. Loggerhead nests for the South Florida Nesting Subpopulation, 1989-1998. A. Nests

counts by region. B. Long-term nesting records for Hutchinson Island, Fla.

Figure 18. Loggerhead turtles captured by the cooling water intake pipe at the St. Lucie Nuclear

Power Plant, St. Lucie, Fla., 1977-1998.

Figure 19. Size distributions and cumulative frequecies for loggerhead turtle strandings. A.

Pooled data for 1986-1987. B. Pooled data for 1996-1997.

Figure 20. Size distributions of loggerheads captured at the St. Lucie Nuclear Power Plant,

1977-1998, represented as proportion of total captures.

Figure 21. Slope method for ISL, showing strandings level required to achieve desired slope of

m=0.069.

Figure 22. Effect of the recovery factor (F) on PBR for a population of 2500.
x NOAA Tech. Mem. NMFS-SEFSC-444

Preface

At the behest of the National Marine Fisheries Service ( NMFS), the Southeast Fisheries

Science Center (SEFSC) established The Turtle Expert Working Group (TEWG) in February

1995. The charge to the TEWG was first identified in the Endangered Species Act Section 7

Consultation Biological Opinion for Shrimp Trawling in the southeastern United States under the

Sea Turtle Conservation Regulations issued in November 1995. Specifically, NMFS was

required to select a team of population biologists, sea turtle scientists, and life history specialists

to compile and examine information on the status of sea turtle species. The team should attempt

to identify: a) the maximum number of individual sea turtles of each species that can be taken

incidentally to commercial fishing activities without preventing the recovery of the species; b)

the maximum number of individuals that can be taken incidentally to commercial fishing

activities without jeopardizing the continued existence of any listed sea turtle species; and c) the

number of stranded sea turtles occurring in each statistical zone that indicates incidental takes are

occurring at levels beyond those authorized. Further clarification was received in a letter of

appointment from Mr. Rolland Schmitten, Assistant Administrator, NMFS.

Since the first meeting in June, 1995, there have been several subsequent formal meetings

which have resulted in the publication of status reports for the loggerhead turtle, Caretta caretta,

and the Kemp’s ridley, Lepidochelys kempii (Turtle Expert Working Group. 1998. An

Assessment of the Kemp’s Ridley (Lepidochelys kempii) and Loggerhead (Caretta caretta) Sea

Turtle Populations in the Western North Atlantic. NOAA Tech. Mem. NMFS-SEFSC-409. 96

pp.) The approaches pursued to evaluate the status and condition of these stocks are considered

preliminary and exploratory and we consider this a work in progress. The current report

continues with this preliminary stock assessment effort and constitutes an update to the

previously published report. It is the TEWG’s desire that the results published in these reports be

interpreted with caution and within the extent of the assumptions both with the data and with the

methodologies applied. We wish to note that the exploratory modeling for the Kemp’s ridley

was completed only in the context of attempting to understand the life history of this species and

to extract population level parameters which duplicate empirical data and numerical trends in

abundance. This effort is considered a first step in exploring the use of stage- or age-specific

models to estimate and evaluate mortality in sea turtles.

The initial charge to the group included a request to develop a method which would

establish take levels especially for the commercial shrimp trawl fishery. Actual removals would

be established at levels that would not jeopardize the recovery of these species as stipulated and

defined by the individual species’ Recovery Plans. Determining the impacts of removal on

population recovery in the absence of information on population abundance proved difficult. 

The TEWG pursued the use of strandings data in developing annual interim stranding limits

(ISL’s), recognizing that this proxy for mortality is biased low. These ISL’s were estimated

annually for 1996, 1997, and 1998. These levels were then monitored by NMFS on a regional

basis for each of these two species. These ISL’s do not constitute take levels per se and are

simply a proxy for take which provides guidance relative to average strandings v. an unusually
NOAA Tech. Mem. NMFS-SEFSC-444 xi

high incidence either over the course of a year or over a short period of time (a so-called

“stranding event”). The pursuit of new methods for estimation of ISL’s has been discontinued,

although the methods used are provided in this and previous reports to NMFS.

It is clear to the TEWG that continued work towards developing estimates of take which

do not negatively impact recovery is limited in meaning without a clear understanding of the

status and condition of these stocks. It is the latter effort which will be continued. The

preliminary modeling work reported here will continue to focus data collection efforts and we are

confident that future assessment teams can make progress as more data become available. The

intention of this report is to guide research efforts to the completion of empirical population

estimates to fully evaluate the status of these stocks and evaluate their condition relative to

current and future management activities and recovery goals.

 Much hard work and deliberation have resulted in all of these reports to NOAA. I

personally want to acknowledge all the contributions of data and hard work that were necessary

to finalize this report. Considerable amounts of data were contributed that were not collected by

the NMFS and these data cannot be used without the explicit permission of the originator. 

Dr. Nancy Thompson, NMFS/SEFSC
xii NOAA Tech. Mem. NMFS-SEFSC-444

Executive Summary

o The Kemp’s ridley population appears to be in the early stage of exponential expansion. While

the number of females nesting annually is estimated to be orders of magnitude less than historical

levels, the mean rate of increase in the annual number of nests has accelerated over the period

1987 to 1999. If assumptions that estimated age to sexual maturity and age specific mortality

rates are correct, preliminary analyses suggest that the intermediate recovery goal of 10,000

nesting females by the year 2020 is achievable.

o It appears that the South Florida loggerhead subpopulation, which is the largest in terms of

nesting females, is stable or may be increasing; the status of the Florida Panhandle and Yucatán

subpopulations is unknown. Current trends for the Northern subpopulation indicate this

subpopulation has declined since the 1970's and may have stabilized, but it is unlikely that the

recovery goals for this subpopulation will be achieved.

o Empirical estimates of the number of turtles that can be removed from the population above 

natural mortality without preventing or impeding recovery cannot be made at this time with the

data available. Stranding limits in the form of Interim Strandings Limits (ISL) were developed

for each species and are used as a proxy for mortality. Relationships between strandings levels,

stock size, and mortality rates are unknown.

o ISL values initially were derived using a risk averse approach for both species based on multiyear averages of stranding totals for each species. For the loggerhead turtles, 5-year running

averages were estimated and, depending on the year or region, standard deviations were added or

subtracted to this average to provide the ISL. This approach recognizes the uncertainty of the

status of this species and accounts for differences in status between subpopulations. Five-year

running averages were estimated for the Kemp’s ridley, which were then multiplied by one half

the current rate of increase observed in nests. In this way, it is recognized that the rate of

increase in strandings for this species should not be greater than the population rate of increase.

Additional methods used in 1998 for ISLs included Potential Biological Removal (PBR) for

loggerhead adults and a method for Kemp’s ridleys which attempted to match the trend in

strandings with the trend in hatchling production over the past 12 years, minus 1 standard

deviation for uncertainty. No method for setting strandings limits was completely satisfactory to

all Group members.

o Significant data gaps exist which limit the pursuit of complete age specific assessments for

these species and subpopulation-specific assessments for the loggerhead turtle.

Recommendations for research to obtain these data are included. Most notably, for both species

empirical estimates of age and age specific survivorship are critical. Continued work to define

subpopulations and rates of mixing of subpopulations on foraging areas of the loggerhead turtle

is also needed.
NOAA Tech. Mem. NMFS-SEFSC-444 xiii

o Estimates of mortality and growth must be developed and based on empirical data collected in

the field. This includes estimates of catch and mortality associated with fisheries and other

human activities throughout the range of these species. The only way to develop accurate

estimates of catch and mortality in fisheries is to provide observer coverage over a statistically

valid portion of the fishing effort throughout the range of these turtles. 
xiv NOAA Tech. Mem. NMFS-SEFSC-444

Turtle Expert Working Group Members (as of October 1998)

Crouse, Deborah

Division of Endangered Species

U.S. Fish and Wildlife Service

4401 N. Fairfax Dr., Rm 452

Arlington, VA 22203

Crowder, Larry B.

Duke University Marine Laboratory

135 Duke Marine Laboratory Road

Beaufort, NC 28516

Epperly, Sheryan P.

Southeast Fisheries Science Center

NOAA, National Marine Fisheries Service

75 Virginia Beach Drive

Miami, FL 33149

Gabriel, Wendy L.

Northeast Fisheries Science Center 

NOAA, National Marine Fisheries Service

166 Water Street

Woods Hole, MA 02543

Harris, Michael

Georgia Department of Natural Resources

2070 Highway 278

Social Circle, GA 30025

Henwood, Terry

Pascagoula Laboratory

NOAA, National Marine Fisheries Service

P.O. Drawer 1207

Pascagoula, MS 39568

Heppell, Selina S.

Duke University Marine Laboratory

135 Duke Marine Laboratory Road

Beaufort, NC 28516
NOAA Tech. Mem. NMFS-SEFSC-444 xv

present address:

Department of Fisheries and Wildlife

Oregon State University

104 Nash Hall

Corvallis, OR 97333

Márquez-Milan, Rene

Programa Nacional de Investigacion de Tortugas Marinas

Centro Regional de Investigacion Pesquera

Playa Ventanas S/N, A.P. 591

Mananillo, Colima, México, 28200

Murphy, Sally

South Carolina Department of Natural Resources

P.O. Box 12559

Charleston, SC 29422

Teas, Wendy

Southeast Fisheries Science Center

NOAA, National Marine Fisheries Service

75 Virginia Beach Drive

Miami, FL 33149

Thompson, Nancy

Southeast Fisheries Science Center

NOAA, National Marine Fisheries Service

75 Virginia Beach Drive

Miami, FL 33149

Witherington, Blair

Florida Fish and Wildlife Conservation Commission

Florida Marine Resources

Department of Environmental Protection

9700 South A1A

Melbourne Beach, FL 32951

Observers:

Conant, Therese

Office of Protected Resources

NOAA, National Marine Fisheries Service

1315 East-West Highway

Silver Spring, MD 20910
xvi NOAA Tech. Mem. NMFS-SEFSC-444

MacPherson, Sandra L.

U.S. Fish and Wildlife Service

6620 Southpoint Dr., South

Suite 310

Jacksonville, FL 32216

Possardt, Earl

U.S. Fish and Wildlife Service

University of West Georgia

Department of Biology

Carrollton, GA 30118

Schmid, Jeffrey R.

Cooperative Institute for Marine and Atmospheric Sciences

University of Miami

4600 Rickenbacker Causeway

Miami, FL 33149

Schroeder, Barbara

Office of Protected Resources

NOAA, National Marine Fisheries Service

1315 East-West Highway

Silver Spring, MD 20910
NOAA Tech. Mem. NMFS-SEFSC-444 1

Part 1. Kemp’s Ridley Status Report

Summary

The Kemp’s ridley sea turtle declined precipitously from tens of thousands of nesting

females in the late 1940's to around 300 nesting females in 1985. Conservation efforts by

Méxican and U. S. agencies have aided this species by eliminating egg harvest, protecting eggs

and hatchlings, and reducing at-sea mortality through fishing regulations. From 1985 to 1999,

the number of nests observed at Rancho Nuevo, Tamaulipas, and nearby beaches has increased at

a mean rate of 11.3% per year. Current totals exceed 3000 nests per year, allowing cautious

optimism that the population is on its way to recovery. At current rates of increase, the

population should reach 10,000 nesting females around 2020, the target given in the Kemp’s

ridley Recovery Plan (USFWS and NMFS, 1992). An analysis of number of nests counted

versus hatchlings released suggests that there has been an increase in survivorship from hatchling

to maturity in the past 5-6 years. An age-based population model was constructed to calculate

potential recovery times, population size, and reasonable take limits. The model was based on

growth curves of recaptured juveniles, a mortality estimate from strandings of juvenile turtles,

and fitted mortality rates for other life stages. A reduction in mortality of the benthic feeding

stages was required to achieve a reasonable model fit; without the mortality reduction, there was

no rapid increase in the predicted number of nests in recent years. The model projections suggest

that the population will continue to increase rapidly if the mortality estimates are reasonable and

remain fairly constant over the next two decades. However, because of uncertainties in the

mortality rates, the range of population size estimates was large (approximately 8-fold). Thus,

the model was unable to adequately calculate the number of turtles in the population nor

reasonable take limits. Continued effort to estimate mortality rates through mark-recapture

studies may reduce the range of possible population sizes.

History of Nesting at Rancho Nuevo

Historic information indicates that tens of thousands of Kemp’s ridley sea turtles nested

at Rancho Nuevo, México, in the late 1940's (Hildebrand, 1963). Before then and until the late

1960's, egg harvesting occurred and nearly all eggs were removed (R. Márquez-M, personal

communication). In 1966, a turtle camp was established in Rancho Nuevo which protected nests

along approximately 25 km of the beach from April to July of each year. Over the past two

decades, poaching on protected beaches has been reduced to nearly zero eggs per year. From

1966-1977, 100-400 nests per year were transplanted to a protective corral while the remaining

nests were left “in situ”. The “in situ” nests experienced high predation and therefore low egg

survival. In 1978, a joint program between the U.S. and México was initiated which increased

the length of the beach sampled and number of nests protected. The proportion of nests left in

situ was reduced to <10%, thereby greatly increasing survival of the eggs. During the first years

of the joint U.S./Méxican conservation effort (1978 – 1983), at least 10 aerial surveys of the

nesting areas were completed from the U.S México border southward to Veracruz. The

observation of Kemp’s ridley tracks outside of the Rancho Nuevo area were few and scattered

(Márquez-M, 1994). However, in recent years, more nests were observed north and south of
2 NOAA Tech. Mem. NMFS-SEFSC-444

Rancho Nuevo and as a result the nesting beach surveys were extended in 1990 and 1991,

respectively (Figure 1). The northern camp, called Tepehuajes/Ostional, covered 45-50 km of

beach. In 1991, another camp was established to the south of Rancho Nuevo in Barra del Tordo,

increasing total beach coverage to over 100 km. Since 1996 more small camps have been added

and now over 120 km of beach is included in this protection, covering nearly all of the state of

Tamaulipas to Lechuguillas, Veracruz. Because new surveys added since 1996 may inflate the

population growth rate estimate, the Working Group decided to include Rancho Nuevo, North

and South camps only in subsequent analyses and model runs. Analysis performed on the nests

and hatchlings released from the Rancho Nuevo beach alone rarely altered the results

qualitatively, and there is reason to believe that females nesting at Rancho Nuevo moved north

and south following a hurricane in 1990 (R. Byles, personal communication).

New Nesting Information

México – Update of nest numbers 1966-1998

The nesting data presented in the previous report (TEWG, 1998) for years 1978-1995

were taken from draft reports of the U.S. Fish and Wildlife Service – Joint U.S.-México

Program. We have updated this information (Table 1). The previous data were considered

incomplete, as they were grouped over a season for some camps, and in some cases included data

from extra camps. Some minor changes were also made to nesting data from 1966-1977 because

of information provided by the Méxican Instituto Nacional de Pesca. Some discrepancies with

U.S. Fish and Wildlife Service reports occur because of uncounted nests that were lost by natural

predation, robbed, or left “in situ” and the exclusion of nests that were not discovered until after

hatching and those that were donated to the U.S. Padre Island-Galveston experiments. The

current nest tally also omits duplication of numbers of nests already included in the numbers

protected in corrals or boxes. The annual number of nests and hatchlings released given in Table

1 were used in all of the trend and model analyses presented in this report. 

Nesting records from South Texas

Isolated Kemp’s ridley nests were documented along the south Texas coast from 1948-

1994, with a total of thirteen nests being found during this forty-six year period. A primary goal

of the joint U.S.-México conservation project was to re-establish a nesting colony at Padre Island

National Seashore, and thousands of hatchlings were “imprinted” to that area (Woody, 1985). 

From 1995-1998, Kemp’s ridley nesting in south Texas showed a marked increase with 4, 6, 9

and 13 nests being documented in each of these years, respectively. Of these thirty-two nests, six

were confirmed by tags to have been laid by turtles which were part of the headstart experiment

and were imprinted at Padre Island National Seashore, seven turtles were examined and found to

be “wild” turtles, and the remaining nineteen turtles were not examined for tags. (Shaver and

Caillouet, 1998).
NOAA Tech. Mem. NMFS-SEFSC-444 3

Life Stage Definitions and Geographic Locations

Post-hatchling stage

We have defined the post hatchling stage as the time the hatchlings enter the water to age

2 years, and have assigned turtles that are 5-20 cm straight-line carapace length (SCL) to this

stage. Post-hatchlings disperse from Méxican beaches of the Gulf of Mexico, presumably by

major oceanic currents (Collard and Ogren, 1990), and development occurs in the epipelagia. 

The presence of juvenile turtles along both the Atlantic and Gulf of Mexico coasts of the U.S.,

where they are recruited to the coastal benthic environment, indicates that post hatchlings are

distributed in both the Gulf of Mexico and Atlantic Ocean. Our stage-length of two years is an

average based on evidence from three sources: 1) growth curves derived from

skeletochronological analysis (Zug et al,. 1997) and mark-recapture data (Schmid and Witzell,

1997), 2) correlation analysis of hatchlings released and the number of strandings of small turtles

(20-30 cm SCL) 1, 2, 3 and 4 years later (Table 2), and 3) preliminary data from 6 dead stranded

turtles that had been tagged as hatchlings at Rancho Nuevo and were recovered 1 and 2 years

after release (C. Fontaine, NMFS-Galveston, personal communication). Variability in growth

rates among individuals likely means that the actual length of time spent in the post-hatchling

pelagic stage varies from 1 - 4 or even more years.

Benthic immature stage (20-60 cm)

We have assigned turtles that are 20 to 60-cm SCL,to the benthic immature stage,

representing the size class that has recruited to the coastal benthic habitat from the pelagic posthatchling stage (Ogren, 1989). The 6 stranded turtles originally tagged on the nesting beach

ranged from 22-28.5 cm SCL, suggesting that turtles of this size are fully recruited to the benthos

(C. Fontaine, NMFS-Galveston, personal communication). We further divided this stage into

small (20-50 cm SCL) and large benthic immatures (50-60 cm SCL) based on the frequency of

strandings used in the mortality estimates (see modeling section below). Benthic immature turtles

are found in coastal habitats of the Gulf of Mexico and the entire U.S. Atlantic coast. The

location and size classes of dead turtles recovered by the Sea Turtle Stranding and Salvage

Network suggests that benthic immature developmental areas occur in many areas along the U.S.

coast and that these areas may change given resource quality and quantity. Growth models from

mark-recapture data suggest a 7-9 year duration for the benthic immature stage (Schmid and

Witzell, 1997; Appendix 1).

Adult stage

Based on size data collected on the Méxican nesting beach since the mid-1960's

(Márquez-M, 1994 and unpublished data), we have defined adults as turtles larger than 60 cm

SCL. The distribution of dead turtles >60 cm SCL is limited to the southeast U.S. coast and the

Gulf of Mexico and are essentially absent from strandings records for the northeast U.S. coast. 

Recent tag returns indicate that individuals from the Atlantic tagged as immature animals do nest
4 NOAA Tech. Mem. NMFS-SEFSC-444

on Méxican Gulf of Mexico beaches and therefore do move between these water bodies (Schmid

and Witzell, 1997; Witzell, 1998).

Status and Trends

Nesting data

Since 1985, the lowest nesting year (740 total nests), the number of Kemp’s ridley nests

counted at Rancho Nuevo, North and South Camps has increased approximately 11.3% per year

(r-square = 0.94, P<0.001, 95% C.I. slope = 0.096 – 0.130; Figure 2). Increase in hatchling

production from 1985-1998 was slightly less, 9.5% per year (r-square = 0.955, P<0.001, 95% C.I.

slope = 0.082 – 0.108). We assume that little nesting occurred North and South of Rancho

Nuevo prior to 1990 (see History of Nesting at Rancho Nuevo, above), so nests at the North and

South camps are included in the totals. Like Rancho Nuevo, the number of nests counted at

North and South Camps has increased since census efforts began in 1990 and 1991 (Figure 3). 

The annual rate of increase for nests at Rancho Nuevo only, 1985 – 1999, is 7.9% per year (rsquare = 0.81, P<0.001, 95% C.I. slope = 0.056 – 0.102). It is uncertain whether the current rate

of increase will continue. Also, the large number of nests counted in 1998 was unprecedented

and may or may not indicate an increase in the population growth rate (preliminary data from the

1999 nesting season suggests that fewer nests were observed this year). However, we can project

the number of nests with the confidence intervals predicted by the exponential trend (Figure 4). 

Given 2.5 nests per female, if the population continues to grow at 9.6 – 13% per year it will

reach the target of 10,000 nesting females around 2014 - 2025. 

Nests vs. hatchlings released 8, 10, and 12 years earlier

When we plotted the number of nests observed against the number of hatchlings released

8, 10 or 12 years earlier, we found an increase in the number of nests per cohort for the past 5-7

years, regardless of time lag (Figure 5, a-c). Heppell et al. (1999) found a similar result when

they restricted the analysis to Rancho Nuevo nests only. Several large cohorts produced in the

first years of the joint U.S.-México recovery effort do not correspond with larger nesting

numbers. While a "build-up" of females is expected with increasing cohort sizes, the graphs

suggest that increased hatchling production alone is not responsible for the current increase in

nest numbers. An increase in hatchling-to-adult survivorship and/or adult female survival in the

late 1980's to early 1990's may be responsible.

Strandings

Complete strandings information by size class, fishing zone and year are provided in

Appendix 2. Figure A2.1 shows the statistical zones for which sea turtle strandings are reported. 

Monitoring effort is not directly comparable between zones but has been reasonably consistent

over this period. Figure A2.2 a, b, and c depict the Kemp’s ridley strandings reported by

statistical zone, by year, 1986-1997. These data do not include incidental captures, posthatchlings, cold-stunned, or head-started animals. Over the 12-year period 1986-97, the Gulf of
NOAA Tech. Mem. NMFS-SEFSC-444 5

Mexico accounted for 69% of total strandings, with the southeast and the northeast accounting

for 25% and 6% respectively (Table 3). Strandings were variable, but lower on average in the

1980's and have increased approximately 70% over this time frame (Table 4). The increase has

been confined to the southeast and Gulf of Mexico. Strandings of Kemp's ridleys have not

increased in the northeast.

Strandings “hotspots”

The western Gulf of Mexico (zones 13-21), where shrimp fishing effort is consistently

high, has accounted for the highest proportion of Kemp’s ridley strandings (Figure A2.2),

particularly of adults. There is no survey effort in zones 15 and 16, due to inaccessibility of

shoreline, and coverage is low in zones 13 and 14. The lack of data from these zones may or

may not reflect a lack of strandings. Over this time frame, zones 17-21 have accounted for

44.8% of all Kemp’s ridleys, including 75% of all adults, stranded over the entire east Coast and

Gulf of Mexico. There appears to have been a shift in the strandings of adults over this time

frame, with adults being more common in zones 18 and 19 prior to 1992, and more common in

zones 20 and 21 from 1994 to the present. In 1993, in zones 13 and 14, a large group of small

Kemp’s stranded in late May-early June. In the eastern Gulf of Mexico (zones 1-12, partial 24-

25), survey coverage is low in zones 1, 3, 6, and 7 due to inaccessibility and zone 2 has very little

land mass. Zones 4 and 5 account for most of strandings in the eastern Gulf. Red tide events

have also been reported periodically from these zones (Foote et al., 1998).

Along the southeast U.S. coast, monitoring coverage is low in zones 24 and 25. 

Shrimping effort is generally high in zones 29-34. Zones 30 and 31 account for the majority of

Kemp’s ridley strandings in this region. In the northeastern U.S. Kemp’s ridley strandings are

rare, except for cold-stunning events in the winter months, which are not included in this

analysis. No adult ridley strandings have been reported from this region.

Systematic surveys for sea turtle strandings do not occur in México but a few (<10),

almost always adult-sized Kemp’s ridleys strand each year at Rancho Nuevo during the nesting

season. In 1998, 30 Kemp’s strandings were recorded, but the area being monitored for nesting

increased in recent years (from 25 km in the 1980's to 120 km in 1998), so it is unclear whether

this represents an actual increase.

Changes in size distributions

Kemp’s ridleys stranded in 1986-98 ranged from about 10-70 cm. Because the

population appears to be growing rapidly, we expected to see a shift in the size distribution

towards smaller turtles. We found no evidence of a shift in the mean size of Kemp’s ridleys

stranded over this period, but the size distribution of stranded Kemp’s ridleys was significantly

different in the 1996-97 compared to 1986-87 (Kolmogorov-Smirnov Test, P<0.0001; Figure 6). 

Many more large benthic immature ridleys stranded in the 1990s relative to the 1980s. While

analysis of stranded turtles can potentially provide information about shifts in size distributions,
6 NOAA Tech. Mem. NMFS-SEFSC-444

and therefore population trends, strandings may not be indicative of actual population trends

because of biases associated with fishing mortality.

Fishery independent, in-water studies of sea turtles have been carried out at multiple sites

in the southeastern U.S. Although the objectives and target species of these studies vary, the

primary goal has been to characterize regional aggregations of sea turtles. Some studies have

been conducted for many years (10+), others have only recently been initiated and still others

have been intermittent or aperiodic. Mark-recapture surveys of Kemp's ridleys reveal important

information about size distributions and growth rates of immature turtles (Appendix 1). Ogren

(1989) first suggested that aggregations of smaller Kemp’s ridley turtles were increasing in U.S.

waters as a result of the protection at the nesting beach. However, no quantitative data were

available to support this claim. Schmid (1998) noted a higher percentage of smaller size classes

captured from 1986-95 when compared to the size distribution of Kemp’s ridleys reported in

1955 by Carr and Caldwell (1956). Furthermore, the seasonal mean carapace lengths of Kemp’s

ridley turtles from the former commercial turtle fishery were 5 – 10 cm greater than those

recorded in recent tagging studies. The differences may be indicative of an increased number of

smaller turtles in this area or could be the result of size bias in the former turtle fishery. The

mean carapace length of Kemp’s ridley turtles captured at Cape Canaveral, Florida from 1986-91

(37.0 cm, n = 107, Schmid, 1995) was slightly smaller than that reported from 1978-84 (38.6 cm,

n = 40, Henwood and Ogren, 1987). However, this difference may not be significant and could

be the result of sampling error owing to dissimilar sample sizes. Reynolds and Sadove (1998)

reported no significant change in the mean carapace lengths of Kemp’s ridley turtles captured in

New York waters from 1986-96. Despite possible temporal shifts in size classes, there are no

quantitative data that conclusively demonstrate demographic shifts in the population.

In-water surveys

In-water studies targeting Kemp’s ridleys either have been or are being conducted at

several sites along the U.S. Gulf of Mexico and Atlantic coasts. These sites are: Sabine Pass

(Tx.), Calcasieu Pass (La.), Big Bend area (Fla.), Cedar Key (Fla.), Ten Thousand Islands (Fla..),

Pamlico Sound (N.C.), Chesapeake Bay (Va.), and Long Island Sound (N.Y.). There are

numerous difficulties in designing in-water studies to ensure that standardized catch per unit

effort (CPUE) methods are employed such that data can be compared not only within site but

also site-to-site. Environmental conditions, including weather, depth strata, current patterns,

visibility, water temperature, by-catch, and investigator experience, can greatly affect turtle

capture rates. Biological conditions, such as prey abundance and habitat availability, complicate

sampling strategies that may be based on random distribution of the target species. Changes in

effort, variability in temporal sampling, and different sampling methodologies make analyses of

population trends particularly difficult.

Evaluating trends in abundance based on in-water survey data would be statistically

unreliable at this time without some method of standardizing the amount of effort. Because most

surveys are restricted spatially or temporally, surveys may monitor local availability at the time

of the survey rather than the relative abundance of the population as a whole or a life stage of the
NOAA Tech. Mem. NMFS-SEFSC-444 7

population. High variability in survey indices due to patchy distributions means that any

subsequent survey estimates will probably continue to have high variances. Trawl surveys may

be a feasible means to monitor status of benthic stages of the population, but are not the only

method. The Working Group recommends the design and implementation of a pilot survey,

bearing in mind the patchy nature of population distribution.

The only in-water trawl surveys that appeared to have a potential for documenting longterm changes in Kemp’s ridley abundance were the NMFS data base from the 1970's and early

1980's (reported in Henwood and Stuntz, 1987) and recent trawl surveys conducted by the Gulf

and South Atlantic Fisheries Development Foundation. A comparison of these data was

conducted by the Foundation (1998; Jamir, 1999).

There were some differences in survey design between the two studies. The original

NMFS observer program was conducted to document the rate of incidental capture and mortality

for all turtle species that might encounter shrimp trawls. The survey approach was very

simplistic; observers were placed aboard cooperative shrimp trawlers, the vessel fished as it

normally would, and all turtle encounters were recorded. The Foundation study was conducted as

a direct result of 1996 Congressional language requiring an investigation of alternatives to TEDs.

The Foundation more or less attempted to repeat the original Henwood and Stuntz trawl surveys,

with a primary intent of demonstrating that offshore shrimp trawlers did not take turtles and

therefore, should be exempt from the TED regulations. The Foundation effort (134 sea days

nearshore Gulf, 463 days offshore Gulf and 125 days Atlantic) clearly reflects the emphasis of

this study, but the distribution of effort does not preclude some comparisons between the two

surveys. The Foundation survey approach used observers to document take and the vessel was

asked to fish as it normally would (except for 1 hour tow time restrictions in nearshore water to

reduce turtle mortality).

Preliminary analysis of the two data sets indicated large increases in CPUE for Kemp’s

ridley and loggerhead turtles when compared to the original Henwood and Stuntz estimates. For

Kemp’s ridley turtles, as much as a 35-fold increase in CPUE was calculated for some areas. A

number of factors must be considered to explain these large changes in CPUE. Most importantly,

at the time of the original NMFS work, Kemp’s ridley population levels were very low, and the

data were 99.9% zeros (tows where Kemp’s were not caught) with a few ones where a rare ridley

was taken. Also, the fleet was not using TEDs so Kemp’s ridleys on the shrimping grounds were

probably extremely rare due to high mortality rates. Thus, the CPUE calculations from the late

1970's and early 1980's were composed of rare event captures of a highly depleted stock. The

Foundation study, on the other hand, was conducted after almost 10 years of full implementation

of TEDs, 30 years of nesting beach protection and 10 years of increases in nests and number of

hatchlings produced at Rancho Nuevo. These factors should result in an increased number of

turtles available for capture, and the Foundation data supported these expectations. Foundation

observers were able to catch more turtles with much less effort which resulted in huge changes in

CPUE.
8 NOAA Tech. Mem. NMFS-SEFSC-444

While the observed increases in CPUE are impressive, caution should be exercised in

interpreting these results. When dealing with small CPUE values (as with the old NMFS

estimates Kemp’s ridley), any increase in number of turtles caught at a given level of effort or

decrease in effort with the same number of turtles caught can result in some “multi-fold” increase

in CPUE estimates. The Foundation study indicates that there are now Kemp’s ridleys in

nearshore waters in measurable quantities, and these turtles were not there or were present in

very low abundance 15 years ago. What CPUE will not explain in this case is how much the total

population has increased. To illustrate this point consider a comparison of CPUE from an

overfished stock where 95% of the fish have been taken from the habitat, with CPUE from that

same habitat after ten years of protection has allowed 100% reoccupation of that habitat. Now

consider that this habitat is only one of many occupied by this species, the rest of which have

never been fished. A twenty-fold increase may have occurred in the overfished habitat, but that

tells us nothing about the total population increase.

The Foundation catch rates of sea turtles indicate that future in-water trawl surveys could

produce enough captures to allow monitoring of population abundance and trends in a cost

effective manner. If such surveys were conducted, a randomized sampling design should be used

to cover the entire coastline and not just the shrimping grounds where active fishing is occurring.

This is an important consideration to note because while the NMFS and Foundation data sets

allow estimation of CPUE on the shrimping grounds, nothing can be said about turtle abundance

in areas or times when shrimp fishing is not occurring.

Trawl surveys are probably the best currently available means of obtaining information on

the in-water abundance of sea turtles. The Foundation study, while suffering from less than

optimal effort allocation, provides some good insights regarding the effectiveness of TED

regulations. The increases in CPUE observed in this study are the strongest evidence to date that

Kemp’s ridleys are coming back on the shrimping grounds where they were almost eradicated

two decades ago. The CPUE values reported are encouraging and believable, although they do

not imply multi-fold increases in the total population. What they do suggest is multi-fold

increases in areas where the species was severely depleted and kept down as a result of shrimp

trawling pressure. Kemp’s ridleys now appear to be repopulating their former habitat. 

Conclusions from nesting and strandings data

There is evidence from nest numbers and in-water surveys that the Kemp's ridley

population is increasing. There has apparently been an increase in survivorship to maturity and

the proportion of large benthic immatures in the strandings records has increased, suggesting that

juvenile mortality has decreased. The number of strandings in the Gulf of Mexico increased

sharply in 1994 and 1995 but remained fairly constant (about 450 turtles per year) thereafter. It is

reasonable to believe that some of the recent increase in strandings is due to an increase in

population size. However, because of the time lag required for each cohort to reach maturity and

recruit to the nesting population, there is not a clear connection between changes in strandings or

in-water surveys and changes in the number of nests, nor is there an easy way to estimate the
NOAA Tech. Mem. NMFS-SEFSC-444 9

effect of changes in juvenile fishing mortality on the nesting female population. We attempted to

elucidate this relationship using a population model, described below.

Nest protection in México – Future strategies

Since 1978, nearly all Kemp's ridley nests have been moved to protected "corrals" for

incubation, thereby greatly reducing the loss of eggs and hatchlings to erosion and predation. As

the population continues to increase and expand north and south, the proportion of nests that can

be completely protected will decrease. From 1987 to 1999, the number of nests reported has

increased annually from less than 750 to approximately 3400 for Rancho Nuevo, Tepehuajes and

Barra del Tordo camps. Based on the relative importance of these camps for nesting and the

number of workers and equipment that can be accommodated, it is estimated that the total

number of nests that can be protected in corrals is about 5,000. This 5,000 total includes about

4,000 from Rancho Nuevo and about another 1,000 at the north and south camps. Another 1,000

nests could be protected “in situ” with wire screens. Any remaining nests will have to remain

where laid with increased protection through increased beach monitoring with additional beach

workers. In addition, the number of beach workers required each year is determined by the

nesting behavior of the turtles. In some years, the turtles aggregate along specific areas of the

beach and fewer beach monitors are required as compared to years when nests are scattered.

While efforts will be made to protect as many nests as possible, it is realistic to anticipate that the

current nest survival rate of approximately 65% will not continue indefinitely.

Population size and trends from a deterministic age-based model

Model modifications from the 1998 report

Models are most useful when they are updated with new data and modified according to

the best information available. The demography of sea turtles is, generally, poorly understood,

and new information must be incorporated into model form and parameterization whenever

possible. Changes to the model that we developed in 1995 are the results of new information and

discussion:

1) The small benthic immature mortality rate is based on a re-analysis of strandings information,

using a variety of cohort-based approaches to account for changes in cohort size and variance in

strandings levels, rather than pooling year classes. This mortality rate is applied to age classes 2 -

6, based on the growth curve from mark-recapture data (Appendix 1).

2) The mortality rate of large benthic immatures is now equal to that of adults, rather than small

benthic immatures, based on the poor fit of the catch curves to strandings of age classes greater

than 6 years and the argument that large immature turtles (>50 cm SCL) are likely to have a

lower mortality rate (and more similar to the adult mortality rate) than smaller turtles.

3) The number of nests used in the model fitting exercise has been updated according to Table 1.
10 NOAA Tech. Mem. NMFS-SEFSC-444

4) The change in nest survival with nest density has been updated according to recent discussions

with the INP and camp managers.

5) The number of years used to calculate the fitted parameters (pelagic immature and adult

survival and the post-1990 mortality decrease) has been varied to achieve the model that best-fits

both past and recent trends in the nesting data.

These changes have altered the fitted parameters, expected growth rates, and population size

estimates of the models. We have done a sensitivity analysis similar to that shown in the 1998

report. These models represent our understanding to date of Kemp's ridley population dynamics,

but will undoubtedly change as more information becomes available to run more complex

simulations.

Model form

A deterministic age-based model was implemented with the short-term objective of

providing a framework to organize and evaluate available data and parameter estimates. The

development of a model to track stock status in terms of abundance and mortality at age, to

project stock abundance levels conditional on current stock status, and to provide management

advice will require reliable estimates of input parameters which are currently unavailable.

The number of female hatchlings produced in each year (Table 1) are projected forward

as a cohort. Each cohort passes through four different life stages (Table 5), where age at sexual

maturity (α) is determined by a growth curve (Appendix 1). All animals aged or higher are

assumed to undergo mortality at the rate of mature individuals and reproduce every other year.

The model is used to estimate three parameters through a least-squares analysis that

compares nests predicted by the model to nests observed each year (Table 1). Instantaneous total

mortality is estimated for two age groups: pelagic immature (Zp), and the combined large benthic

immature-adult stages (Za). A mortality rate for small benthic immature turtles (Zb) is based on

empirical data (strandings, see below) independent of the model. The mortality rate of the large

benthic immature stage was assumed to be more similar to that of the adult stage than to that of

the small benthic immature stage. This is because large benthic immature animals are rare in

strandings in the Gulf of Mexico and are assumed to be distributed in different areas from small

benthic immature animals. This makes the large benthic juveniles potentially less vulnerable to

mortality than small benthic immature animals. A third parameter reflecting a post-1990

decrease in mortality (f) was also estimated. This multiplier was applied to the small benthic

immature and large benthic immature-adult mortality rates after 1990 to address a lack of fit (a

trend from maximum negative residual values in 1989 to maximum positive value in 1998)

which arises when a single mortality rate per life history phase is applied over the duration of the

1966-1999 time series (Figure 7, in Model Results below). The Excel 7.0 Solver algorithm was

used to estimate the parameters by comparing predicted numbers of nests to observed values.
NOAA Tech. Mem. NMFS-SEFSC-444 11

Thus, five different mortality rates are obtained over the time and age range of the model

(Table 6) where Zi

, Za

, and f are estimated using the model fitting exercise and Zb

 is input or

estimated independently. The post-1990 multiplier (f) is a number less than 1.0, representing a

proportional decrease in Zb and Za

. Mortality rates are not re-estimated annually. Observational

data consist of annual numbers of nests at Rancho Nuevo and north and south camps (1966-

1999) and annual hatchling production (1996-1998), which serves as cohort size for each year

(Table 1). Input parameters included adult female remigration rate, mean number of eggs per

nest, female age at maturity, sex ratio in the population, and survivorship of benthic immature

turtles (see below). The temporal resolution is annual, with time periods beginning 1 July,

corresponding with average time of hatchling production and nesting. The number of adults is

modeled without age structure and no maximum age is set.

Starting population sizes were derived from estimates of nests and hatchlings in 1966

(Table 1). Past analyses have shown that the model is relatively insensitive to the nest estimate

used in 1966, although this value does affect the model estimate of adult survival (Appendix 2

and 3; TEWG 1998). The estimated number of nests in 1966 was converted to an estimate of

nesting females in 1966 by applying rates of nests per female and remigration rate. That pool of

nesting females was decremented by adult mortality rates between 1966 and the year the 1966

cohort recruited to the nesting population. This assumes that hatchling production before 1966,

which would have supported recruitment to the nesting population in the intervening years, was

effectively zero, and consequently that there was no recruitment to the nesting population in

1966-1976.

The number of female hatchlings is based on an assumption of a 1:1 sex ratio for

hatchlings produced, and of no other surviving hatchlings from sources other than monitored

nesting beaches at Rancho Nuevo, North and South Camps. The ratio between hatchlings

produced and subsequent contribution of a cohort to the nesting population is assumed to be

constant for all cohorts recruiting by 1990, because age-specific mortality rates are fixed for all

ages over that period, after which the effect of the mortality multiplier is incorporated. Similarly,

reproductive parameters, including remigration rates, number of nests per female, and growth

rates (i.e., stage lengths and age at maturity) are assumed constant over the 1966-1999 series.

Parameterization

Nesting data used

We considered the modeled population to include nesting and hatchling production at

Rancho Nuevo as well as the North and South Camps. Prior to the mid 1980s, aerial surveys

documented few crawls outside Rancho Nuevo; it seems unlikely that many hatchlings were

produced given the small numbers of nests and that there was no protection afforded to the eggs

or hatchlings. So we assume that hatchlings estimated at Rancho Nuevo led to nests

subsequently detected at Rancho Nuevo and elsewhere. Given that Kemp’s ridleys nest numbers

reached their lowest numbers in the mid-1980s, one might expect their range to expand with
12 NOAA Tech. Mem. NMFS-SEFSC-444

increases in the number of females in the nesting stage. Nesting surveys began at North and

South Camp in 1990 (Table 1).

In the 1998 report, the model was fit to years 1978-1995, representing all years for which

there were complete nest counts. Nesting information is now available through 1999 (Table 1).

We wanted to produce a model that fit the time series well but could also be verified by recent

nest counts and used to roughly estimate future population growth. The number of nests observed

in recent years is more variable than in the 1980's (Figure 2), and the number of years used to fit

the model can affect the fitted survival rates, the post-1990 multiplier, and the model projections.

We ran a sensitivity analysis to see how the addition of each new year of data would affect the

model fit for the model fit years (1978-199x), early years (1978-1995), recent years (1996 to

1999) and for the entire available time series (1978-1999) (Table 8). We used Zb = 0.5 for the

analysis (Zb

 does not affect the residuals; TEWG, 1998).

 

Reproduction

For the model fitting exercise, nests deposited per female and years between nestings

(remigration interval) were needed to convert nests to adult females, and sex ratio was needed to

convert females to total population size. Cohort sizes for years 1966 - 1998 were the number of

hatchlings released from Rancho Nuevo, North and South camps, so additional reproductive

parameters, such as eggs per nest and hatch success, were not needed until we projected

population size into future years. For model projections, we examined the effects of reduced

nest survival with nest density, as might be expected as more nests are left in situ (Table 7). We

considered published and unpublished information in developing the following assumptions and

estimates for reproductive parameters:

1. The number of nests per female per season was calculated as 3.075 in a physiological and

ultrasonographic study by Rostal (1991). Pritchard (1990) estimated 2.31 nests per female per

season, and a recent tagging study using PIT (passive integrated transponder) tags (R. MárquezM., personal communication) estimated 1.8 nests per female per season. We used 2.5 nests per

female per season, which is close to the mean of means (2.4 nests per female per season) of

these studies.

2. Adult female remigration rate to the nesting beach was estimated from mark-recapture data

(Márquez-M et al. 1989). We updated these data with current information (R. Márquez-M.,

personal communication). Roughly 20% of the turtles nest annually, 60% nest every 2 years,

15% nest every 3 years, and 5% every 4 years. The weighted mean remigration rate was 2.05

years, which we truncated to 2 years and used in the model. However, we recognized that

estimation of remigration rate must take into account adult mortality rate (Frazer 1989) and the

probability of encountering nesting females (Pritchard, 1990), but neither of these were known.

3. A mean number of eggs per nest of 100.8 was noted by U.S. Fish and Wildlife Service and

National Marine Fisheries Service (1992), which we truncated to 100 eggs per nest.
NOAA Tech. Mem. NMFS-SEFSC-444 13

4. For population projections beyond 1999, hatch success (i.e., egg survival) for nests in corrals

was set at 65%, based on the mean of [annual hatchlings released/(nests*100 eggs/nest)] for

years 1990-1998 (Table 1). Hatch success was reduced step-wise to simulate a reduction in nest

protection with nest density. Protection of nests from predators and poachers at Rancho Nuevo

nesting beaches is effort-dependent, and can change with time due to changes in numbers of nests

laid and availability of resources to protect them. Under previous plans of the Instituto Nacional

de la Pesca (INP) and U.S. Fish and Wildlife Service (USFWS), an agreed upon maximum of

5,000 nests would be protected in corrals, with a maximum of an additional 1,000 nests protected

in situ by screening. Nests in excess of 6,000 would be left in natural surroundings with no

protection, except perhaps predator control and beach patrols. For our model projections, the

hatch rates were imposed on the model according to such best approximation protection

scenarios (Table 7).

Sex ratio (all age classes)

For the model exercise in our first report we assumed a sex ratio of 1:1 (TEWG, 1998). 

Since then two studies in the northern Gulf of Mexico have provided support for that assumption

(Coyne and Landry, 2000; Gregory and Schmid, in review). Although no significant differences

were found from a 1:1 ratio across months and sampling locations, females did comprise the

majority of turtles sampled in these areas.

Age and growth (see Appendix 1)

Age at maturity for Kemp’s ridley turtles was estimated by Márquez-M (1972) as 5-6

years. The U.S. Fish and Wildlife Service and National Marine Fisheries Service (1992) thought

that estimate too low, but offered no better estimate. Zug et al. (1997) estimated Kemp's ridley

maturity between 11 and 16 based on skeletochronological age estimates for stranded turtles and

an assumed mature size of 65 cm SCL. Caillouet et al. (1995) developed a growth model for

known-age head-started Kemp's ridleys and estimated a 10 year age at maturity, assuming size at

maturity of 60 cm SCL. Schmid and Witzell (1997) estimated age to maturity at 8 – 13 years

based on the recapture data of wild subadults in Florida and the minimum and mean carapace

lengths of nesting females. An analysis of the NMFS Miami Laboratory tagging database

(Appendix 1) indicates an 8-12 year age to maturity estimate using the criteria of Schmid and

Witzell. Given the variability in estimates, we assumed the actual age at maturity for Kemp’s

ridley turtles to be within 7 - 15 years. Because size at first nesting and growth rates vary, we ran

preliminary model runs with 8, 10 and 12 years to maturity to test the model's sensitivity to this

parameter. We assumed "knife-edge" maturity (all individuals become mature at the given age)

because there is no information available on the proportion mature at age or size.

Benthic immature mortality from strandings catch curve estimates (Zb

)

Catch curves were constructed both by cohort and by standing stock based on length

composition of strandings data and updated growth curves. Strandings at length by month and

year were re-indexed for consistency with aging by cohort slicing: the “Ridley year” i was
14 NOAA Tech. Mem. NMFS-SEFSC-444

defined from 1 July, calendar year i to 30 June calendar year i+1. An updated von Bertalanffy

growth curve (Appendix 1) was used to assign age based on length, assuming a 1 July birthday. 

Year class was assigned by subtracting (integer) age from Ridley year.

To generate catch curves based on cohorts, observations were weighted to adjust for the

potential interannual variability in strandings based on variable effort. A weighting factor for

each year was calculated as the ratio of strandings in year x /strandings in 1991, the year of the

lowest strandings, and each observation was transformed by dividing by this weighting factor. 

Thus, one individual stranding in year 1991 would be reflected as a 1, while an individual

stranding in other years would be reflected as a number less than 1, with larger reductions in

observations from years with high total strandings.

Weighted observations were then categorized by age and Ridley year, natural log

transformed (ln), and a catch curve fitted for each cohort. The catch curve began with the age

corresponding to the highest left-most point in the ln(frequency)-age plot, and continued to

include the oldest age before an inflection point was detected by eye. For most plots, the

regression line fit ages 2-6. Older age classes had low and variable catches, so we decided to

apply the calculated mortality rate to ages 2-6 only and apply the adult mortality rate (Za) to

immature turtles aged 7 - (a-1).

A similar set of catch curve analyses was generated using standing stock year, with

weighting by hatchling production for a cohort, instead of weighting by strandings levels by

standing stock year. These analyses gave qualitatively similar results but were considered less

reliable because of variance in strandings levels. Overall, the range of Zb was 0.3 to 0.8, with a

mode of 0.5.

As outlined in the Model Form section above, pelagic immature and large benthic

immature and adult instantaneous mortality rates (Zp

 and Za

) were obtained through least-squares

fit of the number of nests predicted by the model and observed nests. In general, pelagic

immature mortality correlated negatively with benthic immature mortality to achieve the

"correct" number of nesting females (see sensitivity analysis below).

Model Results

Model fits to observed nests

We considered 4 factors in the model fitting exercise:

1) age at first reproduction

2) which nesting years to include

3) small benthic immature mortality

4) presence or absence of the post-1990 multiplier.

The lowest sums-of-squares for 1978-1999 were naturally for models fit to the entire time

series, but often these models had poor fits for earlier years (Table 8). The lowest sum-of-squared
NOAA Tech. Mem. NMFS-SEFSC-444 15

residuals for models fit to fewer years were: 1978-1998 for the 8 years to maturity model, 1978-

1996 for the 10 years to maturity model, and 1978-1995 for the 12 years to maturity model. We

used the mortality rates fitted for these years for the model projections. The number of years used

to fit the model primarily affected the post-1990 multiplier, f, which ranged from 0.4 to 0.7, and

the model's future exponential rate of increase, which ranged from 12-17% per year. Without the

post-1990 reduction in mortality for benthic immature and adult turtles, the model cannot fit the

observed trend in nests, regardless of age at maturity or the number of years used to fit the model

(Figure 7, example is for 10 years to maturity, Zb = 0.5, model fit to nests observed from 1978-

1996). 

For simplicity, Figures 8(a, b, and c) show only the best fits and residuals for models with

8, 10 or 12 years to maturity, respectively, and small immature Zb = 0.5. All models shown

include the post-1990 multiplier. The best fit parameters for pelagic immature mortality (Zp

),

large immature/adult mortality (Za), and the post-1990 multiplier are different for each age at

maturity (Table 9). Recall that the fitting algorithm chooses these three parameters to minimize

the sum-of-squares difference between the model nest count and the observed nest count. In

order to achieve the rapid rate of increase observed in nest counts over the past decade, the age

12 model requires a lower post-1990 multiplier, corresponding to a more drastic reduction (50%)

in the mortality rate of benthic feeding juveniles and adults. The resulting survival rates and, in

turn, the population growth rate after 1990, are higher in the age 12 to maturity model (see Bestfit Model Projections).

Sensitivity of model results to changes in small benthic immature Zb

Changes in small benthic immature mortality have virtually no effect on the model fits,

because increases or decreases in one mortality rate are compensated for by the others. To

illustrate this, we plotted the pelagic immature, small benthic immature and large benthic

immature instantaneous mortality rates given by the best fit models for small benthic immature

Zb (pre-1990) ranging from 0.3 to 0.8 (Figure 9). While most of the compensation occurs as a

trade-off between mortality in the pelagic phase and mortality in the subsequent small benthic

immature stage, there is also an increase in the post-1990 multiplier with increasing small

benthic immature mortality (Figure 10). Recall that the multiplier is a proportional change in

mortality, so higher multipliers result in smaller changes in annual survival pre- and post-1990. 

The mortality compensation results in a very similar predicted number of nests, but very different

population size estimates (Figure 11). Because we are trying to fit three parameters with the

current model, there are a large number of feasible solutions to the mortality schedule. Thus,

until empirical estimates can be derived for the post-1990 multiplier and pelagic immature, large

benthic immature and adult mortality, it would be imprudent to use this model to estimate

population size.

Model fits with an empirically-derived estimate of f

We estimated the post-1990 change in small benthic immature mortality by comparing

the strandings catch curves for cohorts that stranded before and after 1990. If pre- and post-1990
16 NOAA Tech. Mem. NMFS-SEFSC-444

estimates of Zb are included in the model, (Zb = 0.52 and 0.34, respectively), the fit and model

outputs are similar to the small benthic immature mortality = 0.5 run with 10 years to maturity

(Figure 12 a). The model predicted a greater post-1990 reduction in large immature and adult

mortality (f = 0.52) than the reduction estimated empirically by the catch curve coefficients of

small benthic immatures (0.34/0.52 = 0.65). It is not known if this is the result of imprecision of

Zb estimates or if there is a larger post-1990 mortality decrease for larger size classes of turtles. 

When the post-1990 multiplier was fixed at 0.65, the model had a lower fit and biased residuals

(sum-of-squared nest residuals 1978-1999 = 791,064; Figure 12 b).

Best-fit model projections

In spite of the large differences in total population size, the model runs with small

benthic immature Zb

 ranging from 0.3 to 0.8 showed a relatively small range of predicted

population growth rates, measured as annual proportional increase and assuming that mortality

rates are constant through time (range 12 – 16% per year). In general, models with lower Zb (=

higher annual survival for 2-6 year-olds) predicted slower population growth rates (these models

also had less reduction in mortality after 1990). Because of their similar post-1990 survival rates,

the age 8 at maturity and age 10 at maturity models showed similar population trajectories

(Figure 13). The age 12 at maturity model, with its higher post-1990 survival rate estimates,

predicted a faster growth rate despite the later age at first reproduction. The time required to

reach 10,000 nesting females varied somewhat in the sensitivity analysis (range 12 – 16 years for

the age at maturity = 10 years model). A reduction in hatch success as the nest population

increased greatly reduced the expected population growth rate for all of the models, but not until

after the 10,000 nesting female mark was reached (example shown is 10 years to maturity, pre1990 small benthic immature Zb

 = 0.5; Figure 14). These model projections are for comparative

purposes only - they are not meant to forecast population size through time. The models do not

include environmental variance and we do not know which, if any, of our models most accurately

matches the mortality schedule and life cycle of the Kemp’s ridley. As more information on

survival and growth is obtained from field studies, we will be able to reduce the number of

unknowns and produce more predictive models.

Literature cited

Caillouet, C. W., Jr., Fontaine, C. T., Manzella-Tirpak, S. A., and Williams, T. D. 1995. 

Growth of head-started Kemp's ridley sea turtles (Lepidochelys kempii) following release. 

Chelonian Conserv. Biol. 1:231-234.

Carr, A. and D. K. Caldwell. 1956. The ecology and migrations of sea turtles: 1. Results of field

work in Florida, 1955. Amer. Mus. Nov. 1793:1-23.

Collard, S. B. and Ogren, L. H. 1990. Dispersal scenarios for pelagic post-hatchling sea turtles.

Bull. Mar. Sci. 47:233-243.
NOAA Tech. Mem. NMFS-SEFSC-444 17

Coyne, M. S. and A. M. Landry. 2000. Plasma testosterone sexing criteria and sex ratio of the

Kemp’s ridley sea turtle (Lepidochelys kempii). U.S. Dep. Commer. NOAA Tech. Mem.

NMFS-SEFSC-436:286

Foote, J. J., N.C. Park, and J. M. Sprinkle. 1998. An increase in marine turtle deaths along the

central coast of Florida (1995-1996); is red tide the culprit? U.S. Dep. Commer. NOAA

Tech. Mem. NMFS-SEFSC-415:169-171.

Frazer, N. B. 1989. Nesting cycles in sea turtles: typical, but not cycles. U.S. Dep. Commer.

NOAA Tech. Mem. NMFS-SEFC-232:61-64.

Gregory, L. F. and J. R. Schmid. in review. Stress responses and sex ratio of wild Kemp’s ridley

turtles (Lepidochelys kempi) in the northeastern Gulf of Mexico. Gen. Comp. Endocrinol.

Gulf & South Atlantic Fisheries Development Foundation, Inc. 1998. Alternatives to TEDS:

final report. Report to NOAA, Contract 50WCNF606083, Tampa, Fla., 19 pp.

Henwood, T. A. and L. H. Ogren. 1987. Distribution and migrations of immature Kemp's ridley

turtles (Lepidochelys kempi) and green turtles (Chelonia mydas) off Florida, Georgia, and

South Carolina. Northeast Gulf Sci. 9:153-159.

Henwood, T. A. and W. E. Stuntz. 1987. Analysis of sea turtle captures and mortalities during

commercial shrimp trawling. Fish. Bull. 85:813-817.

Heppell, S. S., L. B. Crowder and T. R. Menzel. 1999. Life table analysis of long-lived marine

species with implications for conservation and management. Amer. Fisheries Soc. Symp.

23:137-148.

Hildebrand, H. H. 1963. Hallazgo del area de anidacion de la tortuga marina "lora"

Lepidochelys kempi (Garman) en la costa occidental del Golfo de México. Ciencia 

22:105-112.

Jamir, J. V. C. 1999. Revisions to the estimates of incidental sea turtle capture aboard

commercial shrimp trawling vessels. Unpublished report. Gulf & South Atlantic Fisheries

Foundation, Inc., Tampa, Fla. 17 pp.

Márquez-M, R. 1972. Resultados preliminares sobre edad y crecimiento de la tortuga lora,

Lepidochelys kempi (Garman). Mem. IV Congd. Nac. Ocean. (México):419-427.

Márquez-M., R. 1994. Synopsis of biological data on the Kemp’s ridley sea turtle, Lepidochelys

kempi (Garman, 1880). U.S. Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC-343, 91

pp.
18 NOAA Tech. Mem. NMFS-SEFSC-444

Márquez-M., R., A. Villanueva-O., and P. M. Burchfield. 1989. Nesting population and

production of hatchlings of Kemp's ridley sea turtle at Rancho Nuevo, Tamaulipas,

México, p. 16-19. In: Caillouet, C. W., Jr. and A. M. Landry, Jr. (Eds.), Proceedings of

the First International Symposium on Kemp's Ridley Sea Turtle Biology, Conservation

and Management, Texas A & M University, Sea Grant College Program,

TAMU-SG-89-105.

Ogren, L. H. 1989. Distribution of juvenile and subadult Kemp's ridley turtles: results from the

1984-1987 surveys. P. 116-123. In: Caillouet, C. W., Jr. and Landry, A. M., Jr. (Eds.),

Proceedings of the First International Symposium on Kemp's Ridley Sea Turtle Biology,

Conservation and Management. Texas A&M University Sea Grant College Program,

TAMU-SG-89-105.

Pritchard, P. C. H. 1990. Kemp's ridleys are rarer than we thought. Marine Turtle Newsletter

49:1-3.

Reynolds, D. P. and S. S. Sadove. 1998. Wild captures of sea turtles in New York: species

composition shift examined. U.S. Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC415:254-256.

Rostal, D. C. 1991. The reproductive behavior and physiology of the Kemp's ridley sea turtle,

Lepidochelys kempi (Garman, 1880). Ph.D. Diss., Texas A&M University, College

Station, Tx., 138 pp.

Schmid, J. R. 1995. Marine turtle populations on the east-central coast of Florida: results of

tagging studies at Cape Canaveral, Florida, 1986-1991. Fish. Bull., U.S. 93:139-151.

Schmid, J. R. 1998. Marine turtle populations on the west-central coast of Florida: results of

tagging studies at the Cedar Keys, Florida, 1986-1995. Fish. Bull. 96:589-602.

Schmid, J. R. and W. N. Witzell. 1997. Age and growth of wild Kemp’s ridley turtles

(Lepidochelys kempi): cumulative results of tagging studies in Florida. Chelonian

Conserv. Biol. 2:532-537.

Shaver, D.J. and C.W. Caillouet, Jr. 1998. More Kemp’s Ridley turtles return to south Texas to

nest. Marine Turtle Newsletter 82:1-5.

Turtle Expert Working Group. 1998. An assessment of the Kemp's ridley (Lepidochelys kempii)

and loggerhead (Caretta caretta) sea turtle populations in the Western North Atlantic.

U.S. Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC-409, 96 pp.

U.S. Fish and Wildlife Service and National Marine Fisheries Service. 1992. Recovery plan for

the Kemp's ridley sea turtle (Lepidochelys kempii). National Marine Fisheries Service, St.

Petersburg, Fla., 40 pp.
NOAA Tech. Mem. NMFS-SEFSC-444 19

Witzell, W. N. 1998. Long-term tag returns from juvenile Kemp's ridley turtles. Marine Turtle

Newsletter 79:20.

Woody, J. B. 1985. International efforts in the conservation and management of the Kemp’s

ridley sea turtle (Lepidochelys kempi). p. 1-3. In: C. Caillouet and A. Landry (Eds.),

Proceedings of the First International Symposium of Kemp’s Ridley Sea Turtle Biology,

Conservation and Management. Texas A&M University at Galveston, TAMU-SG-89-

105.

Zug, G. R., Kalb, H. J., and Luzar S. J. 1997. Age and growth on wild Kemp’s ridley sea turtles

Lepidochelys kempii from skeletochronological data. Conserv. Biol. 80:261-268.
20 NOAA Tech. Mem. NMFS-SEFSC-444

Table 1. Nests counted at Rancho Nuevo, North Camp and South Camp 1966 - 1998. Nests and

hatchlings released are approximate for 1966 - 1977. Total nests includes estimates (1966 –

1977) or counts (1978 – 1998) of nests not moved to protective corrals. Total hatchlings =

hatchlings released from corrals only. **Hatchling count not yet determined for 1999.

Year Rancho

Nuevo

North Camp South Camp Total Nests Total

Hatchlings

Released

Hatchlings

Released per

Nest

1966 5991 5991 29100 4.9

1967 5519 5519 24100 4.4

1968 5117 5117 15000 2.9

1969 4018 4018 28400 7.1

1970 3017 3017 31400 10.4

1971 2012 2012 13100 6.5

1972 1824 1824 14600 8.0

1973 1642 1642 23500 14.3

1974 1466 1466 23500 16.0

1975 1266 1266 11100 8.8

1976 1100 1100 36100 32.8

1977 1036 1036 30100 29.1

1978 1012 1012 48009 47.4

1979 1019 1019 63996 62.8

1980 900 900 37378 41.5

1981 962 962 53282 55.4

1982 842 842 48007 57.0

1983 795 795 32921 41.4

1984 861 861 58124 67.5

1985 740 740 51033 69.0

1986 752 752 48818 64.9

1987 742 742 44634 60.2

1988 840 0 3 843 62218 73.8

1989 905 11 0 916 66802 72.9

1990 800 204 0 1004 74339 74.0

1991 862 178 74 1114 79749 71.6

1992 932 254 84 1270 92116 72.5

1993 847 251 101 1199 84605 70.6

1994 1126 329 110 1565 107687 68.8

1995 1438 334 149 1921 120038 62.5

1996 1255 444 219 1918 114842 59.9

1997 1514 358 347 2219 141770 63.9

1998 2409 431 642 3482 167105 48.0

1999 2308 273 776 3357 ** **
NOAA Tech. Mem. NMFS-SEFSC-444 21

Table 2. Small turtle strandings (20-30 cm SCL) and hatchlings released 1 - 4 years prior to

stranding year, 1986–1997.

 Hatchlings released

Year Strandings 1 year earlier 2 years earlier 3 years 4 years

1986 105 51033 58124 32921 48007

1987 30 48818 51033 58124 32921

1988 24 44634 48818 51033 58124

1989 27 62218 44634 48818 51033

1990 28 66802 62218 44634 48818

1991 25 74339 66802 62218 44634

1992 24 79749 74339 66802 62218

1993 94 92116 79749 74339 66802

1994 121 84605 92116 79749 74339

1995 40 107687 84605 92116 79749

1996 41 120038 107687 84605 92116

1997 38 114842 120038 107687 84605

Correlation 0.2603 0.4047 0.1419 0.2833

Table 3. Kemp’s ridley strandings, 1986-1997.

1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 Total

N.E. U.S. 27 12 24 10 11 12 20 25 25 16 14 16 212

S.E. U.S. 32 51 116 43 94 48 30 79 95 91 114 92 885

Gulf of

Mexico

246 98 75 96 153 105 106 223 430 288 255 304 2379

Total 305 161 215 149 258 165 156 327 550 395 383 412 3476
22 NOAA Tech. Mem. NMFS-SEFSC-444

Table 4. Kemp’s ridley strandings trends. Percent of total strandings represented by each

location is shown in parentheses.

 1986-88 total 1995-97 total Percent change

N.E. U.S. 63 (9%) 46 (4%) -27%

S.E. U.S. 199 (29%) 297 (25%) +49%

Gulf of Mexico 419 (62%) 847 (71%) +102%

Total 681 1190 0.75

Table 5. Life stages defined in the age-based model for Kemp’s ridleys.

 

Life stage Component ages

 

Pelagic immature 0 through 1

 

Small benthic immature 2 through 6

 

Large benthic immature 7 to ( -1) (e.g., 9 if α=10)

 

Adult + (e.g.,10+)
NOAA Tech. Mem. NMFS-SEFSC-444 23

Table 6. Mortality rate parameters defined for the age-based model.

Stage Pre- 1990 Post-1990

Pelagic immature Zp Zp

Small benthic immature

(age 2-6) Zbf*Zb

Large benthic immature Zaf*Za

Adult Zaf*Za

Table 7. Hypothetical changes in hatching rate expected with increasing numbers of nests, as

applied to density-dependent projections of the population model.

Range in nests Protection regime % survival to

hatchling release

0-5,000 Corral 65

5,001-6,000 in situ with screening 50

6,001-10,000 in situ with predator control 30

>10,000 in situ with no predator

control

20
24 NOAA Tech. Mem. NMFS-SEFSC-444

Table 8. Sum of squared residuals (observed - expected nests) for 3 models fit to various time

series. Small benthic immature Zb

 = 0.5.

8 years to maturity

fit to: model fit yrs

(1978-199x)

1978-1995 1996-1999 1978-1999

1995 126,679 126,670 2,149,482 2,276,151

1996 209,839 185,056 429,312 614,368

1997 235,948 221,965 712,188 934,153

1998 526,370 166,135 428,933 595,069

1999 561,440 171,175 390,264 561,440

10 years to maturity

fit to: model fit yrs

(1978-199x)

1978-1995 1996-1999 1978-1999

1995 60,823 60,823 1,354,351 1,415,174

1996 86,669 77,980 360,795 438,776

1997 89,642 85,029 435,837 520,867

1998 214,790 79,748 742,805 822,554

1999 429,941 73,829 356,111 429,941

12 years to maturity

fit to: model fit yrs

(1978-199x)

1978-1995 1996-1999 1978-1999

1995 161,174 161,174 589,762 750,937

1996 165,800 163,847 923,858 1,087,706

1997 166,084 164,770 997,803 1,162,573

1998 447,024 208,158 720,433 928,592

1999 578,015 165,934 412,081 578,015
NOAA Tech. Mem. NMFS-SEFSC-444 25

Table 9. Parameters for model best-fits with small benthic immature z=0.5.

Age at maturity

Pelagic immature

annual survival

Large benthic immature

and adult annual survival

(age 7+)

Post-1990

multiplier (f)

pre-1990 post-1990 pre-1990 post-1990

8 0.24 0.24 0.85 0.91 0.56

12 0.37 0.37 0.86 0.92 0.50

10 0.31 0.31 0.85 0.91 0.56

10 - w/out post1990 multiplier0.38 0.38 0.83 0.83 1.0
NOAA Tech. Mem. NMFS-SEFSC-444 26Figure 1. Nesting beaches of the Kemp’s ridley sea turtle. Primary nesting beach is Rancho Nuevo
NOAA Tech. Mem. NMFS-SEFSC-444 27 Figure 2. Nests and hatchlings released from Rancho Nuevo, North Camp and South Camp, 1966-1999 (h

are approximate).
NOAA Tech. Mem. NMFS-SEFSC-444 28Figure 3. Nests counted at Rancho Nuevo, North Camp and South Camp 1989-1998.
NOAA Tech. Mem. NMFS-SEFSC-444 29 Figure 4. Kemp's ridley nest projection if current trends continue.
30 NOAA Tech. Mem. NMFS-SEFSC-444

Figure 5. Scatter plots of nests counted vs. hatchlings released 8, 10 or 12 years later. Numbers

denote nesting year.
NOAA Tech. Mem. NMFS-SEFSC-444 31 Figure 6. Size frequency distributions of stranded benthic immature Kemp’s ridleys in 1986-1987 and 1996

across all statistical zones (Appendix 2).
NOAA Tech. Mem. NMFS-SEFSC-444 32Figure 7. Best fit model estimates of nests for 10 years to maturity, Zb = 0.5, with and without the post-199

fitted parameter.
NOAA Tech. Mem. NMFS-SEFSC-444 33

Figure 8. Model fits for 8, 10 or 12 years to maturity with post-1990 multiplier included.

 

 
NOAA Tech. Mem. NMFS-SEFSC-444 34Figure 9. Changes in best-fit parameter estimates for pelagic immature and large turtle instantaneous mo

the small benthic immature mortality rate, estimated from a catch curve of strandings.
NOAA Tech. Mem. NMFS-SEFSC-444 35 Figure 10. Effect of small benthic immature mortality estimate on the post-1990 multiplier.
NOAA Tech. Mem. NMFS-SEFSC-444 36Figure 11. Model estimated population size in 1998, 10 years to maturity model with a range of small ben

survival rates. Small benthic immatures should be read with the right axis.
NOAA Tech. Mem. NMFS-SEFSC-444 37 Figure 12. Model fits and residuals for pre-1990 small benthic immature 

Zb = 0.53, post-1990 

Zb = 0.34. 

fit post-1990 multiplier for large benthic immature and adult turtles f = 0.52. B. Post-1990 multiplier 

f f

benthic immatures and adults. 
NOAA Tech. Mem. NMFS-SEFSC-444 38Figure 13. Model projections with Zb = 0.5, assuming mortality rates remain constant after 1990.
NOAA Tech. Mem. NMFS-SEFSC-444 39 Figure 14. Model projection of nesting female numbers with and without a reduction in nest surviva

years to maturity, 

Zb = 0.5).
40 NOAA Tech. Mem. NMFS-SEFSC-444

Part 2. Loggerhead Status Report

Summary

The loggerhead sea turtle was listed under the Endangered Species Act as threatened in

1978. At the time of listing the species in the North Atlantic was considered to be a single

population, but genetic evidence acquired since has indicated the existence of distinct

subpopulations which form the basis of management units. The trends of each subpopulation

must be assessed separately and the subpopulations may need to be listed separately under the

ESA to afford the smaller or declining subpopulations more protection. The TEWG recommends

that the life history stages of each loggerhead subpopulation be monitored, and acknowledges

that the mix of subpopulations on the foraging grounds makes population-specific assessments

complicated. In-water surveys must be conducted to provide data to evaluate the changes in the

age structure of the subpopulations. 

Stock Definition and Geographic Range

The loggerhead sea turtle (Caretta caretta) occurs throughout the temperate and tropical

regions of the Atlantic, Pacific, and Indian Oceans (Dodd, 1988). Its range of habitat includes

open ocean waters, continental shelves, bays, lagoons, and estuaries. Loggerheads in the

Western North Atlantic nest on high-energy, sandy beaches between the latitudes of 18o and 35o

North. Worldwide, nesting on warm temperate beaches is much more common than nesting in

the tropics. In the western North Atlantic there are at least 4 genetically distinct loggerhead

nesting subpopulations, based on mitochondrial DNA (Bowen et al., 1993; Bowen, 1995;

Encalada et al., 1998) (Figure 15) and possibly there are more (Francisco et al., 1999):

1. The Northern Nesting Subpopulation, occurring from North Carolina to northeast

Florida, about 29o N. (approx. 7500 nests in 1998). 

2. The South Florida Nesting Subpopulation, occurring from 29o N on the east coast to

Sarasota on the west coast (approx. 83,400 nests in 1998). This is the largest loggerhead

nesting assemblage in the Atlantic and is the second largest in the world (Ross, 1982).

3. The Florida Panhandle Nesting Subpopulation, occurring at Eglin Air Force Base and

the beaches near Panama City, FL (approx. 1200 nests in 1998).

4. TheYucatán Nesting Subpopulation, occurring on the eastern Yucatán Peninsula,

Mexico (Márquez-M, 1990) (approx. 1000 nests in 1998). 

Historically, nesting has been minor elsewhere in the western North Atlantic, except in

Honduras, Nicaragua, Columbia, Venezuela, Bahamas, Cuba, Jamaica, and Puerto Rico

(Sternberg, 1981). The present size of these nesting assemblages is unknown and individuals

from these nesting areas have not been assayed for genetic affinity. As a whole, nesting by the
NOAA Tech. Mem. NMFS-SEFSC-444 41

loggerhead subpopulations of the northwest Atlantic account for more than 35-40% of the global

nesting activity of this species (Ross, 1982).

 Cheloniid sea turtle species worldwide have highly structured nesting assemblages

(Bowen et al., 1992; Bass, 1994; Bowen et al., 1994; Broderick et al., 1994). Maternally

mediated gene flow based upon mtDNA analysis among identified populations is low (Bowen et

al., 1993; Bowen, 1995). Natal homing is a dominant force shaping this phylogeographic

structure. The nesting assemblages probably represent distinct genetic entities, but other

research, including analysis of nuclear DNA (inherited from both parents) is necessary to address

stock definition. Analysis of nuclear DNA of green turtles in both the Pacific and in the Atlantic

found moderate rates of male-mediated gene flow (Karl et al., 1992), but still, populations within

ocean basins were structured. Nevertheless, given the high site fidelity of nesting females to

their natal beach and low gene flow between nesting assemblages, most Western North Atlantic

loggerhead nesting assemblages are vulnerable to extirpation. Should an assemblage be

extirpated, regional dispersal will not be sufficient to replenish the depleted nesting assemblage

within thousands of years. This has been demonstrated amply through the loss of assemblages

worldwide; they have not been re-established (e.g. Bermuda, Cayman Island, Alto Velo). Thus,

both genetic evidence (mtDNA) and the historical record indicate that recolonization of

extirpated nesting assemblages does not occur on contemporary time scales (Bowen et al., 1994). 

 For conservation and management, mtDNA data indicate that nesting assemblages must be

considered independent demographically. This conclusion holds even if males should prove not

to be philopatric to their natal site, because the production of progeny depends on female nesting

success. For this reason, we use the term "subpopulation" to describe the nesting assemblages,

consistent with the IUCN definition of the term.

Like other sea turtles, loggerhead post-hatchlings occur in pelagic habitats for a number

of years, eventually settling into coastal habitats as benthic immatures, at approximately 40-60

cm SCL in the U.S. (Carr, 1987). Genetic analyses indicate that turtles from several

subpopulations mix on foraging grounds. Table 10 summarizes the estimated proportional

contributions of the four identified subpopulations to foraging-ground assemblages, as well as the

proportional contributions of Mediterranean haplotypes. The proportional contributions from

Chesapeake Bay, South Carolina, Georgia (in part), and the Mediterranean were determined prior

to the completion of a more thorough survey of Atlantic nesting beaches and may therefore

change by 10% or more, but the qualitative conclusion of mixtures on these foraging grounds

remains valid (B. Bowen, personal communication). Because South Florida is the largest

Nesting Subpopulation contributing to these foraging grounds, the samples are often dominated

by turtles of this haplotype. However, the results of these analyses, conducted at multiple

foraging sites along the Atlantic seaboard, indicate that mixing of subpopulations is not random

and that the contributions of the larger South Florida Subpopulation increase along a general

north-south gradient (northeast U.S. to Florida Bay). Analyses of the genetic identities of

loggerhead turtles inhabiting foraging grounds of the Gulf of Mexico are underway but not yet

available. Preliminary results from analysis of loggerheads stranding in Texas indicate the

Northern Subpopulation may account for up to 10% of the animals stranding there, but most
1Appendix 4 of our first report (TEWG, 1998) details how we arrived at the decision to

use 92 cm SCL to distinguish benthic immature and adult turtles. It was based on an arithmetic

mean which may be an inappropriate measure of central tendency in non-normally distributed

data, which some nesting beach data appear to be. Also, we have not considered the possibility

that the average size of neophyte nesters may be subpopulation-specific.

42 NOAA Tech. Mem. NMFS-SEFSC-444

come from the S. Florida Subpopulation, along with some from the Yucatán, Mexico

Subpopulation (A. Bass, personal communication).

Life stage definitions

We have defined the life history stages for loggerhead turtles as: pelagic immature from

the time a hatchling leaves the nesting beach until it appears on coastal benthic feeding grounds

at around 40-60 cm straight carapace length (SCL); benthic immature from settlement to 92 cm

SCL; and adult which is 92 cm SCL1. Pelagic immature turtles originating from Atlantic

nesting beaches have been observed in the eastern Atlantic near the Azores and Madeira

(Maigret, 1983; Carr, 1986, 1987; Bjorndal et al., 1994; Bolten et al., 1994) and in the

Mediterranean Sea (Laurent et al., 1998). A growth curve derived from a length-frequency

analysis of pelagic immatures in the Azores indicates that loggerheads may stay in the North

Atlantic Gyre for 7-12 years, depending on the size of recruitment to the benthic stage (Bjorndal

et al., 2000). It has been suggested that turtles that originate from Gulf of Mexico beaches also

utilize the North Atlantic Gyre and are distributed to it via the Loop Current of the Gulf of

Mexico and Florida Current (Collard and Ogren, 1990). Strandings records indicate that when

pelagic immature loggerheads reach 40-60 cm SCL they recruit to coastal inshore and nearshore

waters of the continental shelf throughout U.S. Atlantic and Gulf of Mexico. Benthic immatures

have been found from Cape Cod, Massachusetts, to southern Texas, and occasionally strand on

beaches in northeastern Mexico (R. Márquez-M., personal communication). Large benthic

immature loggerheads (70-91 cm) represent a larger proportion of the strandings and in-water

captures (Schroeder et al., 1998) along the south and western coasts of Florida as compared with

rest of the coast, but it is not known whether the larger animals actually are more abundant in

these areas or just more abundant within the area relative to the smaller turtles. Benthic

immature loggerheads foraging in northeastern U.S. waters are known to migrate southward in

the fall as water temperatures cool (Epperly et al., 1995; Keinath, 1993; Morreale and Standora,

1999; Shoop and Kenney, 1992), and likely migrate northward in spring. Given an estimated age

at maturity of 21-35 years (Frazer and Ehrhart, 1985; Frazer et al., 1994), the benthic immature

stage must be at least 10-20 years long. The adult life history stage, 92 cm straight carapace

length, has been reported throughout its U.S. range and throughout the Caribbean Sea. Nesting

occurs primarily from North Carolina southward to Florida with assemblages recorded in the

Florida Panhandle and on the Yucatán Peninsula (Figure 15). Non-nesting females are reported

from the entire area and little is known about the distribution of adult males which are seasonally

abundant near nesting beaches during the nesting season.
NOAA Tech. Mem. NMFS-SEFSC-444 43

Status and Trends

Nesting beaches

Northern Subpopulation

The addition of nesting data from the years 1996, 1997, and 1998, did not change the

assessment of TEWG that the number of loggerhead nests in the Northern Subpopulation is

stable or declining (Figures 16 and 17). The annual number of recorded nests for the period

1989-1998 ranged 4370-7887 nests/year (Table 11). No trends are detectable for North Carolina,

South Carolina or Georgia during that period; nests in northeast Florida may be increasing,

although the data are too variable to detect a significant trend (Table 12). A longer time series of

nesting data for a portion of the Northern Subpopulation comes from Cape Island at the Cape

Romain National Wildlife Refuge, South Carolina (Figure 16B). There, the number of recorded

nests for the period 1975-1998 ranged 579-2654 nests/year. Nesting at Cape Island has declined

an average of -2.7% per year over the 24-year period (r-square=0.284, P = 0.007), but nearly all

of that decline occurred during the 1970's. Since 1989, the Cape Island nest numbers have varied

annually with no significant trend (1989-1998 slope = -1.4% per year, r-square = 0.023, P =

0.675). Annual nesting at Cape Island was found to correlate well with nesting in North Carolina,

South Carolina (remaining beaches), Georgia and the Northern Subpopulation as a whole for the

period 1989-1998, but did not correlate as well with nesting activity of the South Florida and

Florida Panhandle Subpopulations (Table 13). The residuals from the regression model

correlated well with the residuals from northern nesting beaches and southeast Florida, but not

the SW Florida nor the Panhandle residuals.

South Florida Subpopulation

The addition of nesting data from the years 1996 through 1998 did not change the

assessment of TEWG that the number of loggerhead nests in the South Florida Subpopulation is

stable or increasing. The annual number of recorded nests for the period 1989-1998 ranged

48,531-83,442 nests/year (Table 11), with an average rate of increase over the time series of

3.6% per year (r-square = 0.461, P = 0.031, Table 12). The time series suggests a 4-year cycle in

nesting, although the time series is too short to confirm this (Figure 17A). A longer time series of

nesting data for a portion of the South Florida Subpopulation comes from Hutchinson Island,

Florida (Figure 17B). There, the number of recorded nests for the period 1981-1998 ranged

3121-8214 nests/year and had a mean rate of increase of 4% per year (r-square = 0.746, P <

0.001). In more recent years (1989-1998), Hutchinson Island nests have increased 2.7% per year

but the trend is not significant because of data variability (r-square = 0.308, P = 0.096). Variation

in nest counts at Hutchinson Island seem to support the 4-year cycle hypothesis for recent years

but not earlier years in the time series (Figure 17B). The nest count was found to correlate

extremely well with nesting in the remaining parts of the southeast Florida region for the period

1989-1998, but did not correlate as well with the nests counted at northern beaches nor those in

the Florida Panhandle (Table 13). However, the residuals from the ln-transformed regression

correlated fairly well with the residuals from all areas except the Panhandle.
44 NOAA Tech. Mem. NMFS-SEFSC-444

Florida Panhandle Subpopulation

The number of loggerhead nests recorded for the Florida Panhandle Subpopulation is

increasing, but much of this increase is due to expanded beach monitoring. The length of beach

monitored in the Panhandle increased from 186.5 km in 1989 to 359.9 km in 1997. Thus, the

status of this Subpopulation cannot be determined at this time. 

Nest count correlations across states and regions

Table 14 gives the Pearson correlation coefficients for loggerhead nest counts in all U.S.

areas. A second set of correlations was calculated with ln-transformed data residuals to remove

the linear trend of each time series. State nest counts within a nesting subpopulation generally

correlate well with each other and the subpopulation totals. However, the three nesting

subpopulations are less well correlated with each other over the 10 year time period. Nests in

North Carolina and Georgia may follow the 4-year pattern observed in southeast Florida (Figures

16 and 17), but South Carolina nests are much more variable. The Florida Panhandle

Subpopulation is poorly correlated with the other two subpopulations, likely because much of the

recent increase in nest counts for that area are due to an increase in monitoring effort. However,

the Panhandle nests and SW Florida nests are highly correlated.

Yucatán nesting subpopulation

There are few nesting surveys for loggerheads in Mexico. There are 8-12 turtle camps

operated each year in central Quintana Roo on the Yucatán Peninsula. In 1998, 1052 nests were

recorded at 12 camps on Quintana Roo beaches (Xcaret, 1999). Nesting of this Subpopulation is

believed to extend into Belize and nesting appears to be stable or increasing (R. Márquez-M.,

personal communication).

Loggerhead nesting elsewhere in the Gulf of Mexico

Nesting of loggerheads throughout the Gulf of Mexico is sporadic outside of Florida and

few areas conduct systematic surveys of nesting beaches (S. MacPherson, personal

communication). The first estimate of nesting in the Gulf of Mexico outside of Florida indicated

that between 1960 and 1962 a reproductive assemblage of approximately 100 loggerheads nested

on the remote beaches of the offshore islands of Louisiana, Mississippi, and Alabama (Ogren,

1978). Aerial surveys in 1977 of the U.S. Gulf beaches west of Florida indicated that nesting had

declined (Ibid). In recent years about 21-31 nests were found annually at Bon Secour National

Wildlife Refuge in Alabama (1994-1998) and a couple are found each year on Dauphin Island

(1995-1997) (Mays and Shaver, 1998; S. MacPherson, personal communication). Up to 9 nests

were found annually at the Gulf Islands National Seashore in Mississippi 1990-1998 and less

than 10 nests were found annually at the Breton National Wildlife Refuge in Louisiana (1989-

1990) (Ibid). There are few records of loggerheads nesting on the Texas Coast: 18 since 1977

(D. Shaver, personal communication). Eleven of the 18 recorded nests occurred in 1996-1998

and all but one was on the southern Texas coast. Loggerheads also nest occasionally in the
NOAA Tech. Mem. NMFS-SEFSC-444 45

Mexican states of Tamaulipas, Campeche, and Yucatán. During this decade the annual nest

counts for these areas were: 1-5 nests at Rancho Nuevo and Ostionales in Tamaulipas, about 50

nests in Campeche, and about 100 nests in Yucatán (R. Márquez-M., personal communication). 

The genetic identity of these scattered nests is unknown. 

Strandings

Complete strandings information by size class, statistical zone and year are provided in

Appendix 2. As with the analysis of strandings of Kemp’s ridley turtles, the loggerhead

strandings used by the TEWG excluded incidental captures, post-hatchlings, cold-stunned or

head-started animals. Figure A2.1 in Appendix 2 shows the statistical zones for which sea turtle

strandings are reported. Monitoring effort is not directly comparable between zones but has been

reasonably consistent over this period. Figure A2.3 a, b, and c depicts the loggerhead strandings

reported by statistical zone, by year, 1986-1997. There is no survey effort in zones 15 and 16,

due to inaccessibility of shoreline, and coverage is low in zones 13 and 14. In the eastern Gulf of

Mexico (zones 1-12, partial 24-25), survey coverage is low in zones 1, 3, 6, and 7 due to

inaccessibility and zone 2 has very little land mass. The lack of data from these zones may or

may not reflect a lack of strandings. Along the southeast U.S. Atlantic coast, coverage is also

low in zones 24 and 25. In the northeastern U.S. Atlantic, survey coverage is less rigorous. 

However, high human densities along the coast in this area suggests most strandings will get

reported. This is not true for inshore waters, such as the Chesapeake Bay and Pamlico and Core

Sounds of North Carolina, where strandings likely go unreported.

Trends

Table 15 shows the loggerhead strandings by region. Over the 12-year period 1986-97,

the southeast accounted for 63% of the strandings totals, with the Gulf of Mexico and northeast

Atlantic accounting for 21% and 15%, respectively. Strandings in the southeastern U.S. and the

Gulf of Mexico declined in the early 90's, particularly 91-93, but have increased since then. The

combined total for 1995-97 was 13% higher than the combined total for 1986-88 (Table 16). In

particular, strandings in the northeast more than doubled over that time frame, increasing the

relative importance of mortality in the northeast (from 10 to 19%). The northeast is the only

region which has shown significant trends throughout the 1986-1997 period. However, since

1991 all three areas have shown significant increasing trends in strandings, 10-14% per year.

Overall, loggerhead strandings showed an average annual increase of 10.5% per year for the

period 1991-1997.

Strandings "hot spots"

While loggerhead strandings occur throughout the eastern U.S., the largest proportion of

them occur along the southeast U.S. Atlantic coast (Figure A2.3, Appendix 2). Strandings are

high throughout the area, with Zone 30 highest in most years and Zone 26 and the southern

segment of Zone 36 lower in most years. Adults strand throughout the region. Strandings are

lower in the Gulf of Mexico. Zones 4 and 5 have somewhat higher stranding numbers; of
46 NOAA Tech. Mem. NMFS-SEFSC-444

particular concern is the proportion of adult loggerheads stranding in these zones. Zones 18, 19

and 20 also have higher strandings but with few adults. In the northeast U.S. Atlantic loggerhead

strandings are concentrated in Zones 36 and 37 with fewer adults.

Based on NMFS estimates shrimping intensity is much higher in the western Gulf of

Mexico than in the east. It is concentrated nearshore in the west; more than half the shrimping

effort is the 0-18m depth zone. In the eastern Gulf, effort peaks in the 18-37m zone (>60% of the

effort occurs at these depths) (Nance, 1992, 1993; McDaniel et al., 2000). Loggerhead

strandings are consistently highest in zones 4-5 (West Florida) and 18-20 (Texas). Strandings in

West Florida include a high proportion of adults whereas strandings in Texas are exclusively

juveniles. Given the relatively low shrimping effort in the eastern Gulf, the impact on

loggerheads (particularly adults) appears high.

Systematic surveys for sea turtle strandings do not occur in Mexico but a few (<10)

loggerhead strandings have been recorded at Rancho Nuevo in recent years. These have been

primarily large benthic immatures (>63 cm CCL), except for one 95 cm CCL animal in 1997.

TED effect analyses 

Biweekly strandings data from South Carolina and Georgia (1980-97) were analyzed to

determine the significance and magnitude of TED effects in reducing strandings (Royle and

Crowder, 1998). The methods used were similar to those applied to the S.C. data (1980-93) in

which strandings were shown to be reduced by 44% (Crowder et al., 1995). TEDS significantly

reduced strandings over the period 1980-97 by an estimated 40% in South Carolina and 58% in

Georgia relative to strandings estimates without TEDs (Royle and Crowder, 1998). A significant

“TED effect” on strandings is detectable through the time series analysis of biweekly data in spite

of the increasing trend in annual strandings. Recent analysis of the Georgia data utilizing

Poisson regression and taking into account shrimp landings indicates strandings in Georgia (per

unit shrimp catch) were reduced 37% when TEDs where in use compared to periods when TEDs

were not in use (Royle, 2000). 

In-water surveys

Fishery independent, in-water studies of sea turtles have been carried out at multiple sites

in the Gulf of Mexico and U.S. Atlantic. The primary goals and objectives of these studies vary,

and the species targeted vary as well. Some studies have been conducted for many years (10+),

others have only recently been initiated and still others have been intermittent or aperiodic. There

are numerous difficulties in designing in-water studies to ensure that standardized catch per unit

effort (CPUE) methods are employed such that data can be compared, not only within a site but

also site-to-site. Environmental conditions including depth strata, current patterns, visibility,

water temperature, and investigator experience can greatly affect catch rates. Biological

conditions, such as prey abundance, complicate sampling strategies that may be based on random

distribution of the target species. Changes in effort, variability in temporal sampling, and

different sampling methodologies make analyses of population trends particularly difficult. At
2Recent examination of SEAMAP trawl data collected by South Carolina Department of

Natural Resources indicates that those data may also be a source of trend information. The

extensive trawl survey is based on a stratified random sampling design in nearshore waters

between Cape Hatteras and central Florida and has been conducted seasonally since 1989.

NOAA Tech. Mem. NMFS-SEFSC-444 47

this time, only two studies enable relatively long-term comparisons of CPUE on a within study

site basis2. These two sites are the nearshore waters off the south central east coast Florida at the

St. Lucie Nuclear Power Plant (Quantum Resources, 1995, unpublished data) and the Indian

River Lagoon along the central east coast of Florida at Sebastian Inlet (Ehrhart et al., 1996). The

TEWG does not feel that changes in CPUE evaluated with recent data from the naked net survey

conducted by the Gulf & South Atlantic Fisheries Development Foundation, Inc. (1998; Jamir,

1999) indicate a real 8-fold increase in population size in the southeast U.S. because of the

differences in sampling as discussed in the Kemp’s report above, but that it does indicate there

has been an increase of an unknown magnitude.

Ehrhart et al. (1996) analyzed loggerhead CPUE data from his central Indian River

Lagoon study site for three combined sampling periods (1983-1985, 1988-1990, and 1993-1995).

The turtles were captured by gill nets strung across the lagoon, with CPUE measured as turtles

caught per kilometer net hour. Ehrhart concluded that although there was some variation

between the three spans of years, CPUE was not statistically different (Kruskal-Wallace = 1.244,

df = 2, P = 0.5369). Captures of loggerheads at the Indian River Lagoon site are immature turtles

and hence the CPUE statistics would not be affected by influxes of adult turtles during the

nesting season.

The intake pipe of the St. Lucie Nuclear Power Plant captures sea turtles each year,

depositing them relatively unharmed in an enclosed lagoon. The species and size of each turtle is

recorded, after which the turtles are tagged and released. Because the intake pipe operates

continuously, the capture effort may be considered relatively constant. Turtles of all size classes

have been caught, but it is not known whether the St. Lucie turtles represent an unbiased sample

of the nearby turtle population. Over the entire time series (1977-1998), the exponential trend in

total loggerheads captured at St. Lucie has a slope of 4.2% per year (r-square = 0.40, P = 0.002)

(Table 17). However, the early part of the time series shows little overall growth and a

potentially cyclical pattern in abundance (Figure 18). In recent years, the trend has increased to

17% per year (r-square = 0.76, P = 0.005), although the number of loggerheads captured in 1997

dropped dramatically. The TEWG (1998) reported significant increases in numbers for adult

loggerheads ( 92 cm SCL) at St. Lucie which track increases in the nesting population, as

determined by nesting surveys. These trends continue when post-1995 data (through 1998) are

included (Figure 18, Table 17). Increased numbers of benthic immature turtles captured in 1996

and 1998 resulted in a significant increasing trend for these turtles as well, which was not

detected in the 1977-1995 time series due to data variance (TEWG, 1998). Adult and immature

captures were generally well correlated across years (Pearson’s r 0.65, P<0.05), without

evidence of a time lag.
48 NOAA Tech. Mem. NMFS-SEFSC-444

The two studies which provide the longest and most consistent time series of data do not

agree in their assessments of trends in the population of benthic immature loggerheads. Because

juvenile abundance and capture probability are highly variable, it will be difficult to detect trends

without long time series from multiple sites. Both of the surveys discussed here consist of

samples that are likely to be dominated by the South Florida Subpopulation, thereby giving little

indication of trends in benthic immature loggerheads originating from the Northern

Subpopulation. A survey of benthic immature turtles in Pamlico and Core Sounds, North

Carolina, will provide more information regarding trends in the different subpopulations. More

comprehensive surveys are needed to monitor the status of benthic stages of the population and

multiple, index area studies that yield statistically valid CPUE data are urgently needed. Trawl

surveys may be one method of accomplishing this goal. The Working Group recommends the

design and implementation of a pilot trawl survey, bearing in mind the patchy nature of

population distribution (see Research Recommendations). 

Sex ratio

Studies of loggerheads along the Atlantic coast suggest that these populations exhibit a

female bias. Laparoscopy of 47 turtles examined during fall 1997 in North Carolina found 68%

were female and 32% were males (Braun-McNeill et al., 2000). Serum androgen sexing of

turtles captured in waters 20 ° C in the same area, 1996-1998 indicated 72% female, 27% male,

and 1% unknown (n=114) (Braun-McNeill et al., in press). Wibbels et al. (1987) reported a

female bias (66%), based on serum testosterone titres, across Atlantic coast sites sampled during

the 1980's: Chesapeake Bay (52%), Cape Canaveral (60%), Hutchinson Island (74%), and Indian

River (58%). It can be speculated that female-biased sex ratios are likely given the

predominance of turtles from the South Florida Nesting Subpopulation, whose nest incubation

temperatures are generally higher than the pivotal temperature required for an even sex ratio in

loggerheads (Mrosovsky, 1988). 

Size distributions

Mean size of stranded loggerheads 1986-87 vs. 1996-97

The mean size of benthic immature loggerheads that stranded from 1986-1997 varied

from region to region with the largest turtles in the eastern Gulf (about 80 cm SCL) and

somewhat smaller turtles elsewhere (about 60-70 cm SCL). One might expect that a growing

population would experience a shift in size distribution towards smaller turtles. We found no

evidence of a shift in the mean size of loggerheads stranded over this period, but the size

distribution of stranded loggerheads was significantly different in 1996-97 compared with

1986-87 (Kolmogorov-Smirnov Test, P<0.0001) (Figure 19). Benthic juvenile strandings in the

55-65 cm SCL size range increased in abundance in the 1990s relative to the 1980s.
NOAA Tech. Mem. NMFS-SEFSC-444 49

Size distributions from the St. Lucie Power Plant

The size composition of loggerheads captured at the St. Lucie nuclear power plant has

varied from year to year with some long-term trends (Figure 20). In the 1980's, the proportion of

small immatures captured decreased, becoming as low as 50% in 1989 and 1990. Since 1991,

this proportion has stabilized at around 70%. The proportion of adult loggerheads in the capture

records reached a peak of 17% in 1990 and has fluctuated since 1991 (5-13.5%). The average

proportions for 1991-1998 are 69% small immatures, 22% large immatures and 9% adults.

Population size estimates 

Nesting data collected on index nesting beaches 1989-1998 represent the best dataset

available to index the population size of loggerhead sea turtles. These data provide annual

estimates of the number of nests laid each year and indirectly provide estimates both of the

number of females nesting in a year (based on an average of 4.1 nests per nesting female,

Murphy and Hopkins (1984)) and of the number of adult females in the entire population (based

on an average remigration interval of 2.5 years, (Richardson et al., 1978). The total number of

nests laid along the U.S. Atlantic and Gulf coasts, 1989-1998, has ranged from 53,016-89,034

annually, representing, on average, an adult female population of 44,780 (Table 11) [(nests/4.1) *

2.5]. On average, 90.7% of the nests were from the South Florida Subpopulation, 8.5% were

from the Northern Subpopulation, and 0.8% were from the Florida Panhandle Subpopulation.

The ratio method used by TEWG (1998) to estimate minimum population size of

post-pelagic loggerhead turtles in nearshore U.S. waters required assumptions about the accurate

representation of loggerhead size classes in the stranding data. A recent, cursory appraisal of

capture data and of TED design information leads the TEWG to believe that this assumption may

not hold true. Data from the naked-net study conducted by the Gulf & South Atlantic Fisheries

Development Foundation, Inc. (1998; Jamir, 1999) show a significantly smaller (KolomogrovSmirnov Test, P<0.005) proportion of larger loggerheads (large immatures and adults) than

appears in the stranding data (Table 18). A size bias in the stranding data might be expected if

large loggerheads are less likely to be excluded by TEDs than small loggerheads and because of

this, they incur higher mortality and are more likely to strand. A recent examination of

loggerhead morphometrics and TED exit opening dimensions shows that loggerheads greater

than approximately 70 cm may not be readily excluded by TEDs (Epperly and Teas, 1999). Also,

nesting female loggerheads that occur in nearshore habitats may be more prone to incidental

capture and repeat captures, thereby increasing their mortality rate and probability of stranding.

In light of this information, we have chosen not to recalculate population size estimates using the

ratio method.

Conclusions from nest numbers, strandings, and in-water surveys

In a species with a long age to maturity, such as the loggerhead, nesting trends alone may

give an incomplete picture of population status. It is conceivable for a population with no new

recruitment to the benthic juvenile stage to continue to show increases in nesting for a number of
50 NOAA Tech. Mem. NMFS-SEFSC-444

years as benthic juveniles from past cohorts mature. Conversely, a population could continue to

show declines in nesting over time due to losses of adults while the immature population is

increasing. Thus multiple lines of evidence must be considered in order to determine true

population status.

The loggerhead recovery plan established recovery goals primarily in terms of nesting

activity. While the loggerhead recovery plan was drafted before the mitochondrial DNA

evidence of distinct nesting subpopulations became available, goals in terms of numbers of nests

per season were given for North Carolina, South Carolina and Georgia, roughly encompassing

the range of the Northern Subpopulation, with a goal for Florida of an increasing adult female

population.

Taken together, the data reviewed by the TEWG would suggest that the South Florida

Subpopulation of loggerheads is increasing, while no trends are evident in the Northern

Subpopulation in recent years. Long term trends at Hutchinson Island, an index nesting beach for

the South Florida Subpopulation, has increased by approximately 4% per year since 1981. While

stranding data and trends cannot be attributed to subpopulations, the size of the South Florida

Nesting Subpopulation taken together with the genetic evidence that animals from this

Subpopulation disperse along the entire Atlantic coast suggest that it predominates in the

strandings. Strandings have increased in recent years and there is evidence of a shift in size class

distribution toward smaller turtles, supportive of an increasing population. Again, the in-water

data are not conclusive, but apparent increases at the St. Lucie Power Plant, where in 1999 an

estimated 69% were from the South Florida Subpopulation (Witzell et al., in review), also

support the increasing population hypothesis. Apparent increases in nesting in the Florida

panhandle area are confounded by increases in monitoring, making interpretation uncertain.

Correlation analysis of the overlapping time series of available data (1989-1997) suggest

that, in general, nest counts and strandings do not correlate well (Table 19). High correlation

coefficients do occur between time series that are rapidly increasing (e.g., Panhandle nests and

northeast strandings), and these correlations disappear when the residuals are compared. The

notable exception to the lack of correlation is between Northern nests and adults stranding in the

Southeast Region (r = 0.822, P=0.011), a relationship that holds in the residuals of the lntransformed data (r = 0.822, P = 0.007). Adult strandings in this region were also well correlated

with nest counts in each of the 4 Northern Subpopulation States. Further analysis of strandings

correlations by size were not possible because a much higher proportion of stranded turtles in

other regions are reported without measurements (SE region reporting average is 52.5%, range

45-59%). Southeast strandings and Gulf of Mexico strandings were well correlated (r=0.762,

P=0.017) but failed to correlate with northeast strandings. St. Lucie captures of adults and

immatures correlated fairly well with South Florida nests and Southeast strandings, although the

correlations were not significant at the P<0.05 level (raw data from the St. Lucie captures also

correlated with Panhandle nests, but the residuals did not correlate). As more years of data are

collected, more detailed time series analyses may reveal which sources of loggerhead population

data correlate through time and are likely to be reasonable proxies for population size

distributions and trends.
NOAA Tech. Mem. NMFS-SEFSC-444 51

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Foraging Area Life History Stage(s)

N

Nesting Subpopulation

Western North Atlantic

Sou

Atla

N.C. to

Northeast

Florida

South

Florida

NW

Florida

Yucatán,

Mexico Bra

Northeast U.S.1, 13 Primarily Benthic Immature 82 0.25 0.59 0.00 0.16 0.

 Chesapeake Bay*,2 Primarily Benthic Immature 63 0.54 0.46

Pamlico and Core Sounds, N. C.3 Benthic Immature 97 0.32 0.64 <0.01 0.03 0.

S. Carolina*,4 Benthic Immature 33 0.50 0.50

Georgia*,5 Benthic Immature 97 0.59 0.41

Georgia6 Benthic Immature 192 0.24 0.73 0.03 0.

Hutchinson Island, Florida12 Benthic Immature 109 0.10 0.69 0.20

Florida Bay7 Benthic Immature and Adult 51 0.08 0.84 0.08

Mediterranean Sea*8,9 Unknown 59 0.57 

Azores and Madeira10 Pelagic Immature 183 0.19 0.71 0.00 0.11 0.

Western Mediterranean11 Pelagic Immature 59 0.02 0.45 

Eastern Mediterranean11 Pelagic Immature 52 0.02 0.47 Eastern Mediterranean11 Benthic Immature and Adult 58 

* Studies were cond ucted before full c omplement of geneti c analyses of Atlantic nesting beaches were available.** Includes rookeries of Turkey and Cyprus , in addition to Greece.1 Rankin-Baransky, 1997; 2 Norrgard, 1995; 3 Bass et al., 1998; 4 Sears et al., 1995; 5 Sears, 1994; 6 Bass et al., 1999; 7 B. Schroeder, unpublished data; 8 Laur

1995, 10Bolten et al., 1998, 11Laurent et al., 1998, 12Witzell et al., in review; 13 Rankin -Baransky et al., in review.

NOAA Tech. Mem. NMFS-SEFSC-444 57
NOAA Tech. Mem. NMFS-SEFSC-444 58Table 11. Loggerhead sea turtle nests in the U.S., 1989-1998. Minor nesting occurring north of North Caroli

Florida is not included. Data are from Meylan et al. (1995), Schroeder (1994), Maley (1995), Maley and Mu

1994), Maley and Harris (1991, 1992), Harris and Maley (1990), Florida Department of Environmental Protedata), Georgia Department of Natural Resources (unpubl. data), South Carolina Department of Natural Resoudata) and North Carolina Department of Environment, Health and Natural Resources (unpubl. data). Nesting are based on an average of 4.1 nests/female in all subpopulations. Area Year 

1989 1990 1991 1992 1993 1994 1995 1996 1997 

Northern Nesting Subpopulation: North Carolina 457 806 931 729 485 1021 662 776 5

South Carolina 2444 4491 3657 3943 2757 4136 2959 3892 21

Georgia 691 1085 1209 1054 475 1375 1028 1111 8

N.E. Florida 778 1355 952 759 809 1355 1149 1142 8

Total 4370 7737 6749 6485 4526 7887 5798 6921 44

Nesting Females 1066 1887 1646 1582 1104 1924 1414 1688 10

South Florida Nesting Subpopulation: S.E. Florida 46295 62071 63416 59677 50618 64410 71394 68174 562

S.W. Florida 2236 3085 3959 4186 3836 5395 6973 6461 67

Total 48531 65156 67375 63863 54454 69805 78367 74635 630

Nesting Females 11837 15892 16433 15576 13281 17026 19114 18204 153

Florida Panhandle Nesting Subpopulation: Total 113 174 287 351 560 772 928 891 11

Nesting Females 28 42 70 86 137 188 226 217 2

Total U.S. Nests 53014 73067 74411 70699 59540 78464 85093 82447 685

Total U.S. Nesting Females 12930 17821 18149 17244 14522 19138 20754 20109 167
NOAA Tech. Mem. NMFS-SEFSC-444 59 Table 12. Mean annual percent change in loggerhead nest numbers in the U.S. 1989-1998, based on slopes 

transformed data. Bold F statistics are significant at P < 0.05; variance in nest counts from year-to-year and nesting reduced the power to detect trends for many areas. Data are for calculations of minimum number of be accurate for assessments of trends due to yearly changes in nesting beach survey effort. See text for a disloggerhead nesting numbers for each subpopulation. North Carolina South Carolina Georgia N.E. Florida Total Northern S.E. Florida S.W. Florida Total S. Florida Florida Panhand le 

Percent change per year 1989- 1998 1.7% -0.1% 1.5% 2.9% 0.8% 2.9% 13.2% 3.6% 2.6% 

SE of slope 0.031 0.030 0.036 0.026 0.027 0.014 0.014 0.014 0.025 

r-square 0.036 0.000 0.021 0.136 0.012 0.344 0.919 0.461 0.932 

F 0.297 0.002 0.175 1.261 0.095 4.197 90.805 6.840 109.060 
NOAA Tech. Mem. NMFS-SEFSC-444 60Table 13. Correlations (Pearson) of loggerhead nests from index beaches at Cape Island, Cape Romain Nati

Refuge, S.C., and Hutchinson Island, Fla. with nest totals (excluding each index beach) from each state and rBold correlations are significant at P < 0.05. A. Raw nest numbers. B. Residuals from ln-transformed regress

nest numbers from Cape Island, S.C. and Hutchinson Island, Fla. were not significantly correlated (r = 0.466their regression residuals were highly correlated (r = 0.7602, P = 0.01). North Carolina South Carolina (excluding Cape Is.) Georgia NE Florida Northern Subpopulation (excluding Cape Is.) SE Florida (excluding Hutchinson Is.) SWFlorida S. Flo

Subpo(excluHutch

A. Raw nest numbers Cape Island, S.C. 0.8132 0.8274 0.8412 0.6966 0.9017 0.5615 -0.0251 0.462

Hutchinson Island, Fla. 0.5307 0.5107 0.5750 0.6511 0.6213 0.9547 0.7225 0.945

B. Regre ssion residuals Cape Island, S.C. 0.8898 0.8247 0.9047 0.7809 0.9470 0.8813 0.6312 0.871

Hutchinson Island, Fla. 0.6501 0.6470 0.6663 0.6281 0.7305 0.9414 0.8206 0.939
NOAA Tech. Mem. NMFS-SEFSC-444 61 Table 14. Correlations (Pearson) of U. S. loggerhead nests from all states and regions, 1989-1998. Bold co

significant at P < 0.05. A. Raw nest numbers. B. Residuals from ln-transformed regression analysis. North Carolina South Carolina Georgia N.E. Florida Northern Total S.E. Florida S.W. Florida S. Florida Total 

A. Raw nest numbers North Carolina 1.000 South Carolina 0.818 1.000 Georgia 0.935 0.733 1.000 N.E. Florida 0.694 0.683 0.649 1.000 Northern Total 0.920 0.963 0.865 0.799 1.000 S.E. Florida 0.688 0.617 0.702 0.742 0.723 1.000 S.W. Florida 0.271 0.101 0.285 0.459 0.232 0.758 1.000 S. Florida Total 0.635 0.543 0.650 0.717 0.658 0.992 0.832 1.000 Florida P anhandle 0.103 -0.091 0.096 0.354 0.042 0.570 0.960 0.665 

B. Regre ssion residuals North Carolina 1.000 South Carolina 0.855 1.000 Georgia 0.905 0.706 1.000 N.E. Florida 0.695 0.693 0.633 1.000 Northern Total 0.940 0.967 0.844 0.787 1.000 S.E. Florida 0.828 0.807 0.822 0.722 0.873 1.000 S.W. Florida 0.709 0.525 0.758 0.431 0.635 0.863 1.000 S. Florida Total 0.824 0.790 0.826 0.713 0.863 0.999 0.876 1.000 Florida P anhandle 0.103 0.012 0.029 -0.037 0.040 0.108 0.447 0.115 
62 NOAA Tech. Mem. NMFS-SEFSC-444

Table 15. Loggerhead strandings by region.

1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 Total

N.E.U.S.

Atlantic

93 175 166 134 207 158 206 221 227 299 274 346 2506

S.E.U.S.

Atlantic

828 1179 904 896 1091 606 714 558 777 859 1127 902 10441

Gulf of

Mexico

278 371 298 392 288 199 176 186 302 245 415 376 3526

Total 1199 1725 1368 1422 1586 963 1096 965 1306 1403 1816 1624 16473

Table 16. Loggerhead stranding trends. The annual rate of change for each region is based on a

regression slope of ln-transformed strandings, with the model fit (r-square) given in parentheses.

Slope values in bold are significant at P < 0.05.

 Annual percent change

1986-88 1995-97 Percent

change

1986-1997

(r-square)

 1989-1997

(r-square)

1991-1997

(r-square)

N.E.U.S.

Atlantic

434 

(10%)

919

(19%)

+ 112% +8.9%

(0.801)

+10.0%

(0.846)

+11.5

(0.911)

S.E.U.S.

Atlantic

2911

(68%)

2888

(60%)

- 1% -0.8%

(0.017)

+1.5%

(0.030)

+9.1% 

(0.659)

Gulf of Mexico 947

(22%)

1036

(21%)

+ 9% 0.0%

(0.000)

+3.1%

(0.068)

+13.9%

(0.759)

Total 4292 4843 + 13% 0.0%

(0.017)

+3.1%

(0.138)

+10.5%

(0.835)
NOAA Tech. Mem. NMFS-SEFSC-444 63

Table 17. Trends in loggerhead abundance at the St. Lucie Nuclear Power Plant, St. Lucie, Florida

(Quantum Resources, 1995, unpublished data), over various time intervals to compare with nests

and strandings. Bold values are significant at the P<0.05 level.

 Annual perce nt increase

1977-1998

 (r-square)

1989-1998

 (r-square)

1991-1998

 (r-square)

All loggerheads 4.2%

(0.401)

14.0%

(0.783)

17.0%

(0.764)

Adults ( 92 cm SCL) 12.3%

(0.605)

11.2%

(0.2282)

25.3%

(0.546)

Large Benthic Immatures (70-

91.9 cm SCL)

5.5%

(0.377)

10.1%

(0.467)

17.0%

(0.687)

Small Benthic Immatures (<70

cm SCL)

3.6%

(0.297)

17.0%

(0.890)

16.5%

(0.805)

Table 18. Observed size distributions for loggerheads in zones 29-32 of the southeast U.S., May

1997 - May 1998 (except January and February).

Data source <70 cm SCL 70-91.9cm SCL 92cm SCL

Strandings Data (n=395) 77% 15% 17%

Naked Net Captures (n=20)192% 8% 0%

1Gulf & South Atlantic Fisheries Development Foundation, Inc. (1998), Jamir (1999)
NOAA Tech. Mem. NMFS-SEFSC-444 64Table 19. Pearson correlation coefficients between loggerhead nests and strandings in the southeastern U. S. an

and loggerhead captures at the St. Lucie Power Plant , 1989-1997. Bold coefficients are significant at P < 0.05data. B. Residuals from ln-transformed regression analysis. Northern nests South Florida nests Panhand le nests NE US Strandings SE US Strandings Gulf of Mexico Strandings SE US Adult Strandings SE 

ImStr

A. Raw data Northern nests 1.000 S. Florida nests 0.604 1.000 Panhand le nests -0.146 0.542 1.000 NE US Strandings -0.128 0.530 0.933 1.000 SE US Strandings 0.263 0.294 0.166 0.311 1.000 Gulf of Mexico Strandings -0.121 -0.026 0.300 0.261 0.762 1.000 SE US Adult Strandings 0.790 0.549 0.219 0.250 0.592 0.310 1.000 SE US Immature Strandings 0.268 0.186 0.041 0.220 0.977 0.727 0.556 1.0

St. Lucie All 0.072 0.645 0.735 0.658 0.501 0.464 0.412 0.3

St. Lucie Ad ults 0.089 0.458 0.399 0.370 0.655 0.522 0.433 0.5

St. Lucie Immatures 0.065 0.665 0.785 0.701 0.458 0.442 0.395 0.2

B. Ln-transfo rmed residuals Northern nests 1.000 S. Florida nests 0.855 1.000 Panhand le nests 0.285 0.322 1.000 NE US Strandings 0.126 0.218 -0.048 1.000 SE US Strandings 0.251 0.223 -0.728 0.366 1.000 Gulf of Mexico Strandings -0.138 -0.256 -0.824 -0.165 0.766 1.000 SE US Adult Strandings 0.822 0.534 0.009 0.286 0.556 0.237 1.000 SE US Immature Strandings 0.237 0.194 -0.761 0.385 0.988 0.775 0.541 1.0

St. Lucie All 0.193 0.319 -0.182 -0.068 0.532 0.393 0.330 0.4

St. Lucie Ad ults 0.005 0.068 -0.275 0.202 0.633 0.509 0.369 0.5

St. Lucie Immatures 0.249 0.389 -0.118 -0.172 0.450 0.316 0.294 0.3
NOAA Tech. Mem. NMFS-SEFSC-444 65 Figure 15. Loggerhead nesting areas, Western North Atlantic.
66 NOAA Tech. Mem. NMFS-SEFSC-444

Figure 16. Loggerhead nests for the Northern Nesting Subpopulation, 1989-1998. A. Nest counts by

state. B. Long-term nesting records for Cape Island, S.C.
NOAA Tech. Mem. NMFS-SEFSC-444 67

Figure 17. Loggerhead nests for the South Florida Nesting Subpopulation, 1989-1998. A. Nest

counts by region. B. Long-term nesting records for Hutchinson Island, Fla.

A.

B.
NOAA Tech. Mem. NMFS-SEFSC-444 68Figure 18. Loggerhead turtles captured by the cooling water intake pipe at the St. Lucie Nuclear Power Pla

1977-1998.
NOAA Tech. Mem. NMFS-SEFSC-444 69 Figure 19. Size distributions and cumulative frequencies for loggerhead turtle strandings. A. Pooled data for 

Pooled data for 1996-1997.
NOAA Tech. Mem. NMFS-SEFSC-444 70 Figure 20. Size distributions of loggerheads captured at the St. Lucie Nuclear Power Plant, 1977-1998, repres

proportion of total captures. Time series of total captures is included (right y-axis).
NOAA Tech. Mem. NMFS-SEFSC-444 71

Part 3. Anthropogenic Mortality Sources

Summary

A number of anthropogenic mortality sources have been identified for Kemp’s ridley and

loggerhead sea turtles (National Research Council, 1990; NMFS & USFWS, 1991; USFWS and

NMFS, 1992) but few, outside drowning in bottom trawls, have been quantified with any degree of

confidence. While they still cannot be quantified, new information in recent years leads the TEWG

to be particularly concerned about longline fisheries and coastal gillnet fisheries, and about marine

debris and pollution, mortality sources that primarily affect the pelagic immature stage. A more

thorough assessment of anthropogenic mortality sources is provided in the first TEWG report

(1998).

Mortality Sources for Pelagic Stage

Important sources of mortality occur in the pelagic phase of loggerhead turtles and are not

revealed by numbers of stranded loggerheads. Likewise, this pelagic mortality is not accounted for

in estimates of population size and mortality presented in TEWG (1998). This pelagic phase is

approximately a decade in length (Bjorndal et al., 2000; Bolten et al., 1994). 

Pelagic longline fisheries

The loggerhead model (Crouse et al. 1987; Crowder et al. 1994) indicated that the pelagic

juvenile stage is likely the second most sensitive stage with respect to population growth rates, after

the benthic juvenile stage. Pelagic loggerhead juveniles are believed to circumnavigate the North

Atlantic, via the Atlantic Gyre, over a period of several years and have been reported from the

Azores, and the Mediterranean as well as the eastern Caribbean. During this transit they are

exposed sequentially to a series of longline fisheries that primarily target swordfish and tuna.

Aguilar et al. (1995) estimated the Spanish swordfish longlining fleet in the Mediterranean, one of

many fleets working in the region, alone captures more than 20,000 juvenile loggerheads (killing as

many as 10,700) annually. About half of these are from southeastern U.S. beaches (Bowen et al.,

1993; Laurent et al., 1993, 1998). Bolten et al. (1994, 1998) reported that virtually all of the pelagic

immature loggerheads taken in the Azorean longline fleet are from western North Atlantic nesting

subpopulations. In one study on the Mexican tuna longline fishery in the Gulf of Mexico, which

included the capture of a Kemp’s ridley, sea turtle mortality was 33% (Ulloa Ramirez and Gonzáles

Ania, 2000), comparable to the mortality rates reported by Aguilar et al. (1995). Estimated bycatch

of marine turtles by the U.S. Atlantic tuna and swordfish longline fisheries, based on observer data,

was significantly greater than reported in logbooks (Johnson et al., 1999; Witzell, 1999). Observer

records indicated that an estimated 4808 loggerheads and no Kemp’s ridleys were captured by the

U.S. fleet 1992-1997 of which 21 were released dead (Johnson et al., 1999). Logbooks indicated

that loggerheads readily ingested hooks (Witzell, 1999).
72 NOAA Tech. Mem. NMFS-SEFSC-444

Marine debris

An additional source of mortality that has not been adequately assessed is the ingestion of

anthropogenic debris by pelagic turtles. Preliminary indications are that approximately 15% of

pelagic post-hatchling loggerheads from Florida beaches have ingested plastics and approximately

46% have ingested tar within the first few weeks of pelagic foraging (n=168) (Witherington 1994,

in review). 

Mortality Sources for Benthic Feeding Stages

Trawls

Pursuant to Endangered Species Act Amendments of 1988, the National Research Council

Committee on Sea Turtle Conservation assessed the status of the U.S. sea turtles (National Research

Council, 1990). Incidental capture in shrimp trawls was identified as the major cause of sea turtle

mortality associated with human activities, killing more sea turtles than all other human activities

combined. An estimated 500-5000 benthic immature and adult Kemp's ridley annual mortalities

and an estimated 5,000-50,000 loggerhead mortalities were attributed to shrimp trawling (Ibid.),

prior to implementation of TED regulations in 1989. This compares to 75-750 and 500-5000

estimated mortalities due to all other known human causes, respectively. The Council identified

other fisheries, particularly the winter trawl fishery for summer flounder off North Carolina and

Virginia as a possible source of turtle losses and estimated those as 50-200 per year. Among the

findings of the National Research Council Committee on Sea Turtle Conservation (1990) were:

(1) mortality of benthic immature and adult stages must be reduced to prevent extinction and effect

recovery, under the condition that large numbers of hatchlings continue to be produced;

(2) shrimp trawling kills more benthic immature and adult sea turtles than all other human causes

combined;

(3) in Texas and South Carolina, strandings increased with the onset of shrimping activities and

decreased with the closure of Texas waters, indicating that 70-80% of the sea turtles stranding

during the shrimping seasons were killed in shrimp trawls; and

(4) shrimping can be compatible with sea turtle conservation if adequate controls are placed on

trawling activities, especially the mandatory use of TEDs at most times and places.

Federal regulations requiring TEDs in trawls used by offshore (seaward of the COLREGs

demarcation line) shrimp trawlers longer than 25 ft were published in 1987 (Federal Register, Vol.

52, No. 124, p. 24247, 28 June 1987). However, TEDs were first required in shrimp trawls

beginning in May 1989, but were used only sporadically until May 1990, and then only seasonally

(McDonald, 1990; Crouse et al., 1992; Henwood et al., 1992; Crouse, 1993; Weber et al., 1995). 

Full implementation of seasonal TED requirements became effective essentially in May 1991. 

Beginning in December 1992, year-round use of TEDs was required by most trawlers operating in
NOAA Tech. Mem. NMFS-SEFSC-444 73

Southeastern U.S. waters (57 FR p. 48861, Sept. 8, 1992) and this regulation was fully implemented

by December 1994 (57 FR p. 18446, Apr. 30, 1992; 57 FR p. 57348, Dec. 4, 1992). In the three

years before Federal Regulations (1986-1988), U.S. sea turtle strandings averaged 33% higher than

in TED years 1991-1993 (Crouse, Crowder and Heppell unpubl. as cited by Crowder et al., 1995).

The halt in the decline in numbers of nests at Rancho Nuevo can be attributed to a

combination of factors, including a decrease in shrimp fishing effort in Méxican waters when the

U.S. shrimping fleet was prohibited from shrimping there in the late 1970s, and the overall decline

and deterioration of the Mexican shrimping fleet. Threats to mating and nesting adult Kemp's

ridleys from shrimping off Rancho Nuevo were reportedly high until the U.S. and Méxican

governments negotiated a bilateral agreement in 1976. Under this agreement U.S. shrimping in

Méxican waters was phased out through 1979 (Iversen et al., 1993). After 1979, U.S. vessels

continued to shrimp off México, although illegally and at reduced levels, through the mid-1980s,

when the U.S. enforced the Lacey Act.

The reduction in shrimping effort caused by the departure of the U.S. fleet from Western

Gulf waters of México was furthered by the subsequent decline and deterioration of the Mexican

fleet. In 1995, the Mexican shrimp fleet in the Gulf was approximately 660 vessels, many of which

were not actually fishing. There has been a decline in the number of vessels operating in the

Campeche fleet, which currently represents about half of the operating Méxican vessels, but the

Tamaulipas fleet appears to be increasing (Table 20). Additionally, since 1978, waters out to 4 nm

along approximately 14 km of the beach at Rancho Nuevo have been closed to fishing during the

nesting season. However, this closure has not been strictly enforced, and until a few years ago,

beach workers reported observing 10-20 trawlers operating 2-5 mi off the beach at night. Since

1995, Mexican waters of the Gulf of Mexico are closed to shrimping in concert with the Texas

Closure, from May 15 through July 15, through binational agreement.

A study of the Mid-Atlantic winter trawl fishery for summer flounder identified the fishery

as a significant source of sea turtle mortality, estimating that 1063 turtles, mostly Kemp’s ridleys

and loggerheads, were caught in a 4 month period and that 89-191 were killed (Epperly et al., 1995,

1996). Consequently, since 1992, TEDs have been required in the fishery south of Cape Charles,

Virginia. At first these regulations were implemented through emergency and interim rules and

were seasonal requirements (57 FR p. 53603, Nov. 12, 1992; 57 FR p. 60135, Dec. 18, 1992; 58 FR

p. 4088, Jan. 13, 1993; 58 FR p. 8554, Feb. 16, 1993). They became year round requirements (58

FR p. 48797, Sep. 20, 1993; 59 FR p. 19584, Mar. 7, 1994), and subsequently were codified in a

final rule (61 FR p. 1846, Jan. 24, 1996) requiring TEDs in all summer flounder trawls operating

south of Cape Charles, Va. There is a temporary reprieve of the regulation automatically occurring

each year north of Oregon Inlet from Jan. 14-Mar. 15, when waters predictably are very cold.

Gill nets

Turtles are particularly susceptible to entanglement and drowning in gill net gear, especially

when nets are set and left unattended. There are a number of gill net fisheries currently operating

along the mid- and southeast U.S. Atlantic coastline that are known to incidentally capture
74 NOAA Tech. Mem. NMFS-SEFSC-444

loggerhead turtles. Although all or most nearshore gill netting in state waters of South Carolina,

Georgia, Florida, Louisiana, and Texas is prohibited by state regulations, gill netting in other states’

waters and in federal waters does occur. Of particular concern are the nearshore and inshore gill net

fisheries of the mid-Atlantic operating in New Jersey, Delaware, Maryland, Virginia, and North

Carolina state waters and/or federal waters offshore thereof. Annual peaks in loggerhead strandings

on ocean beaches and just inside the mouths of bays and inlets regularly occur in early summer and

late fall, coinciding with increased gill netting activity and observers have documented lethal takes

of both loggerhead and Kemp’s ridley turtles in these fisheries (D. Christensen, personal

communication). In December 1999 the large mesh gill net fishery for southern founder in Pamlico

Sound, N.C. was closed to fishing for 30 days, based on elevated levels of stranded Kemp’s and

loggerhead turtles in inshore waters and the documentation of takes by the fishery (64 FR, p. 70196,

Dec. 16, 1999). On May 12, 2000 the large mesh gill net fishery along eastern North Carolina and

Virginia, north of Cape Hatteras, including the ocean, was closed for 30 days because of elevated

strandings off northern North Carolina, mostly loggerhead turtles, observers’ documentation that the

large mesh gill net fisheries for monkfish and dogfish do take turtles (D. Christensen, personal

communication), and the fact that 4 of the stranded turtles were entangled in large mesh gill nets

consistent with these fisheries (65 FR p. 31500, May 18, 2000). NMFS made a determination that

the level of mortality was “severely impacting the northern nesting subpopulation of loggerheads”,

the first time management has identified a subpopulation of sea turtles as a management unit. 

Additionally, the gill net fishery targeting sharks in federal waters offshore Georgia and Florida

warrants close observation as these fisheries are known to take loggerheads (Trent et al., 1997) and

likely could take Kemp’s ridleys. Observers are required in this fishery during certain times of the

year; however, it is unclear whether periods of expected highest turtle abundance are adequately

covered.

NMFS has initiated discussions with coastal states of the mid-Atlantic in an attempt to

address these take issues, however, little progress has been made (T. Conant, personal

communication). There is concern by vessel captains over liability issues, and many have refused to

take observers on board (S. Epperly, personal communication) despite the fact that many are

classified as Category 2 fisheries under the Marine Mammal Act and are required to take observers

if asked.

Hook and line

Loggerhead and Kemp’s ridleys are known to bite a baited hook, frequently ingesting the

hook. Hooked turtles have been reported by the public fishing from boats, piers, the beach, banks,

and jetties (Cannon et al., 1994; J. Braun McNeill, personal communication; A. Cannon personal

communication; S. Epperly, personal communication) and from commercial fishermen fishing for

reef fish and for sharks with both single rigs and bottom longlines (S. Epperly, personal

communicaton). Necropsies of turtles have revealed hooks internally (Witzell and Teas, 1994),

which often were the cause of death (W. Teas, personal communication). An investigation of

injuries and mortalities related to fish hook ingestion is underway at the National MarineFisheries

Laboratory in Galveston, Texas (T. Fontaine, personal communication) and NMFS currently is

exploring adding questions about encounters with sea turtles to intercept interviews of recreational
NOAA Tech. Mem. NMFS-SEFSC-444 75

fishermen conducted by the Texas Parks and Wildlife Department and under the auspicies of the

Marine Recreational Fishery Statistics Surveys conducted throughout the Gulf of Mexico and along

the Atlantic Coast (T. Fontaine, personal communication; T. Conant, personal communication). 

NMFS also is considering questioning recreational fishermen aboard headboats throughout the

southeast U.S. Atlantic and the Gulf of Mexico to quantify their encounters with sea turtles, also (J.

Braun McNeill, personal communication).

Power plants

Power plants are known to entrain loggerhead and Kemp’s ridley sea turtles at the intake

canals to their cooling systems. In some cases, the turtles may become trapped on intake screens

and drown if they are unable to surface. The cumulative effect of such mortalities is unclear, but is

believed currently to be relatively small. Some plants, such as the Brunswick Steam and Electric

Plant in Southport, N.C., placed permanent diversion screens over the mouths of their intake canals

to keep turtles away from the intake screens of the plant but this has not been 100% effective (W.

Pollard, personal communication). Sea turtle entrainment has also been reported at four other

power plants in eastern Florida (National Research Council, 1990) and recently at the Crystal River

plant in western Florida. This latter facility is of particular concern given its proximity to

aggregations of subadult Kemp’s ridley turtles in the eastern Gulf of Mexico (Schmid and

Barichivich, submitted). Sea turtle monitoring at the Crystal River Energy Complex increased

substantially in 1998 and 40 live takes (37 were Kemp’s ridleys) and 5 lethal takes (all Kemp’s),

were observed that year (NMFS, 1999), mainly from February-May. The turtles were subadults,

with carapace lengths of 21-55 cm. Through March 1999, 4 subadult Kemp’s ridleys were removed

from the intake bars and released alive (D. Bruzek, personal communication).

In some cases the turtles pass unharmed through intake pipes into holding ponds, where they

are then captured and released. The best documented situation of this type of interaction is the St.

Lucie Nuclear Power Plant (Florida Power and Light Co.) where, in recent years, capture rates have

exceeded 200 loggerheads annually and the annual capture of Kemp’s ridleys is generally less than

6 (Quantum Resources, 1995; J. Gorham, personal communication). 

Marine debris

Of 1710 turtles necropsied between 1980 and 1992, 197 (11.5%) had ingested debris,

including plastic pieces and balloons (Witzell and Teas, 1994): a greater proportion of loggerheads

were affected than were Kemp’s ridleys, and in both species the percentage impacted by digested

debris was highest in the Gulf of Mexico. Studies conducted in the western Gulf of Mexico

reported debris ingestion rates of 51% for loggerheads and 34% for Kemp’s ridleys (summarized in

Witzell and Teas, 1994). Apparently Kemp’s ridleys in the northeast U.S. consume little or no

plastic debris (Sadove and Morreale, 1990; Burke et al., 1994).
76 NOAA Tech. Mem. NMFS-SEFSC-444

Literature Cited

Aguilar, R., J. Mas and X. Pastor. 1995. Impact of Spanish swordfish longline fisheries on the

loggerhead sea turtle Caretta caretta populations in the Western Mediterranean. U.S. Dep.

Commer. NOAA Tech. Mem. NMFS-SEFSC-361:1-6.

Bjorndal, K.A., A.B. Bolten, and H.R. Martins. 2000. Somatic growth model of juvenile

loggerhead sea turtles: duration of the pelagic stage. Mar. Ecol. Prog. Ser. 202:265-272.

Bolten, A.B., K.A. Bjorndal, and H.R. Martins. 1994. Life history model for the loggerhead sea

turtle (Caretta caretta) populations in the Atlantic: Potential impacts of a longline fishery. 

U.S. Dep. Commer. NOAA Tech. Mem. NMFS-SWFC-201:48-55.

Bolten, A. B., K. A. Bjorndal, H. R. Martins, T. Dellinger, M. J. Boscoito, S. E. Encalada, and B.

W. Bowen. 1998. Transatlantic developmental migrations of loggerhead sea turtles

demonstrated by mtDNA sequence analysis. Ecol. Applic. 8:1-7.

Bowen, B.W., J.C. Avise, J.I. Richardson, A.B. Meylan, D. Margaritoulis, and S.R.

Hopkins-Murphy. 1993. Population structure of loggerhead turtles (Caretta caretta) in the

northwestern Atlantic Ocean and Mediterranean. Sea. Conserv. Biol. 7:834-844.

Burke, V.J., S.J. Morreale, and E.A. Standora. 1994. Diet of the Kemp’s ridley sea turtle,

Lepidochelys kempii, in New York waters. Fish. Bull. 92:26-32.

Cannon, A.C., C.T. Fontaine, T.D. Williams, D.B. Revera, and C.W. Caillouet, Jr. 1994. 

Incidental catch of Kemp’s ridley sea turtles (Lepidochelys kempi), by hook and line, along

the Texas Coast, 1980-1992. U.S. Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC341:40-42.

Crouse, D. 1993. Victory! TEDs required in all U.S. shrimp trawls. Marine Turtle Newsletter

61:3-5.

Crouse, D.T., L.B. Crowder, and H. Caswell. 1987. A stage-based population model for

loggerhead sea turtles and implications for conservation. Ecol. 68:1412-1423.

Crouse, D. T., M. Donnelly, M. J. Bean, A. Clark, W. R. Irvin and C. E. Williams. 1992. The

TED experience: claims and reality. Center for Marine Conservation, Environmental

Defense Fund and National Wildlife Federation, Washington, D.C., 17 pp.

Crowder, L.B., D.T. Crouse, S.S. Heppell, and T.H. Martin. 1994. Predicting the impact of turtle

excluder devices on loggerhead sea turtle populations. Ecol. Applic. 4:437-445.
NOAA Tech. Mem. NMFS-SEFSC-444 77

Crowder, L. B, S. R. Hopkins-Murphy, and J. A. Royle. 1995. The effects of turtle excluder

devices (TEDs) on loggerhead sea turtle strandings with implications for conservation. 

Copeia 1995:773-779.

Epperly, S.P., J. Braun, A. J. Chester, F. A Cross, J.V. Merriner, and P.A. Tester. 1995. Winter

distribution of sea turtles in the vicinity of Cape Hatteras and their interactions with the

summer flounder trawl fishery. Bull. Mar. Sci. 56:547-568.

Epperly, S.P., J. Braun, A. J. Chester, F. A Cross, J.V. Merriner, P.A. Tester, and J.H. Churchill. 

1996. Beach strandings as an indicator of at-sea mortality of sea turtles. Bull. Mar. Sci.

59:289-297.

Henwood, T., W. Stuntz and N. Thompson. 1992. Evaluation of U.S. turtle protective measures

under existing TED regulations, including estimates of shrimp trawler related turtle

mortality in the wider Caribbean. U.S. Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC303, 15 pp.

Iversen, E. S., D. M. Allen and J. B. Higman. 1993. Shrimp capture and culture fisheries of the

United States. Halsted Press, New York, N.Y., 247 pp.

Johnson, D.R., C. Yeung, and C.A. Brown. 1999. Estimates of marine mammal and marine turtle

bycatch by the U.S. Atlantic pelagic longline fleet in 1992-1997. U.S. Dep. Commer.

NOAA Tech. Mem. NMFS-SEFSC-418, 70 pp.

Laurent, L., J. Lescure, L. Excoffier, B. Bowen, M. Domingo, M. Hadjichristophorou, L. Kornaraki,

and G. Trabuchet. 1993. Étude génétique des relations entre les populations

méditerranéenne et atlantique d’une tortue marine (Caretta caretta) à l’aide d’un marqueur

mitochondrial. C.R. Acad. Sci. Paris 316:1233-1239.

Laurent, L, P. Casale, M.N. Bradai, B.J. Godley, G. Gerosa, A.C. Broderick, W. Schroth, B.

Schierwater, A.M. Levy, D. Freggii, E.M. Abd El-Mawla, D.A. Hadoud, H.E. Gomati, M.

Domingo, M. Hadjichristophorou, L. Kornaraky, F. Demirayak, and Ch. Gautier. 1998. 

Molecular resolution of marine turtle stock composition in fishery bycatch: a case study in

the Mediterranean. Molecular Ecol. 7:1529-1542.

McDonald, D. 1990. Shrimpers arrested for non-compliance as efforts to enforce TED regulations

intensify. Marine Turtle Newsletter 51:10-12.

National Marine Fisheries Service and U.S. Fish and Wildlife Service. 1991. Recovery plan for

U.S. population of loggerhead turtle Caretta caretta. National Marine Fisheries Service,

Washington, D.C., 64 pp.

National Research Council. 1990. Decline of the sea turtles: causes and prevention. National

Academy Press, Washington, D.C., 259 pp.
78 NOAA Tech. Mem. NMFS-SEFSC-444

Quantum Resources. 1995. Annual environmental operating report for 1994. Unpublished report. 

Florida Power and Light Company, Juno Beach, Fla., 49pp.

Sadove, S.S. and S.J. Morreale. 1990. Marine mammal and sea turtle encounters with marine

debris in the New York Bight and the northeast Atlantic. U.S. Dep. Commer. NOAA Tech.

Mem. NMFS-SWFSC-154:562-570.

Schmid, J.R. and W.J. Barichivich. submitted. Kemp’s ridley turtle, Lepidochelys kempii. In:

Turtles of Florida. Chelonian Conserv. Biol. Monogr.

Trent, L., D.E. Parshley, and J.K. Carlson. 1997. Catch and bycatch in the shark drift gillnet

fishery off Georgia and East Florida. Mar. Fish. Rev. 59:19-28.

Turtle Expert Working Group. 1998. An assessment of the Kemp’s ridley (Lepidochelys kempii)

and loggerhead (Caretta caretta) sea turtle populations in the Western North Atlantic. U.S.

Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC-409, 96 pp.

Ulloa Ramirez, P.A. and L.V. González Ania. 2000. Incidence of marine turtles in the Mexican

longline tuna fishery in the Gulf of Mexico. U.S. Dep. Commer. NOAA Tech. Mem.

NMFS-SEFSC-436:110.

U.S. Fish and Wildlife Service and National Marine Fisheries Service. 1992. Recovery plan for the

Kemp's ridley sea turtle (Lepidochelys kempii). National Marine Fisheries Service, St.

Petersburg, Fla., 40 pp.

Weber, M., D. Crouse, R. Irvin, and S. Indicello. 1995. Delay and denial: a political history of sea

turtles and shrimp fishing. Center for Marine Conservation, Washington, D.C., 46 pp.

Witherington, B.E. 1994. Flotsam, jetsam, post-hatchling loggerheads, and the advecting surface

smorgasbord. U.S. Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC-351:166-168.

Witherington, B.E. in review. Ecology of neonate sea turtles inhabiting debris lines near the Gulf

Stream front. Mar. Biol.

Witzell, W.N. 1999. Distribution and relative abundance of sea turtles caught incidentally by U. S.

longline fleet in the western North Atlantic Ocean, 1992-1995. Fish. Bull. 97:200-211.

Witzell, W.N. and W.G. Teas. 1994. The impacts of anthropogenic debris on marine turtles in the

Western North Atlantic Ocean. U.S. Dep. Commer. NOAA Tech. Mem. 355:1-21.
NOAA Tech. Mem. NMFS-SEFSC-444 79

Table 20. Total shrimping vessels and vessels operating in the Campeche and Tamaulipas fleets, 1982-1995.

 Campeche Fleet Tamaulipas Fleet

Year Total Operating Total Operating

1982 605 578

1983 575 507

1983 544 481

1985 512 357

1986 488 354

1987 469 368 unknown 124

1988 450 324 unknown 116

1989 423 294 unknown 139

1990 412 259 unknown 134

1991 356 225 unknown 145

1992 398 293 unknown 169

1993 403 325 unknown 163

1994 347 268 unknown 154

1995 352 292 unknown 199
80 NOAA Tech. Mem. NMFS-SEFSC-444

Part 4. Take Limits

Introduction

One of the initial charges to the TEWG was to determine maximum take levels for each

species. This is the maximum number of turtles that can be removed by human activities (i.e., above

losses due to natural mortality) without jeopardizing recovery goals specified in each Species

Recovery Plan. While allocation of take is a regulatory action clearly beyond the scope of the

Group, a scientific assessment of the population-level effects of current and maximum allowable

anthropogenic mortality to achieve recovery goals is needed. In general, it was agreed that in order

to meet recovery goals, “take” should not increase at a rate equal to or greater than the population

growth rate for either species. However, reliable indicators of take levels and population growth

were generally unavailable. The Group determined early on that observer coverage on shrimp

vessels was an inadequate measure of fishing mortality; less than 1% of shrimp vessel trips include

observers and most observers are on offshore vessels that catch few turtles, and there are few

observer programs designed to document turtle take by other fisheries. Thus, strandings became the

Group’s proxy for take. Because the following key information was contentious or lacking, no

method for setting strandings limits was completely satisfactory to all Group members:

Population size. We do not have a reliable estimate of the number of turtles that are

susceptible to fishing mortality in a given year. While there is a fairly good estimate of nest

numbers, many assumptions must be made to translate nests into adult females, adult males,

and juveniles of both sexes. For loggerheads, an approach that utilized the ratios of

males:females and adults:juveniles in the strandings records gave estimatef of 183,627 -

316,113 turtles greater than 40 cm SCL (Table 3; TEWG, 1998). Because we have a time

series of cohort sizes for Kemp’s ridleys through records of hatchlings released at Rancho

Nuevo, a population model was developed to determine total population size annually (see

Part I., above). However, the estimate is completely dependent on unknown mortality rates,

and varies by 10-fold. Both methods of estimating population size suffer from a lack of

corroborating empirical estimates – we do not even know if we are within the correct order

of magnitude. 

 

Proportion of total mortality represented by strandings. In order to estimate population size

with the ratio method and to set take limits based on strandings, assumptions must be made

about the relationship of strandings and the turtle population. There are two issues: whether

the size distribution of stranded turtles is representative of the population as a whole, and

what proportion of total mortality is represented by stranded turtles. The probability that a

dead turtle will wash up on a beach is undoubtedly highly variable, depending on distance

from shore, water temperature, wind and currents, and possibly body size. Fishing mortality

may be size-selective and is also variable. The group agreed that a majority of turtles that

are killed in fishing gear fail to strand, but an actual proportion could not be agreed upon

because existing data are not adequate to evaluate this.
NOAA Tech. Mem. NMFS-SEFSC-444 81

Population distribution. Turtles of different sizes are not evenly distributed in time and

space. Migrating turtles may be exposed to different levels of risk from fishing mortality as

their overlap with fishing fleets varies through the year. Juvenile as well as adult

loggerheads from genetically distinct subpopulations appear to overlap disproportionately in

space, but the genetics of turtles caught by various fishing fleets has not been determined. 

This has profound implications for the ratio method of population size estimation and for

setting take limits based on strandings.

Several other unknowns also impeded development of take limits, including age at maturity (to

determine time lags). Any one of these data deficiencies could have been surmounted with careful

analysis of the best information available; the lack of consensus about take limits primarily rose

from the unknown consequences of “stacking” so many uncertainties. Setting a take limit or

estimating the impact of take is necessarily a quantitative exercise that requires a substantial amount

of information and a good understanding of how data uncertainty will affect the results (Mangel,

1993; Taylor et al., 2000). Although three methods to set strandings limits were used at various

times by the TEWG, each was presented as an Interim Stranding Limit (ISL) to be used for a single

year, rather than an agreed-upon method to be incorporated into management.

Three Interim Strandings Limits (ISL) calculations: Methods and Limitations

1) Running average

In 1996 and 1997, Kemp's ridley ISLs were calculated according to a 5 year running average

of strandings + 1 standard deviation, which was then multiplied by one-half the observed rate of

increase in the nesting population. The observed rate of increase in nests was estimated as the slope

of the natural log of nest levels over the past 5 years, transformed to a linear (proportional) scale

(TEWG, 1998). Stranded Kemp's ridley turtles greater than 10 cm, including those from coldstunning events, were used in these analyses. In 1997, the stranding data were partitioned according

to geographic regions (western Gulf of Mexico, zones 13 - 21; eastern Gulf, zones 1-12 and partial

zones 24 and 25; southeastern U.S. Atlantic, partial zones 24 and 25, 26-35, partial zone 36;

northeastern U.S. Atlantic, partial zone 36, 37 - 44). Similar methods were used for calculating

loggerhead ISLs in 1996 and 1997 with the exception of the nesting multiplier and the inclusion of

cold-stunned turtles. During both years, stranded loggerhead turtles greater than 40 cm were

partitioned among the four regions. In 1997, the standard deviation was not added to the running

averages of loggerheads on the Atlantic coast in order to give more conservative ISL estimates.

Furthermore, log-normal means were investigated for both species in 1997, but the NMFS

Southeast Regional Office, when presented with two sets of ISLs, chose to continue with arithmetic

means. In 1998 the Group agreed that the running average method was inadequate for calculating

ISLs, primarily because this method does not account for changes in fishing mortality, growing

concern about the status of the Northern Subpopulation of loggerheads, and the large number of

loggerheads stranding and the size distribution of those animals.
82 NOAA Tech. Mem. NMFS-SEFSC-444

2) Slope method

For Kemp’s ridleys, it appeared that in many years the rate of increase of strandings

exceeded the observed rate of increase of nests and hatchlings at Rancho Nuevo. A new method

was introduced in 1998 that worked on the key assumption that to maintain and enhance population

growth that rate of increase of strandings should be lower than the population rate of increase. This

method relied on the estimated slopes of hatchling production and strandings and set the ISL for the

following year to obtain a desired slope. It was used on a trial basis to calculate an ISL for Kemp’s,

as a full analysis of potential changes in population size and strandings on the ISLs had not been

attempted. As a measure of population growth rate, the group agreed to use hatchling production 2

years prior to the strandings year. Best estimates from correlation studies suggest that turtles spend

two years at sea (see Life Stage Definitions in the Kemp’s Ridley Status Report, above). This is

corroborated by the new catch curve analysis, which shows peak strandings age of 2-3 years old in

most strandings years (see Parameterization in the modeling section of the Kemp’s Ridley Status

Report, above). The strandings years used in the analysis were 1986-1997 for all regions, and did

not include headstarted, TED-tested, or cold-stunned turtles and also excluded post-hatchlings (< 10

cm). 

Slope calculation

Hatchlings released 2 years prior and strandings were ln-transformed to calculate and

compare slopes (Table 21). Slopes and regression statistics were calculated with Excel 5.0. Over

the 12 year time period the two slopes were nearly identical (m = 0.8 for hatchlings and m = 0.78

for strandings). Strandings were much more variable than hatchlings, but the regression was still

significant at the 0.05 level (p = 0.028). The standard error of the hatchling slope was 0.011.

Setting the ISL 

Strandings in 1998 should not exceed a value that maintains or lowers the slope to a desired

level. To set the ISL conservatively, the group agreed that the desired strandings slope should be 1

standard error below the slope observed for hatchling production. Using a minimization program in

Excel (Solver), the number of allowable strandings in 1998 was calculated to give a strandings

slope of 0.069 over the years 1986 - 1998 (0.069 = observed slope of 0.08 - 0.011, the standard

error). That point estimate was 334 turtles (Figure 21). 

Caveats of the slope method

The group agreed that this was an interim method in need of further research, but was a

more appropriate and conservative estimate than the running average method. Several concerns

were raised. First, the goal of the method must be determined: do we want to reduce the rate of

strandings increase to assure that the population will continue to increase rapidly, or is the

“conservative factor” of reducing the slope by 1 standard error unnecessary? Regardless of whether

the average slope of the strandings is greater than the population slope, if the slopes are not

statistically different from each other a reduction in the slope may not be needed. One problem with
NOAA Tech. Mem. NMFS-SEFSC-444 83

determining statistical differences between the two slopes is that their variances are not equal (p =

0.35, one-tailed F test), which make standard slope comparisons (t-tests) invalid. Second, setting

the strandings limit for a single year may not be feasible if the desired slope is very different from

the observed strandings slope. This will depend on the number of years used in the analysis; 12

years may be inappropriate for Kemp’s because it includes hatchling production years when the

population was declining and spans more than the current predicted generation time (10 years). 

3) Potential Biological Removal (PBR)

Potential Biological Removal (PBR) was developed for marine mammal stocks (Barlow et

al., 1995), which share many life history traits and data restrictions with sea turtles. PBR is based

on the concept that human-caused mortality of a protected mammal stock should not exceed onehalf the potential net productivity rate of the population, adjusted by a recovery factor (F) that varies

from 0 to 1. It is a simple equation that requires a minimum population size estimate (Nmin), the

maximum rate of increase predicted (or measured) for a population (Rmax), and pre-determined risk

criteria for the recovery factor:

PBR = Nmin x (0.5 x Rmax) x F 

Wade (1998) ran simulations for a number of marine mammals to determine appropriate recovery

factors for endangered and threatened populations. The default F values are: widespread

populations with little risk of depletion, F = 1.0, threatened species F = 0.5, and endangered species

F = 0.1. It has been determined that a stock experiencing a level of human-caused mortality

exceeding PBR could become depleted (Wade 1998). Although data do not exist to conduct

comparable simulations for sea turtle stocks, there are enough life history similarities between the

two taxa (relatively low fecundity, advanced age at maturity, longevity, etc.) that the majority of the

TEWG felt the PBR concept could be applied to sea turtle stocks. Discussions with the developers

of the marine mammal PBR methodology have suggested that with modification, such as

calculating separate PBRs for each life stage, PBR is potentially applicable to sea turtles.

Modifications should account for the large difference in the reproductive value and abundance of

different life stages and the extreme delay in maturity for some species.

PBR calculations for sea turtles

The PBR estimate varies depending on each of the three required parameters. As an

example, Figure 22 shows changes in PBR with various recovery factors when Rmax is set at 0.12

(i.e., observed rate of increase for Kemp’s ridleys, 1985-1998) or 0.04 (observed rate of increase for

loggerhead nests in South Florida, 1989-1998). Note that these rates of increase are probably below

Rmax because of incidental mortality. For a population of 2500 turtles, total human-induced take

would range from 10 to 1000 individuals. The lowest PBRs occur for very slow-growing

populations, where additions to the population (recruitment) each year are only a small proportion

of the total population size. While PBR numbers can be very low if conservative recovery factors

are used, it is important to realize that PBR represents a conservative estimate of allowable take that

will not prevent population recovery or cause a population decline (Wade, 1998). Also, if PBR is
84 NOAA Tech. Mem. NMFS-SEFSC-444

used to determine ISLs, additional multipliers are required to reduce total take by the proportion of

animals killed that are likely to be counted as strandings. 

Application to adult loggerheads

Because loggerheads have such a long lag time between hatching and maturation, the slope

and running average methods were judged to be inadequate for protection of the threatened

Northern Subpopulation by a majority of the Group in 1998. PBR was used to calculate the number

of adult loggerhead strandings allowed in southern Atlantic fishing zones. Strandings for all of the

zones within an area were used to determine sex ratios of adults. Also, nesting data were combined

for the S. Florida and the Florida Panhandle subpopulations. PBR for adult loggerheads was

calculated assuming the minimum population size was the estimated size of the adult populations in

U.S. waters, the maximum rate of increase was 5%, and the recovery factor was 0.5, given the

species is listed as threatened. It was assumed that strandings represent, at best, 25% of the at-sea

nearshore mortality (Murphy and Hopkins-Murphy, 1989). Thus, the ISL was set as 25% of the

calculated PBR for each area/subpopulation (Table 22) The result was a mean of 9 adult strandings

in the Northern Subpopulation zones and 98 adult strandings in all remaining areas. In addition to

the assumptions of the ratio method which are listed in TEWG (1998), this method assumes that

adult individuals of the Northern Subpopulation are found in zones 29-44 only, or are outside U.S.

waters. ISLs for turtles <92 cm SCL (and unmeasured) were calculated by region using the running

average method.

Caveats of PBR 

PBR has yet to be evaluated for sea turtles, contains a number of simplifying assumptions,

and the default recovery factors are considered arbitrary by some researchers. A minority opinion to

the ISL recommendations for 1998 strongly criticized the use of PBR and the emphasis on

separating the loggerhead ISLs by subpopulation. None of the parameters in the PBR equation have

been fully evaluated, including the strandings:deaths ratio required to set ISLs. Finally, PBR does

not include much of the important biological information available for sea turtles, such as life

history and sensitivity of various life stages.

Summary of results 1996 – 1998

Strandings of Kemp’s ridleys during 1996-1998 for comparison to the ISLs set were 393,

458, and 412, respectively (Table 23). Strandings of Kemp’s ridleys in 1996 and 1997 were less

than the ISLs set for those years. However, strandings of Kemp’s ridleys in 1998 (412) exceeded

the ISL of 334 animals by nearly 25%.

Total strandings of loggerheads during 1996-1998 exceeded total ISLs by 12-82% each year

(Table 23). In 1996 ISLs were exceeded in all regions except the northeast U.S. and in 1997 ISLs

were exceeded in all regions. In 1998 ISLs for juvenile loggerheads were exceeded in all regions

except the Eastern Gulf of Mexico. The number of adults stranding in zones 29-44, the area where

all adults are presumed to be from the Northern Subpopulation, was extraordinary (83) exceeding
NOAA Tech. Mem. NMFS-SEFSC-444 85

ISL (9) by 922%, even exceeding the number of adults stranding in other zones where they are

presumed to be from the much larger (nearly 10x) S. Florida Subpopulation and from the Florida

Panhandle Subpopulation. 

Research recommendations for future evaluation of take limits

Take limits are a contentious issue because data on sea turtles are extremely limited and

scientific opinions vary on the importance of conservative recovery factors. Better estimates of

means and uncertainty in population size, the strandings:deaths ratio, and population distribution

will help, but the level of allowable risk must be established first. As long as sea turtle populations

appear to be recovering, it is unlikely that disagreements about risk aversion will be resolved. The

distinctness of the loggerhead subpopulations is critical, as the nests of Northern Subpopulation

beaches do not appear to be increasing. Because the goals stated in the Recovery Plan specify nest

increases for these areas, thoughtful evaluation of management priorities is warranted.

A simulation analysis is underway to examine the effects of population change and variance,

number of years analyzed, and parameter uncertainty in the ISLs calculated using the slope method,

running averages, and PBR (S. Heppell, personal communication). It may also be possible to use

the Kemp’s ridley age-based population model for projections through time to see how the

population might be affected by these take levels, but that will require an estimate of the

strandings:total mortality proportion. A future meeting of population modelers familiar with a

variety of methods for analyzing sea turtle and fishery data may also be beneficial, particularly if a

working simulation model could be provided to test various take limits. 

Literature Cited

Barlow, J, S.L. Swartz, T.C. Eagle, and P.R. Wade. 1995. U.S. Marine Mammal Stock

Assessments: Guidelines for Preparation, Background, and a Summary of the 1995

Assessments. U.S. Dep. Commer. NOAA Tech. Mem. NMFS-OPR-6, 73 pp..

Mangel, 1993. Effects of high seas driftnet fisheries on the northern right whale dolphin

Lissodelphis borealis. Ecol. Applic.3:221-229.

Murphy, T. and S.R. Hopkins-Murphy. 1989. Sea Turtle & Shrimp Fishing Interactions: a

summary and critique of relevant information. Center for Marine Conservation,

Washington, D.C., 52 pp.

Taylor, B.L., P.R. Wade, D.P. DeMaster, and J. Barlow. 2000. Incorporating uncertainty into

management models for marine mammals. Conserv. Biol. 14:1243-1252.

Turtle Expert Working Group. 1998. An assessment of the Kemp’s ridley (Lepidochelys kempii)

and loggerhead (Caretta caretta) sea turtle populations in the Western North Atlantic. U.S.

Dep. Commer. NOAA Tech. Mem. NMFS-SEFSC-409, 96 pp.
86 NOAA Tech. Mem. NMFS-SEFSC-444

Wade, P. 1998. Calculating limits to the allowable human-caused mortality of cetaceans and

pinnipeds. Mar. Mammal Sci. 14:1-37.
NOAA Tech. Mem. NMFS-SEFSC-444 87

Table 21. Hatchling production and strandings levels used to calculate slopes.

Year Hatchlings produced two

years earlier

Ln

(hatchlings)

Strandings Ln(strandings)

1986 58124 10.970 306 5.724

1987 51033 10.840 162 5.088

1988 48818 10.796 216 5.375

1989 44634 10.706 150 5.011

1990 62218 11.038 258 5.553

1991 66752 11.109 165 5.106

1992 74339 11.216 156 5.050

1993 76238 11.242 327 5.790

1994 92116 11.431 550 6.310

1995 84210 11.341 396 5.981

1996 107687 11.587 382 5.945

1997 120038 11.696 412 6.021

slope 0.080 0.078

standard error 0.011 0.030

p-value <0.001 0.027

Lower 95% C.I. 0.056 0.010

Upper 95% C.I. 0.104 0.146
88 NOAA Tech. Mem. NMFS-SEFSC-444

Table 22. PBR and ISL estimates for adult loggerheads, 1998.
NOAA Tech. Mem. NMFS-SEFSC-444 89 1.Unmeasured turtles were included. 2.In 1997, methods based on both geometric and arithmetic 5-yr averages were used. However, the NMFS Southeast Regional Office chose to 

 arithmetic means.

Table 23. Summary of Incidental Stranding Limits (ISL) and total strandings for 1996-98. ISL methods for 

1998 are from the majority report.

Year Species

Region

(zones)

Size

classes1

Cold-stunned

included Calculation method 

1996 L. kempii All > 10 cm yes (Arithmetic mean + 1 std dev)*multiplier 

C. caretta Western Gulf

(13 - 21)

> 40 cm no Arithmetic mean + 1 std dev 

Eastern Gulf

(1 - 12, part 24 & 25)

> 40 cm no Arithmetic mean + 1 std dev 

Southeast US coast

(part 24 & 25 - 35, part 36)

> 40 cm no Arithmetic mean + 1 std dev 

Northeast US coast

(part 36, 37 - 44)

> 40 cm no Arithmetic mean + 1 std dev 

19972 L. kempii Western Gulf > 10 cm yes (Arithmetic mean + 1 std dev)*multiplier 

Eastern Gulf > 10 cm yes (Arithmetic mean + 1 std dev)*multiplier 

Southeast US coast > 10 cm yes (Arithmetic mean + 1 std dev)*multiplier 

Northeast US coast > 10 cm yes (Arithmetic mean + 1 std dev)*multiplier 

C. caretta Western Gulf > 40 cm no Arithmetic mean + 1 std dev 

Eastern Gulf > 40 cm no Arithmetic mean + 1 std dev 

Southeast US coast > 40 cm no Arithmetic mean without std dev 

Northeast US coast > 40 cm no Arithmetic mean without std dev 

1998 L. kempii All > 10 cm no Slope method 

C. caretta

adults

S. Fla & Panhandle subpop.s

(1 - 28)

92 cm no PBR estimate 

Northern subpopulation

(29 - 44)

92 cm no PBR estimate 

C. caretta

juveniles

Western Gulf 40 - 91 cm

& unmeasured

no Arithmetic mean without std dev 

Eastern Gulf 40 - 91 cm

& unmeasured

no Arithmetic mean without std dev 

Southeast US coast 40 - 91 cm

& unmeasured

no Arithmetic mean - 2 std dev 

Northeast US coast 40 - 91 cm

& unmeasured

no Arithmetic mean - 2 std dev 
NOAA Tech. Mem. NMFS-SEFSC-444 90 Figure 21. Slope method for ISL, showing strandings level required to achieve desired slope of m=0.0
NOAA Tech. Mem. NMFS-SEFSC-444 91 Figure 22. Effect of the recovery factor (F) on PBR for a population of 2500.
92 NOAA Tech. Mem. NMFS-SEFSC-444

Part 5. Research Recommendations

Progress towards developing realistic population level models is dependent upon obtaining

empirically derived parameters which reflect and incorporate the life history of each species. To this

end, the information required includes empirically estimated survival rates for each life history stage

defined by age. The assignment of ages and stages requires the application of age-length keys

developed from growth models, which are used to develop the age structure for the population. The

evaluation of changes in the age structure of the population can be completed by examining shifts in

size frequency distributions over time and between areas. Shifts in abundance and the age structure

must be evaluated in the context of migration in and out of areas identified as developmental or

important foraging areas. This information can only come from research on turtles in the water and

requires the accumulation of mark-recapture information over several years in multiple areas of

known utilization. These in-water areas must be sampled to ensure that growth models, abundance

indices, and size/age frequency distributions can be obtained. The distributions for these must be

described with some measure of variability so that trends and shifts can be evaluated. Problems with

data collected from strandings to estimate age and growth and survivorship have been demonstrated

and underscored the need to conduct long term in-water studies that are comparable between years

and between areas.

In-water surveys will provide data to evaluate changes in the age structure of populations.

This trend analysis should enable us to evaluate progress towards recovery as defined in the

individual recovery plans. Currently, data on the number of nesting females or nests is used to

evaluate recovery and to ensure recovery. The timeline to the fully recruited benthic stage is shorter

than that for returning to the nesting beach, especially for loggerhead turtles. Tracking recovery

requires tracking juvenile turtles to ensure recovery rather than waiting the many years to maturity,

which may be too late to reverse any declines.

Strandings data continue to provide information about unusually large mortality events that

can be attributed to either a natural event or anthropogenic activity. These data also continue to

provide information on sex ratios, maturity, breeding condition, and size frequencies. Strandings

can be used to determine population demographics and to estimate mortality rates, but only if they

are representative of the live turtle population. This assumption must be verified, as it continues to

limit the interpretation of results that rely on these data. The relationship between the location and

occurrence of strandings and environmental factors such as tides, currents, water temperature,

resource condition and quantity must be determined to fully understand where the mortality

occurred relative to where the stranding is reported. Stranding protocols have been recently

standardized to provide comparable information between areas and time periods; accumulation of

these data over several years may provide trend information if strandings are representative of the

living population.

Research must continue to pursue ways to age turtles and to validate these ages. Body parts,

such as the humerus, ear bone, or other hard structures that may harbor growth rings should be

carefully preserved for all specimens collected by the Sea Turtle Stranding and Salvage Network,

particularly body parts from known-age turtles. Validation of ages determined by humerus or sclera
NOAA Tech. Mem. NMFS-SEFSC-444 93

growth should be attempted through the use of hard part markers. The use of internal wire coded

tags placed in small Kemp’s ridleys should be continued to provide aging information. However,

the success of this effort can only be ensured through the distribution of tag detectors throughout the

range for this species and for use with both dead and live turtles.

While estimating total mortality rates empirically is an immediate and critical requirement,

the partitioning of mortality by cause is imperative for the management of these species if recovery

goals are not being met. When mortality can be partitioned and identified by cause, then

management measures can be implemented to eliminate or reduce controllable anthropogenic

sources of mortality. Interactions with fisheries gear can be fully evaluated and quantified with the

placement of observers at a statistically valid sampling level to measure incidental take and

mortality by this component on a fishery by fishery basis. There is no substitute for the direct

observation of take and mortality as evidenced by the effort to quantify these in the commercial

shrimp trawl fishery of the southeast U.S. Repeated mortality incidents occur where high levels of

strandings are reported coincident with anecdotal observations of nearshore trawling activities, but

the lack of observer coverage on nearshore vessels prevents verification and quantification of a

causal relationship between trawling activity and subsequent strandings.

Genetic studies must be continued to fully understand and quantify the stock structure of the

loggerhead turtle within U.S. Atlantic waters. Currently, the TEWG describes sub-populations

based on nesting location as supported by recent genetic studies. However, these studies are limited

to female turtles and do not provide any measure of fidelity by males. Foraging ground studies

provide information on the mixing of subpopulations. The extent to which cross over by males

occurs likely will determine the actual stock structure for this species. However, the TEWG notes

that the natal homing of female turtles, as verified by existing genetic data, suggests that a nesting

assemblage could be lost for ecological time periods if a subpopulation declines to extinction.

The pursuit of stock assessments is contingent upon these recommendations to enhance

existing research programs and develop new ones that can determine the status and condition

relative to recovery of these species. Limitations of the existing data and time series continue to be

the critical impediment to completing these analyses. Fine-tuning of data that have not been

collected for the purpose of providing empirically based population parameters will not further the

efforts towards a stock assessment. Peer reviewed results are certainly preferable and ideal for

inclusion in such stock assessments. It is the conclusion of the TEWG that further pursuit of stock

assessment analyses is unwarranted until these data and results are available to the group. 
94 NOAA Tech. Mem. NMFS-SEFSC-444

Appendix 1. Von Bertalanffy Growth Models for Wild Kemp’s Ridley Turtles:

Analyses of the NMFS Miami Laboratory Tagging Database

Jeffrey R. Schmid and Amy Woodhead

National Marine Fisheries Service

Southeast Fisheries Science Center

75 Virginia Beach Drive

Miami, FL 33149

Prepared for

The Turtle Expert Working Group

Originally submitted:

November 17, 1997

Modified and resubmitted:

October 21, 1998

A number of von Bertalanffy growth models have been developed for the Kemp’s ridley

turtle, Lepidochelys kempi. Zug and Kalb (1989), Zug (1990), and Schmid (1995, 1998) prepared

von Bertalanffy growth models for wild, subadult Kemp's ridley turtles, but the range of carapace

lengths and sample sizes for these studies were insufficient to accurately describe growth

parameters. Caillouet et al. (1995) provided a growth model for head-started Kemp's ridley turtles

released off the Texas coast and subsequently recaptured in the Gulf of Mexico. Zug et al. (1997)

recently refined their skeletochronological growth model by increasing the sample size and the size

range of turtles. Schmid and Witzell (1997) provided von Bertalanffy equations for the combined

data of tagging studies in Florida (Schmid, 1995, 1998) and subsequent recaptures at Ranch Nuevo,

Mexico. This paper presents von Bertalanffy growth model analyses for Kemp's ridley turtles using

the data available from the National Marine Fisheries Service (NMFS) Miami Laboratory

Cooperative Marine Turtle Tagging Program.

Materials and Methods

Tagging records for wild Kemp’s ridley turtles were compiled from several sources: NMFS

Miami laboratory tagging database1,2,3,4,5,6,7,8,9,10, Cape Canaveral from 1986 to 1991 (Schmid, 1995),

and Cedar Keys from 1986 to 1995 (Schmid, 1998). Carapace measurements were converted to

metric units, as necessary, and transformed to standard straight-line carapace length (SSCL; nuchal

notch to posterior end of postcentral scute) using the regression equations of Teas (1993), Schmid

(1998), and Schmid and Witzell (1997). Records for a NMFS Galveston headstart turtle, a negative
NOAA Tech. Mem. NMFS-SEFSC-444 95

growth rate (-5.6 cm/yr), and an exceedingly large growth rate (493.1 cm/yr) were omitted from the

analyses. Means are followed by ± one standard deviation.

The von Bertalanffy growth interval equation (Fabens, 1965) was fitted to Kemp’s ridley

recapture data with a non-linear least squares regression procedure (SAS Institute, 1989) as

described by Schmid and Witzell (1997). Growth models were constructed using different intervals

of time between initial tagging and recapture, in order to minimize the effects of measurement

errors on short-term recaptures. Furthermore, separate growth models were calculated for Kemp’s

ridley turtles tagged in the Gulf of Mexico and turtles tagged in the Atlantic Ocean.

Results and Discussion

Ninety-six recaptures were recorded for Kemp’s ridley turtles tagged in cooperative NMFS

tagging projects. Of this total, fifty-eight Kemp’s ridley turtles were initially tagged in the northern

Gulf of Mexico (Fla. – 31, Ms. – 1, La. – 4, and Tx. – 22), and thirty-eight were initially tagged

along the Atlantic coast (Fla. – 18, Ga. – 1, N.C. – 3, Va. – 4, and N.Y. – 12). Nine Kemp’s ridley

recaptures were at or near Rancho Nuevo, Mexico (Table 1); four of these were initially tagged in

the northern Gulf (La. – 1 and Tx. – 3) and five on the Atlantic coast (Va. – 1, N.C. – 1, and Fla. –

3). In the Gulf, a maximum recapture interval of 3.7 years was recorded for a turtle tagged in Texas

and recovered in Mississippi. In the Atlantic, turtles tagged in Florida and North Carolina were

observed nesting at Rancho Nuevo, Mexico 8.3 and 8.5 years after last capture.

Kemp’s ridley turtle carapace lengths ranges from 22.0 to 65.9 cm for all initial captures and

from 22.0 to 66.7 cm for all recaptures. A mean growth rate of 6.7 ± 6.2 cm/yr (range = 0.0 – 29.1

cm/yr) was calculated for the entire database. Carapace lengths for turtles tagged in the Gulf ranged

from 22.2 to 65.8 cm for initial captures and 22.4 to 66.7 cm for recaptures. Carapace lengths for

turtles tagged in the Atlantic ranged from 22.0 to 65.9 cm for initial captures and 22.0 to 66.0 cm

for recaptures. The mean growth rate for Kemp’s ridley turtles tagged in the Gulf (7.5 ± 6.2 cm/yr)

was significantly larger ( 2 = 4.3, df = 1, p =0.04) than that of turtles tagged in the Atlantic (5.5 ±

6.2 cm/yr).

The fundamental assumption of the von Bertalanffy model is that growth rate steadily

decreases with increasing size and age. Straight-line regression of individual growth rates on the

mean of initial and recapture carapace lengths produced negative slopes for all records, Gulf

records, and Atlantic records (Fig. 1). This suggests that Kemp’s ridley turtle growth rates decreased

with increasing mean carapace lengths, but correlation coefficients for the regressions were low

owing to considerable variation in growth rates for 20 – 50 cm turtles.

There was very little change in the estimates of von Bertalanffy growth parameters with

increasing recapture intervals, although there were increases in the standard errors of the estimates

and the residual mean square errors (MSE) of the growth models (Table 2). The model for all tag

records had the lowest MSE for the recapture interval data treatments and was therefore considered

the most appropriate growth model (Dunham, 1978). The fitted von Bertalanffy growth equation for

all recaptures combined was:
96 NOAA Tech. Mem. NMFS-SEFSC-444

Lt = 70.7 (1 - 0.9378 e-0.2035 t) (1)

The estimated asymptotic length for this model is larger than the mean size (64 cm converted

SSCL) and slightly smaller than the maximum size (72.5 cm converted SSCL) for nesting females

observed at Rancho Nuevo, Mexico (Burchfield et al., 1988).

The fitted von Bertalanffy growth equation for Kemp’s ridley turtles tagged in the Gulf of

Mexico was:

Lt = 71.1 (1 - 0.9381 e-0.2095 t), MSE = 6.6244, (2)

and the equation for turtles tagged along the Atlantic coast was:

Lt = 73.2 (1 - 0.9399 e-0.1665 t), MSE = 4.9517. (3)

The estimated asymptotic length for turtles captured in the Atlantic is slightly larger than that of the

Gulf. This observation was probably the result of the different growth rates between the two areas

and the inverse relationship of the von Bertalanffy parameters in the absence of larger individuals

(Knight, 1968; Schmid and Witzell, 1997).

Estimating age to maturity has been the primary application of the von Bertalanffy growth

models derived for Kemp’s ridley turtles. The growth curve for all recaptures combined (equation 1)

is illustrated in Figure 2. By using the mean size of nesting females (64 cm) as the size at maturity,

estimates of age to maturity ranged from 11-12 years for all tagging records, 12-13 years for

Atlantic records, and 10-11 years for Gulf records. Minimum size of nesting females (56.0 cm

converted SSCL; Burchfield et al., 1988) has also been used as an estimate of size at maturity and

resulted in age to maturity estimates as early as 7-8 years for all tagging records, 8-9 years for

Atlantic records, and 7-8 years for Gulf records. Another application of von Bertalanffy models has

been to estimate the number of years required for turtles to grow within a specified size range

(Bjorndal et al., 1995; Schmid and Witzell, 1997). Accordingly, the duration of Ogren’s (1989)

Kemp’s ridley coastal-benthic subadult stage (20 – 60 cm) would be 7 – 8 years for all tagging

records and the Gulf records, and 8 – 9 years for the Atlantic records.

Literature Cited

Bjorndal, K. A., Bolten, A. B., Coan, A. L., and Kleiber, P. 1995. Estimation of green turtle

(Chelonia mydas) growth rates from length-frequency analysis. Copeia 1995:71-77.

Burchfield P. M., Byles, R., Mongrell, J. V., Bartlett, M., and Rostal, D. 1988. Report on Republic

of Mexico/United States of America conservation effort on behalf of Kemp's ridley sea turtle

at Playa de Ranch Nuevo, Tamaulipas, Mexico, 1988. Gladys Porter Zoo, Brownsville, Tx.,

83pp. 
NOAA Tech. Mem. NMFS-SEFSC-444 97

Caillouet, C. W., Jr., Fontaine, C. T., Manzella-Tirpak, S. A., and Williams, T. D. 1995. Growth of

head-started Kemp's ridley sea turtles (Lepidochelys kempii) following release. Chelonian

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Proceedings of the First International Symposium on Kemp's Ridley Sea Turtle Biology,

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98 NOAA Tech. Mem. NMFS-SEFSC-444

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Text Footnotes

1 Bartol, S. Virginia Institute of Marine Science, P.O. Box 1346, Gloucester Point, VA 23062-1346.

2 Branstetter, S. Gulf and South Atlantic Fisheries Development Foundation, Inc., Lincoln Center,

Suite 997, 5401 West Kennedy Boulevard, Tampa, FL 33609.

3 Corcoran, G. J.L. Scott Marine Education Center & Aquarium, P.O. Box 7000, Ocean Springs,

MS 39566-7000.

4 Epperly, S. Southeast Fisheries Science Center, Beaufort Laboratory, 101 Pivers Island Road,

Beaufort, NC 28516-9722.

5 Landry, Jr., A.M. Texas A & M University at Galveston, Department of Marine Biology,

Galveston, TX 77551-5923.

6 Martin, E. Ecological Associates, Inc., P.O. Box 405, Jensen Beach, FL 34958-0405.

7 Nelson, D. A. U.S. Corps of Engineers, Waterways Experiment Station, 3909 Halls Ferry Road,

Vicksburg, MS 39180.

8 Rudloe, A. Gulf Specimen Marine Laboratories, Inc., P.O. Box 237, Panacea, FL 32346.

9 Rudloe, J. Panacea Institute of Marine Science, P.O. Drawer AB, Panacea, FL 32346.

10 Sadove, S. Puffin Consulting, Inc. P.O. Box 361, Jamesport, NY 11947; Okeanos Ocean Research

Foundation, Inc., 431 E. Main Street, Riverhead, NY 11901-2556.
NOAA Tech. Mem. NMFS-SEFSC-444 99

Table 1. Recoveries of Kemp’s ridley turtles initially tagged in the U.S. and subsequently

recaptured at or near Rancho Nuevo, Mexico.

Tagging Location

Date

Tagged

Initial

SCL

Recapture

SCL

Years at

Large

Atlantic

Virginia16/19/89 46.6 64.9 6.9

Florida2

7/24/89 54.8 64.9 4.8

Florida3

1/11/89 42.8 63.8 8.3

Florida39/30/90 36.2 62.1 5.6

North Carolina4, * 10/26/89 40.5 63.0 8.5

Gulf of Mexico

Texas511/15/92 33.9 35.070.1

Texas55/1/94 56.0 65.8 3.0

Louisiana5

8/13/94 65.8 66.2 0.7

Texas6,* 3/5/98 62.2 63.0 0.2

* Not included in von Bertalanffy growth models.

1 Bartol, S. Virginia Institute of Marine Science, P.O. Box 1346, Gloucester Point, VA 23062-1346.

2 Martin, E. Ecological Associates, Inc., P.O. Box 405, Jensen Beach, FL 34958-0405.

3 Schmid and Witzell (1997)

4 Epperly, S. Southeast Fisheries Science Center, Beaufort Laboratory, 101 Pivers Island Road,

Beaufort, NC 28516-9722.

5 Landry, Jr., A.M. Texas A & M University at Galveston, Department of Marine Biology,

Galveston, TX 77551-5923.

6 Branstetter, S. Gulf and South Atlantic Fisheries Development Foundation, Inc., Lincoln Center,

Suite 997, 5401 West Kennedy Boulevard, Tampa, FL 33609.

7 Recaptured in Laguna Madre de Tamaulipas, Mexico.
100 NOAA Tech. Mem. NMFS-SEFSC-444

Table 2. Estimated values of asymptotic length (a), intrinsic growth rate (k), and means square error

(MSE) from non-linear regression of von Bertalanffy growth interval equation for Kemp’s

ridley turtles (one asymptotic standard error in parentheses).

Data Treatment n a k MSE

All tag records 96 70.7 cm

(2.7)

0.2035

(0.0251)

5.9436

All tag records > 90 days 62 70.9 cm

(3.3)

0.2005

(0.0305)

8.7846

All tag records > 180 days 50 71.2 cm

(3.6)

0.1953

(0.0318)

9.4246

All tag records > 365 days 20 69.7 cm

(5.1)

0.2237

(0.0572)

19.5439

NMFS Miami tag records1

58 70.7 cm

(3.4)

0.2135

(0.0325)

7.1282

NMFS Miami tag records > 90 days 38 71.0 cm

(4.2)

0.2098

(0.0390)

10.3059

NMFS Miami tag records > 180 days 32 70.9 cm

(4.5)

0.2103

(0.0424)

11.8087

NMFS Miami tag records > 365 days 13 69.7 cm

(6.5)

0.2343

(0.0794)

26.3726

1

 Excuding the tag records used in Schmid and Witzell (1997).
NOAA Tech. Mem. NMFS-SEFSC-444 101

Figure 1. Kemp’s ridley turtle growth rates plotted by the mean of the initial and recapture

carapace lengths for a) all records, b) Gulf of Mexico records, and c) Atlantic Ocean

records.
102 NOAA Tech. Mem. NMFS-SEFSC-444 Figure 2.Growth curve for the combined database of all Kemp's ridley turtle recaptures. Dashed portion of the curve represents extrapolation outside the size range of the database. Numbers on the curve indicate the estimated age at 5 cm intervals.
NOAA Tech. Mem. NMFS-SEFSC-444 103

Appendix 2. Sea Turtle Strandings

Figure A2.1. Statistical zones along the U.S. Atlantic and Gulf of Mexico coasts.

 
104 NOAA Tech. Mem. NMFS-SEFSC-444

Figure A2.2a. Kemp’s ridley strandings in the U.S. Gulf of Mexico
NOAA Tech. Mem. NMFS-SEFSC-444 105
106 NOAA Tech. Mem. NMFS-SEFSC-444

Figure A2.2b. Kemp’s ridley strandings in the Southeast U.S Atlantic.
NOAA Tech. Mem. NMFS-SEFSC-444 107
108 NOAA Tech. Mem. NMFS-SEFSC-444

Figure A2.2c. Kemp’s ridley strandings in the Northeast U.S Atlantic.
NOAA Tech. Mem. NMFS-SEFSC-444 109
110 NOAA Tech. Mem. NMFS-SEFSC-444

Figure A2.3a. Loggerhead strandings in the U.S Gulf of Mexico.
NOAA Tech. Mem. NMFS-SEFSC-444 111
112 NOAA Tech. Mem. NMFS-SEFSC-444

Figure A2.3b. Loggerhead strandings in the Southeast U.S Atlantic.
NOAA Tech. Mem. NMFS-SEFSC-444 113
114 NOAA Tech. Mem. NMFS-SEFSC-444

Figure A2.3c. Loggerhead strandings in the Northeast U.S Atlantic.
NOAA Tech. Mem. NMFS-SEFSC-444 115