f 

AN ASSESSMENT OF THE KEMP'S RIDLEY 

(Lepidochelys kempii) AND LOGGERHEAD (Caretta 

caretta) SEA TURTLE POPULATIONS IN THE 

WESTERN NORTH ATLANTIC 

A Report of the Turtle Expert Working Group 

U. S. Department of Commerce 

National Oceanic and Atmospheric Administration 

National Marine Fisheries Service 

Southeast Fisheries Science Center 

75 Virginia Beach Drive 

Miami, FL 33 149 

March 1998 
f stable population dynamics processes ove 

k size and parameter estimates. 
AN ASSESSMENT OF THE KEMP'S RIDLEY 

(Lepidochelys kempii) AND LOGGERHEAD (Caretta 

caretta) SEA TURTLE POPULATIONS IN THE 

WESTERN NORTH ATLANTIC 

A Report of the Turtle Expert Working Group 

U. S . Department of Commerce 

National Oceanic and Atmospheric Administration 

National Marine Fisheries Service 

Southeast Fisheries Science Center 

75 Virginia Beach Drive 

Miami, FL 33 149 

March 1998 
NOAA TECHNICAL MEMORANDUM 

NMFS-SEFSC-409 

AN ASSESSMENT OF THE KEMP'S RIDLEY 

(Lepidochelys kempii) AND LOGGERHEAD (Caretta 

caretta) SEA TURTLE POPULATIONS IN THE 

WESTERN NORTH ATLANTIC 

A Report of the Turtle Expert Working Group 

U.S. DEPARTMENT OF COMMERCE 

William M. Daley, Secretary 

NATIONAL OCEANIC AND ATMOSPHERlC ADMINISTRATION 

D. James Baker, Administrator 

NATIONAL MARINE FISHERIES SERVICE 

Rolland A. Schrnitten, Assistant Administrator fbr Fisheries 

March 1998 

The Technical Memorandum Series is used for documentation and timely communication of 

preliminary results, interim reports, or special-purpose information. Although the Memoranda are 

not subject to complete formal review, editorial control, or detailed editing, they are expected to 

reflect sound professional work. 
I 

NOTICE 

L 

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 WS, 

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 mentioned herein, or which has as its purpose or intent to cause directly or indirectly the 

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

This publication should be cited as follows: 

Turtle Expert Working Group. 1998. An Assessment of the Kemp'z; Ridley (Lepidochelys kempii) 

and Loggerhead (Caretta caretta) Sea Turtle Populations in the Western North Atlantic. NOAA 

Technical Memorandum NMFS-SEFSC-409. 96 pp. 

Copies may be obtained in writing: 

National Marine Fisheries Service 

Miami Laboratory 

Sea Turtle Program 

75 Virginia Beach Drive 

Miami, FL 33 149 

National Technical Information Service 

5285 Port Royal Road 

Springfield, VA 22 1 6 1 

(703) 487-4650 

(800) 336-4700 (rush orders) 
Preface 

At the behest of NMFS, the SEFSC established the Turtle Expert Working Group (TEWG) 

in February 1995. The charge to the TEWG was first identifled in the Endangered Species ActSection 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 formation on the status of sea turtle species. The team should attempt to 

iden* 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 occumng in statistical zone that indicate incidental takes are occurring at levels beyond 

those authorized." Further clarification was received in a letter of appointment from Mr. Rolland 

Schrnitten, Assistant Administrator, NMFSrNOAA. 

Since the first meeting in June, 1995, there have been 3 subsequent formal meetings which 

have culminated in the two stock level reports that constitute this publication. 'There is embodied in 

this publication a status report for the loggerhead turtle, Caretta saretta, within the southeast U.S. 

and the Kemp's ridley, Lepidochelv~ kempii, which is found in both Mexican and U.S. waters. This 

effort which represents the beginning of an evaluation of the status and condition of these stocks is 

considered preliminary and exploratory. We consider these efforts a work in progress. The TEWG 

is continuing in it's initial endeavors to develop methods to proceed with stock assessments for turtles 

and we are optimistic that this will be realized. It is the TEWG's desire that the results published in 

this report be interpreted with caution and within the bounds of the assumptions both with the data 

and with the methodologies applies. We wish to note that the exploratory modelling 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 assess the status of sea turtles and particularly progress towards recovery. Much hard work and 

deliberation has resulted in these reports to NOAA. I personally wish to acknowledge all the 

contributions of data and numerous late nights that were necessary to finalize tiis report. 

Considerable amounts of data were contributed that were not collected by NMFS and these cannot 

be used or duplicated without the written permission of the originator. 

Nancy Thompson, Southeast Fisheries Science Center 
Executive Summary 

o The loggerhead stock within the southeast U.S. is subdivided into nesting assemblages based on 

genetic information. The extent that these represent separate stocks is not known. The South Florida 

Subpopulation which is currently the largest in terms of nesting females is stable and may be 

increasing; the status of the Florida Panhandle Subpopulation is unknown. Current trends for the 

Northern Subpopulation indicate this subpopulation has declined since the 1980's and that it is 

unlikely that the recovery goal will be achieved. The Yucatan Subpopulation may be decreasing but 

data are lacking for any time series for this assemblage. 

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 rate of increase in the annual number of nests has accelerated over the period 1986 to 1995. 

Under assumptions that estimated age to sexual maturity and age specific mortality rates are correct, 

preliminary analysis suggest that the intermediate recovery goal of 10,000 nesting females by the year 

2020 is achievable. 

o Empirical estimates of the number of turtles that can be removed f?om the population above natural 

mortality without preventing recovery cannot be made at thus time with the data available. Stranding 

levels in the form of Interim Strandings Limits (ISLs) 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 were derived using a risk averse approach for both species based on 3 and 5 year 

averages of stranding totals for each species. For the loggerhead turtles 5 year running averages were 

estimated and one standard deviation was added to this average to provide the ISL. This approach 

recognizes the uncertainty of the status of this species and accounts for differences in the status 

between subpopulations. Both 3 and 5 year running averages were estimated for the Kemp's ridley 

plus one half the current level of increase observed in the annual number of nests. In this way, it is 

recognized that the rate of increase in strandings for this species should not be greater than the 

measured rate of increase of the population. 

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

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

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

subpopulations and rates of mixing on foraging areas of the loggerhead turtle need to be determined. 

o Without a doubt current management which includes the use of TED'S in the shrimp trawl fishery 

must be maintained for recovery of these stocks to be realized as defined by their Recovery Plans. 

Other sources of mortality must be empirically estimated and for fishery related mortality, the 

optimum method for estimation remains through the placement of observers on vessels. 
Turtle Expert Working Group 

Byles, Richard 

Caillouet, Charles W., Jr. 

NOAA NMFS SEFSC 

Galveston Laboratory 

4700 Avenue U 

Galveston, TX 775 5 1 

Crouse, Deborah 

Center for Marine Conservation 

1725 DeSales Street NW #600 

Washington, DC 20036 

Crowder, Larry B. 

Duke Marine Laboratory 

13 5 Duke Marine Laboratory Road 

Beaufort, NC 285 16 

Epperly, Sheryan 

NMFS SEFSC Beaufort Laboratory 

10 1 Pivers Island Road 

Beaufort, NC 285 16 

Gabriel, Wendy L. 

NMFS NEFSC Woods Hole Laboratory 

166 Water Street 

Woods Hole, MA 02543 

Gallaway, Benny J. 

LGL Ecological Research Associates, Inc. 

1410 Cavitt Street 

Bryan, TX 77801 

Harris, Michael 

Georgia Department of Natural Resources 

2070 Highway 278 

Social Circle, GA 30025 

Henwood, Terry 

NOAAINMFSISEFSC 

Mississippi Laboratories 

P.O. Drawer 1207 

Pascagoula, MS 39568-1207 

Heppell, Selina 

Duke Marine Laboratory 

13 5 Duke Marine Laboratory Road 

Beaufort, NC 285 16 

Mairquez-Milan, Rene 

INP-CRIP Manzanillo 

Playa Ventanas S/N 

A.P. 591, Manzanillo 

Colima, Mexico 28200 

Murphy, Sally 

South Carolina Department of 

Natural Resources 

P.O. Box 12559 

Charleston, SC 29422 

Teas, Wendy 

NO AA/NMFS/SEFSC 

Miami Laboratory 

75 Virginia Beach Drive 

Miami, FL 33 149 

Thompson, Nancy 

NO AA/NMFS/SEFSC 

75 Virginia Beach Drive 

Mami, FL 33 149 

Witherington, Blair 

Florida Marine Resources 

Department of Environmental Resources 

19 100 Southeast Federal Highway 

Tequesta, FL 33469 
*.- - 

CONTENTS 

Citation Information . . 

11 

. . Ordering lnformation 11 

Preface 

Executive Summary iv 

List of Participants v 

Table of Contents vi 

ASSESSMENT OF THE KEMP'S RIDLEY SEA TURTLE (Lepidochelys kempii) 1 

IN THE WESTERN NORTH ATLANTIC 

ASSESSMENT OF THE LOGGERHEAD SEA TURTLE (Cawetta cawetta) 47 

IN THE WESTERN NORTH ATLANTIC 
ASSESSMENT OF THE KEMP'S RIDLEY SEA TURTLE (Lepidochelys kempii) IN THE E WESTERN NORTH ATLANTIC 

A Report of the Turtle Expert Working Group 

June 28, 1996 

(Revised November 1997) 

STOCK DEFINITION AND GEOGRAPHIC RANGE 

Kemp's ridleys (Lepidochelys kempii) occur in the Gulf of Mexico and northern half of the 

Atlantic Ocean (Pritchard 1989; Mkquez 1994) and are assumed to constitute a single stock. 

The range of Kemp's ridleys includes the Gulf coasts of Mexico and the U.S., and the Atlantic 

coast of North America as far north as Nova Scotia and Newfoundland. They also have been 

reported from Bermuda, European Atlantic waters, the Mediterranean Sea, Madeira, the Azores 

and Nicaragua (Marquez 1994; USFWS and NMFS 1992). 

Adult Kemp's ridleys are essentially restricted to the Gulf of Mexico (Chavez 1969; National 

Research Council 1990; USFWS and NMFS 1992). Adult-sized ridleys are occasionally found in 

Eastern U.S. waters but rarely nests outside the Gulf (i.e., in Florida, South Carolina and North 

Carolina, Meylan el a1 1990; Palmatier 1993). Most Kemp's ridleys nest at the Rancho Nuevo 

nesting beach (Pritchard 1989). Pritchard and Marquez (1973) reported that adult females 

tagged at the nesting beach were captured by shrimp fishers in nearly equal proportions from the 

northern and southern Gulf. The distribution of adults appears to be mostly in the shallow, near 

shore zone (i.e., waters less than 50 m deep) as evidenced by satellite monitoring (Byles 1989; 

Renaud 1995). 

Post-hatchling pelagic stages are assumed to become entrained in eddies within the Southwestern, 

Western and Eastern Gulf of Mexico (Collard 1990; Collard and Ogren 19901, where they are 

dispersed within the Gulf and Atlantic by oceanic surface currents until they reach about 20 cm in 

straight carapace length (SCL), at which size they enter coastal shallow water habitats (Ogren 

1989). We define benthic immature Kemp's ridleys as those individuals of 20-60 cm straight 

carapace length (SCL) corresponding to the sizes generally found in the near shore habitat and not 

yet adults. Typically, they inhabit estuaries and adjacent near shore coastal waters (Ross et al. 

1989; Rudloe et al. 1989; Epperly et al. 1992; Schmid and Ogren 1992; Epperly et al. 1995% b). 

In the Gulf of Mexico, the habitat occupied by ridleys has been characterized by Hildebrand 

(1982) as white shrimp (Penaeus setiferus) and portunid crab habitats. 

Some investigators believed Kemp's ridleys that exited the Gulf of Mexico were lost to the 

population, especially if they traveled to European Atlantic waters (Carr 1963, 1980; Hendrickson 

1980; Brongersma 1982). However, others (Pritchard 1989; Collard and Ogren 1990; Schmid 

1995) have concluded that the U.S. East coast is important developmental habitat for the benthic 
immature stages which return to the Gulf at some later period in their development. In 1994, a 

wild Kemp's ridley, tagged on the Southeast coast of Florida in 1989, nested at Rancho Nuevo 

(Schmid 1995). Prior to that time, the only indications of possible return to the Gulf of Mexico 

were seasonal north-south migrations along the Atlantic coast (Ogren 1989; Epperly et al. 1995a) 

and observations that the average size of Kemp's ridleys increases southward along the Eastern 

U. S. (Carr 1980; Henwood and Ogren 1987; see Epperly et al. 1995b for review). 

Kemp's ridley is the most seriously endangered of the sea turtles. Yet, its three to six million year 

existence (Bowen et al. 1991) shows it had been a very successful, well adapted species. Five 

decades ago, Kemp's ridley was very abundant in the Gulf of Mexico (USFWS and NMFS 1992). 

Its precipitous decline since 1947 (Figure 1) most likely was caused by human impacts at the 

primary nesting beach near Rancho Nuevo, Tamaulipas, Mexico and at seal (Francis 1978; 

Hildebrand 1963; Chavez et al. 1968; National Research Council 1990; USFWS and NMFS 

1992). 

The Mexican government began protecting the Rancho Nuevo nesting beach from poachers in 

1966, but the nesting population continued to decline (Chavez et al. 1968; Marquez et al. 1989; 

Marquez 1994). The U. S. Government listed Kemp's ridley as endangered in 1970 (USFWS and 

NMFS 1992), and the species received additional protection under the U.S. Endangered Species 

Act of 1973 (ESA), the Convention on International Trade in Endangered Species of Wild Fauna 

and Flora (CITES), and various laws, regulations, decrees and acts promulgated in Mexico 

(Mhrquez 1994). In 1978, U.S. agencies joined Mexico in protecting the nesting beach and in 

other efforts to prevent Kemp's ridley extinction (Woody 1986, 1989; Marquez 1994), but the 

number of nesters at Rancho Nuevo continued to decline by about 4% annually (Frazer 1986). In 

the mid-1980's, the number of Kemp's ridley nests laid at Rancho Nuevo began to increase in what 

now appears to be an exponential trend of 11% per year (Figure l), but the species' population is 

still far below historical levels. The annual level of total human-caused mortality and serious 

injury (benthic imrnatures and adults) was estimated by the National Research Council (1990) to 

be 575 to 5,750 Kemp's ridleys. 

The Kemp's ridley population appears to be in the early stage of exponential expansion. Due to 

the multi-year time lag between release of hatchlings and appearance of neophyte nesters of the 

same cohort (year-class), nesting beach management pra~ctices which enhance the annual 

production of hatchlings do not affect the annual number of nesting females or nests until years 

later. This time lag is determined by how long it takes the turtles to mature and nest for the first 

time, and it is not necessarily of the same duration for all surviving nesters of the same cohort. In 

contrast, management practices that reduce human-caused mortality and serious injury in turtles at 

sea can be expected to have both immediate and long-term effects which increase the annual 

number of nesters. Reductions in human-caused mortality and serious injury rates at sea (e.g., 

with Turtle Excluder Devices) can be expected to increase survival in all post-pelagic stages, 
Population Trends 

The forty thousand nesting females estimated from a siigle mass nesting emergence (arribadu) in 

1947 reflected a much larger total number of nesting turtles in that year than exists today, perhaps 

by more than a 100 times (Carr 1963; Hildebrand 1963). For the two decades thereafter, a 

precipitous decline occurred in the Kemp's ridley population (National Research Council 1990; 

USFWS and NMFS 1992), during which production of hatchlings probably was low if not nil due 

to near total exploitation of eggs (Hildebrand 1963, 1982). Such exploitation certainly must have 

reduced, if not eliminated, recruitment of neophyte nesters to the beach at Rancho Nuevo for 

many years. The sharp decline in the Kemp's nesting population continued through the 1960s, as 

depicted by the decline in reported annual number of nests, imd was followed by a decelerating 

decline to the minimum nests observed in 1985 (Figure 1). 'To a large extent, the annual number 

of nests reflects the number of nesting females, including new recruits (neophytes) to the nesting 

population. The decline in nests fiom 1966 through the mid 1980s can be explained by a decline 

in nesting females. There should have been a marked reduction in new recruits to the nesting 

population, based on the paucity of previous hatchling production. The decline in nesting females 

reflected both natural and anthropogenic mortalities, the latter of which predominately occurred at 

sea. The Kemp's Ridley Recovery Team (USFWS and NMFS 1992) attributed the decline to 

years of egg exploitation and the growth of the shrimp fishery post World War 11. The brown 

shrimp (Penaeus aztecus) fishery in the Western Gulf was expanding during the period of decline 

in the Kemp's ridley population (Klima et al. 1982; Brunenmeister 1984; Rothschild and 

Brunenmeister 1984; Caillouet et al. 1991; Nance 1992). 

The Kemp's ridley population decline might well have been irreversible had the Mexican 

government not begun to protect the nesting beach fiom poachers, beginning in 1966. When, in 

1978, the U.S. Government began cooperating with Mexico in protecting the Rancho Nuevo 

beach, the annual number of hatchlings released more than doubled (Figure 2). The Magnuson 

Fisheries Conservation and Management Act of 1976 (MFCMA) extended U. S. fisheries 

jurisdiction to 200 NM offshore. Mexico extended its fisheries jurisdiction in 11 978, thereby 

reducing shrimping effort in Mexican waters to the extent represented by exclusion of U.S. 

shrimpers. 

Years 1987-1995 represent a period of accelerating increase in annual number of nests (Figure 1). 

We expect this upward trend in annual number of nests to continue. And it could accelerate, 

provided hatchling production and protection of sea turtles at sea are maintained. 
ANNUAL HUMAN-CAUSED MORTALITY AND SERIOUS INJURY 

Pursuant to Endangered Species Act Amendments of 1988, the National Research Council (1990) 

assessed the status of the Kemp's ridley sea turtle. 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 to 5000 benthic immature 

and adult Kemp's ridley mortalities were attributed to shrimp trawling, prior to implementation of 

TED regulations. This compares to 75 to 750 estimated mortalities due to all other known human 

causes (NRC 1990). 

Among the findings of the National Research Council (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, stranding 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; (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. Furthermore, the Kemp's Ridley Recovery Team (USFWS and NMFS 

1992) concluded that incidental mortality caused primarily by commercial shrimping had 

forestalled recovery of the species by preventing adequate recruitment to the breeding population. 

In addition to shrimp trawling, National Research Council (1990) identified many natural and 

anthropogenic causes of sea turtle mortality in waters and on the shores of the U.S. and Mexico 

including predation, parasitism, diseases, environmental changes, effects of beach manipulations 

on eggs and hatchlings, collisions with boats, entrapment in fishing nets and other gear, dredging, 

oil-rig removal, power plant entrainment, ingestion of plastics and toxic substances, and incidental 

capture in shrimp trawls. It is unlikely that any major non-shrimping cause of sea turtle mortality 

could have escaped detection during the many years of study of factors causing sea turtle injury 

and mortality at sea (National Research Council 1990, USFWS and NMFS 1992). 

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. 24247-24262,28 June 1987). However, TED regulations were not implemented 

until May 1989, and TEDs were used only sporadically until May 1990 (McDonald 1990; Crouse 

et al. 1992; Henwood et al. 1992; Crouse 1993b; Weber et al. 1995). Year-round TED 

requirements became effective essentially in May 199 1. After December 1992, year-around use of 

TEDs was required by most trawlers operating in Southeastern U.S. waters (57 FR 48861-48871, 

Sept. 8, 1992). 

In the northwestern GuK Caillouet et al. (1996) found no change in the strength of the statistical 

association between nearshore shrimping intensity and sea turtle (all species) stranding rates in 
TED years as compared to pre-TED years (Caillouet et al. 1991, 1996). Gallaway et al. (1995), 

following National Research Council (1990) methods, compared the reduction in strandings 

between the open and closed parts of the Texas shrirnping season during pre-TED (1986-89) and 

post-TED years (1990-94). The magnitude in the reduction in strandigs from the open to the 

closed period did not differ between post-TED and pre-TED years. 

The Kemp's ridley population declined and remained very low during the early 1980s. Nesting 

stopped declining in the mid 1980s, before TEDs were required in U.S. shrimp trawls. The halt in 

the decline in the mid 1980s and the subsequent early stages of recovery of the Kemp's ridley 

population likely can be attributed to a number of factors, including protection of the nesting 

beach, associated enhancement of hatchliig production, decrease in shrimp fishing effort in 

Mexican waters when the U.S. shrimping fleet was prohibited fiom shrimping in the late 1970s, 

and the overall decline in the Mexican shrimp fleet. 

The shrimp fishery in the Gulf of Mexico 

Threats to mating and nesting adult Kemp's ridleys from shimping activity in Mexican waters 

have been decreasing. Shrimping effort off Rancho Nuevo was reportedly high until the U.S. and 

Mexican governments negotiated a bilateral agreement in 1976, phasing out U.S. shrimping in 

Mexican waters through 1979 (Iverson et al. 1993). Clearly, the reproductive component of the 

Kemp's ridley population was vulnerable to capture by this fishery. After 1979, U.S. vessels 

continued to shrimp off Mexico, although illegally and at reduced levels, through the mid 1980s, 

when the U. S. enforced the Lacey Act. 

The reduction in effort in Mexican western Gulf waters caused by the departure of the U.S. fleet 

was hrthered by the subsequent decline in the Mexican fleet. Currently, the entire Mexican fleet 

in the Gulf is estimated to number approximately 600 vessels, many of which are not presently 

fishing. 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 observations of ten 

to twenty trawlers operating fiom 2 to 5 mile off the beach at night. During 1995, Mexican 

waters of the Gulf closed to shrimping in concert with the Texas closure, from May 15 through 

July 15. 

Significantly increased strandings, including a very high proportion of Kemp's ridleys, occurred 

throughout the southeast in 1994, but particularly in Texas and Louisiana. This was followed by 

elevated strandings again, but not as high, in 1995. Technical and intentional violations in gear 

installation, intense pulses of fishing at certain times and places, and use of TEDs with less 

effective at excluding turtles (NMFS, 1994) were all implicated as causes of the increased 

mortality. In addition, it is possible that an increase in the number of benthic immature Kemp's 

ridleys, due to the combined factors of increased hatchling production and TEDs in most US 

shrimp trawls, may be providing an increased pool of tur;hles. 
Other sources of mortality include: Flounder fishing with otter trawls, off the coasts of North 

Carolina and Virginia, and gill netting for a number of species have also been implicated in high 

rates of sea turtle mortality on a more localized scale than shrimp trawling (National Research 

Council 1990). TEDs are currently required seasonally in all flounder trawls off North Carolina 

and Virginia and observers are documenting take rates for that fishery in the rest of its range. 

POPULATION SIZE, VARIABILITY AND TRENDS 

Kemp's ridley sea turtles provide a unique opportunity for population modeling because although 

they are endangered, we can monitor numbers of nests and hatchling output for nearly the whole 

population. Our approach, then, was to conduct statistical estimation procedures to estimate 

unknown parameters (e.g. stage-specific survival) for a preliminary population model. Other 

parameters were derived from empirical data and sensitivity analyses were conducted. Finally, 

population projections under various scenarios were conducted. 

The primary indices of population status of Kemp's ridleys are currently limited to annual 

numbers of nests, observed nesting females, eggs and hatchlings produced at Rancho Nuevo 

(Mbrquez 1994)' and size and imputed age composition and abundances of beach stranded turtles. 

The data include a 30-year time series of nests counted and hatchlings released (Table I), and an 

empirical estimate of mortality rate for benthic immature turtles from a catch curve analysis of 

strandings (See 7 below). 

The values we used as fixed parameters were: adult female remigration rate, the mean number of 

eggs per nest, the mean number of nests per female per year, female age at maturity, sex ratio in 

the population, and survivorship of benthic immature turtles. We considered published and 

unpublished information in developing the following values for the fixed parameters: 

1. The remigration rate of females to the nesting beach was estimated from mark-recapture 

data (Marquez et al. 1982). We updated these data with current information provided by R. 

Marquez. Twenty percent of the turtles nest annually, 60% nest every 2 years, 15% nest every 3 

years, and 5% nest every 4 years. This results in a mean remigration rate of 2.05 years, which we 

rounded to 2 years. 

2. The mean number of eggslnest incorporated into the calculations was 100.8 noted by the 

Kemp's Ridley Recovery Team (USFWS and NMFS 1992), which we truncated to 100 eggdnest. 

3. Nests/female/season were calculated by Rostal(1991) to be 3.075 fiom a physiological / 

ultrasonographic study. Pritchard (1990) estimated 2.3 1 nests/female/season and a recent pit-tag 

study by Marquez (unpublished data) suggests the mean is 1.8 nests/female/season. We used the 

mean of means fiom these studies (2.4 nestdfemale/season), which we rounded to 2.5 

nests/female/season. 
4. The age at maturity has been estimated by Marquez (1972) at 5-7 years. The Kemp's Ridley 

Recovery Team (USFWS and NMFS 1992) thought that estimate too low, but offered no better 

estimate. Zug et al. (1 997) estimates Kemp's ridley maturity at 10 to 15 years based on analysis 

of cross-sections of humerus bones. He preferred the upper end of that estimate. From these 

data, we placed bounds of 7 to 15 years on age at maturity. Model runs were conducted with 8, 

10 and 12 years input as the age to maturity parameter in order to test the model's sensitivity to 

the parameter. 

5. A linear regression analysis was run to estimate years in the pelagic stage from hatchling 

production and number of stranded juveniles 1,2,3, and 4 years later in the stranding records. 

The analysis used stranding records for 20-30 cm, 30-40 cm, and 20-40 cm turtles. The resultant 

estimate of 2 years in the pelagic stage was used in the model. 

6. The sex ratio in wild, benthic immature turtles, as determined by testosterone assay, has been 

found to approximate 1 : 1 at several locations throughout the U.S. @. Owens, pers comm), but 

Stabeneau et al. (1996) found a sex ratio significantly skewed to females among stranded turtles. 

This female skew could be due to nesting females spending more time nearshore than males. We 

assumed a 1: 1 sex ratio for this analysis. 

7. Instantaneous mortality rate, Z, for benthic immatures was estimated using "catch curve" 

analysis of age composition derived by converting size distributions of post-pelagic stranded and 

measured Kemp's ridleys recorded in the Sea Turtle Stranding and Salvage Network (STSSN) 

database (1980-1995) to age distributions with a von Bertalanffjr growth curve solved for age at 

size (see Sparre and Venema 1992). Curved carapace length (CCL) is the predominate 

measurement in the stranding database, so it was converted to straight carapace length (SCL), 

using a linear regression equation derived from stranded Kemp's ridleys for which both SCL and 

CCL were measured. When only SCL had been measured it was also included in the data 

analyzed. Size composition of post-pelagic stages fully recruited to the strandings was assumed 

to represent the size composition of the population of benthic immature and adult Kemp's ridleys. 

It was also assumed that growth of male and female Kemp's ridleys can be represented by the 

same growth curve. 

Because of the scarcity of mark-recapture data for wild Kemp's ridleys, especially those in the 

benthic immature stage, mark-recapture data for head-started turtles (Caillouet et al. 1995a) were 

used to fit a von BertalanffL growth curve in its In-linear fon-m, but SCL at infinite age was fixed 

at 75 cm instead of finding its value by iteration: 
None of the recaptured turtles had a SCL equal to or greater than 75 cm, so division into zero 

was not encountered in fitting the equation. The resulting von Bertalw equation was solved 

for t, as follows: 

Since some stranded Kemp's ridleys had a SCL equal to or greater than 75 cm, whenever this 

condition was encountered, t was assigned a value of 15, because benthic immatures were 

assumed to fall well below 75 cm in SCL. Otherwise, t was estimated fiom SCL then grouped by 

class intervals into ages represented by integer values. For example, age 1 represented hatchlings 

and pelagic imrnatures less than 1 year old, age 2 represented pelagic imrnatures equal to or 

greater than 1 year old but less than 2 years old, age 3 represented benthic immatures equal to or 

greater than 2 years old but less than 3 years old, and so on. An age-frequency distribution was 

prepared for all year-classes. Year-classes were assigned by subtracting age from the calendar 

year. For example, for calendar year 1990, a turtle of age 1 year represented the 1989 year-class. 

The natural logarithm of the total number of turtles at each age 2-6 years was regressed on age to 

estimate Z for all year-class combined. This estimate of % was 0.38 and was rounded to 0.4. It 

may have been biased by small sample sizes, pooling age distributions across year-classes, and the 

unknown relationship between strandings and total mortality. 

8. We assumed 70% survival of eggs to hatchlings based on recent survival rates in the corrals 

at Rancho Nuevo. 

Parameter Estimation 

In base model runs, the number of female hatchlings produced each year was entered as a starting 

cohort size, assuming a 5050 sex ratio and no surviving hatchlings fiom other sources. Numbers 

of turtles in each age class in each year were projected fiom hatching production, assuming a two 

year pelagic phase with an associated mortality rate, and an eight-year benthic immature phase (in 

the case of age at sexual maturity of 10). Recruitment to the nesting population was partial in the 

first year of sexual maturity, determined by the rlemigration rate for the fblly recruited population. 

Thus, at age of sexual maturity of 10, with a population remigration rate of 2 years, half of the 10 

year-old turtles were added to the nesting population in the year they recruited. This assumption 

was required in order to scale the observed number of nests to the estimated nesting population, 

without knowing the percentage of new recruits in the population. The remaining adult 

population (e.g., age 11+ for age at sexual maturity of age 10) was modeled without additional 

age structure: number of nesting females at time t+l was a hnction of number of adult females at 

time t and number of female recruits at time t, adult mortality rate, and remigration rate. The total 

number of predicted nests each year depended on constant assumed rates of nests produced per 

female, on the number of adults surviving from the preceding year, and the number of recruits to 

the nesting population. 
Starting population sizes were derived from estimates of nests and hatchlings in 1966 (Table 1). 

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. 

The model was updated on an annual basis, keyed to the reproductive rather than calendar year, 

i.e., estimates of stock size are as of the season of peak nesting rather than as of 1 January. 

Hatchling production was assumed to enter the population at that time as well. 

Mortality rates of hatchlinglpelagic turtles and adults, and a post-1989 decrease in benthic 

'mature and adult mortality rates (see results) were estimated by comparing predicted nests to 

observed nests (1978-1995), using least-squares fit criteria (Excel 5.0 SolverTM algorithm). We 

chose to model the 1978-1995 period because it included the most complete census of the nesting 

region. 

Initial Model Runs 

Initially, a single mortality rate per life history phase (hatchling, benthic immature and adult) was 

assumed to apply over the duration of the 1966-1995 time series. But inspection of residuals 

fiom a model estimating hatchling and adult mortality rates for the entire time series (combined 

with the fixed benthic immature mortality rate estimated empirically) revealed a strong trend in 

residuals &om maximum negative values in 1989 to maximum positive values in 1995 (See 

Appendices). To address this problem, an additional term was fitted to the model as a single 

multiplier operating on both benthic immature and adult instantaneous total mortality rates during 

the 1990-1995 period. The period was chosen to reflect years in which TEDs were required. But 

it must be noted that the affect reflected in this term may arise fiom any combination of factors 

operating over the 1990- 1995 period which in aggregate would decrease mortality rates, e.g., 

TEDs, favorable environmental factors, reduction in anthropogenic mortality fiom other sources, 

changes in fishery patterns driven by non-TED factors, etc. 

Trends in residuals after inclusion of the post-1989 effect may indicate an increase between 

1978-1984, with a possible decrease thereafter. This qualitative serial autocorrelation could be a 

direct result of the model formulation, in which number of nesters at time t+l is a function of 

number of nesters at time t. It also could arise from the application of single stage-specific 

survival rates over the entire 1966- 1989 time series, when more realistically, survival rates varied 

over time. Thus, residual patterns indicate a rough but not especially acceptable fit. Key 

diagnostic indicators of model performance such as matrices of correlations between parameter 

estimates were not considered for this analysis. Consequently, there is a significant possibility that 

too many parameters are being estimated given the model structure and the solution presented 

here may not be unique. 
Mortality rates estimated fkom the model indicate a 44% annual proportionate survival rate of 

hatchlings during ages 0 and 1. This means that if 100,000 hatchlings were released, only 44,000 

would reach age 1, and 19,400 (0.44 * 44,000) would reach age 2. In the benthic immature 

phase, proportionate survival rates of 66% were estimated fiom strandings data and input into the 

model. Of the 19,400 hatchlings reaching age 2, 12,800 would reach age 3 (0.66 * 19,400). With 

an age at sexual maturity of 10, turtles would remain in the benthic immature phase for 8 years 

(from ages 2 to 10). The 12,800 turtles at age 3 would decline to only 700 by age 10. Survival 

rate of sexually mature turtles was estimated at 84%. Thus, less than 600 turtles would remain by 

age 11, and less than 300 by age 15. With a 5050 sex ratio, less than 350 females would survive 

fiom the initial 100,000 hatchings to reach sexual maturity, and fewer than 150 would be alive at 

age 15. 

The survival rate for benthic immature and adult turtles was estimated to have increased 

substantially after 1989. The model was formulated so that lower initial survival rates would be 

increased faster than higher survival rates. The change in the rate was modelled as a multiplier of 

instantaneous mortality, (estimated as about a 33% reduction in instantaneous mortality). 

Instantaneous rates of mortality and post-1989 multipliers for those rates are reported in the 

Appendices. This meant that although mortality in the pelagic phase would continue during ages 

0 and 1 at the same rate, survival rate of immature individuals was estimated to have increased to 

75% and survival rate of adults was estimated to have increased to 89%. Thus, although 19,400 

individuals would reach age 2, now 14,600 would reach age 3. Those 14,600 individuals would 

decline to 2,000 by age 10. With an adult survival rate of 89%, 1,800 turtles would remain by age 

11, and 1,100 would remain by age 15. With a 5050 sex ratio, less than 1,000 females would 

survive fiom the initial 100,000 hatchlings to reach sexual maturity, and perhaps 550 would be 

alive at age 15. Of those 550 to 1,000 females, half would nest each year. 

It should be noted that although large numbers of turtles may occur at immature stages, (e.g., the 

19,400 turtles at age 2 and the 14,600 turtles at age 3 in the example above) the number in a year 

class that survive to contribute to the nesting population in any year is small. Under recent 

improved survival conditions, 500 out of 100,000 hatchlings would contribute nests 10 years 

later, and only 225 would contribute 15 years later. 

Based on model results, the abundance of benthic immature turtles (ages 2 - 9) has trended 

upward in three phases since 1966 (Appendix 1). Between 1966-1977, abundance initially 

increased fairly rapidly then leveled off, as hatchling production in 1966 began to contribute to the 

benthic immature population. Hatching production averaged about 23,000 per year between 

1966-1977. A second period of increase and subsequent leveling off occurred between 

1978-1989, as hatchling production trended upward over the 1976- 1979 period, averaging 

49,000 per year between 1978-1987. The third period, fiom 1990 on, was one of steady 

increase. It began in 1990 as hatchling production increased fiom about 62,000 in 1988 to 84,000 

in 1993, and as estimated survival rates of immature turtles increased as well. Abundance of 

benthic immature turtles in 1995 was estimated at 54,000 individuals, 55% of which were age 2 -3 

and <2% of which were aged 9. 
A qualitative relationship may emerge between estimated number of dead benthic immature 

animals and numbers of strandings fiom corresponding size categories (Appendix 1). Relatively 

uniform and widespread monitoring of strandings occurred after 1985, and indicates a variable 

pattern without trend between 1986- 1993, but sharp increases in 1994- 1995. Model estimates of 

numbers of dead benthic immature individuals were fairly stable between 1986 and 1991, and 

trended upward thereafter, as a function of increasing population size. Any statistical relationship 

would be driven by the co-occurrence of both maximum strandigs levels and maximum estimates 

of deaths in 1994- 1995. The increasing number of deaths estimated between 199 1 - 1993 is not 

reflected in an increase in strandings during that period, however. 

Abundance of adults declined from the population that produced about 6000 nests in 1966 to a 

population that produced 924 nests in 1978 and a low of 702 nests in 1985. The historical 

trajectory of adult abundance tracks trends in nest abundance, fiom an estimate of 9,600 in 1966 

(fi-om conversion of nests to population size, using ratios as described above) to 1050 in 1985 

(Appendix 1). The adult population has increased to about 3,000 in 1995, consistent with 1,930 

nests observed in 1995. 

Based on model simulations, the percentage of neophyte nesters in the population increased 

steadily fiom just 13% to about 56% over the 198 1-1 989 period. The lower levels estimated for 

1990 and 1993 (30%) are the result of low hatchling production in 1980 and 1983 projected to 

age at sexual maturity. Estimates for other recent years have been in excess of 40%. Direct 

observations of percentage of neophyte nesters from nesting beaches may be considered upper 

bounds (because some animals may not have been tagged during previous nesting seasons), and 

have ranged fiom 23- 28% in 1990-1991 to 38-41% in 1992-1994. There thus is fair 

correspondence between recent direct observations and model estimates. 

I 

Sensitivity Analyses: Number of nests in 1966 

Because the estimated adult survival rate must be consistent with the decline in number of nesting 

females estimated from nests in 1966, and because number of nests in 1966 was an approximate 

estimate rather than a complete census, the effect of a +/- 25% difference in 1966 nest level was 

examined. The model was re- fitted using 1966 nest levels of 7500 and 4500 (Appendix 2, 

Appendix 3). 

If the 1966 nest level were closer to 7500 nests (Appendix 2), estimated adult survival would be 

lower, to produce a steeper decline between 1966 and 1978. That steeper decline in the remnant 

nesting population after 1978 would be compensated for by a slightly increased contribution of 

neophytes in order to obtain nest production thereafter. The slight increase in contribution by 

neophytes would be driven by an increase in estimated survival rate of pelagic juveniles in earlier 

years of the recruiting cohorts' life. Lower increases in benthic immature and adult survival rates 

after 1989 would be required to match increased nest levels, because the standing stock of benthic 

immature individuals (and subsequent contributions to the nesting population) would be higher 
due to increased pelagic survival. The combination of higher pelagic juvenile survival rates with 

slightly lower survival rates of benthic immature and adult turtles would lead to a similar (albeit 

marginally smaller) recovery rate in future years. The residual sum of squares for this model was 

higher than the baseline run of 6000 nests (age at sexual maturity = 10). 

Similarly, if the 1966 nest level were closer to 4500 nests (Appendix 3), adult survival would be 

higher and survival of pelagic juveniles would be correspondingly lower, with mechanisms exactly 

opposite from those described above. The combination of lower pelagic juvenile survival rates 

and higher benthic immature and adult survival rates would lead to a similar (albeit marginally 

larger) recovery rate in fbture years. The residual sum of squares for this model was lower than 

the baseline run of 6000 nests, and residual patterns showed less pronounced trends over time. 

i Sensitivity Analyses: Age at Sexual Maturity 

Estimates of survival rates of benthic immature turtles from strandings data were independent of 

age at sexual maturity. The effect of increasing the age at sexual maturity (relative to the base 

level of 10) is to increase the number of years within the cohort's lifespan in which that rate is 

applicable. The aggregate mortality between pelagic juvenile and adult phase of a cohort is 

increased. The fitted results from this formulation indicate that the minimum residual sum of 

squares is obtained through an increase in pelagic juvenile survival, to compensate for aggregate 

decreased survival of benthic immature turtles over the longer immature period. 

In a model formulation assuming age at sexual maturity of 12 years, estimated proportionate 

survival of pelagic immature turtles increased to 71% from 44% estimated for age 10. Estimated 

survival in the pelagic phase was higher than survival in the benthic immature phase, and on the 

same scale as survival in the adult phase (85%). The increased pelagic survival rate together with 

a greater number of contributing age classes led to greater abundance of benthic immature 

individuals. The increase in survival rate post-1989 was thus estimated to be lower. The 

percentage of neophytes in the nesting population was lower, as cohorts were decremented by 

additional years of mortality before recruiting to the nesting population. Recovery rates were 

similar to the base run for near fbture years, albeit marginally smaller. The residual sum of squares 

was almost three times that of the base level run (Appendix 4). 

In a model formulation assuming age at sexual maturity of 14 years, estimated proportionate 

survival of pelagic immature turtles increased to the constraining upper bound of loo%, and 

unconstrained estimates of survival exceeded 100%. 

If the estimate of survival of benthic immature turtles from strandings data is reasonable, then 

model results obtained by assuming age at sexual maturity of 12 years and greater appear 

unreasonable given the current model formulation. Other formulations, objective knctions or 

estimation algorithms may give more acceptable results for older ages at maturity. Alternatively, 

I 
if survival of benthic immature turtles were overestimated, meaningfbl results would be obtained 

fiom formulations based on older ages at sexual maturity. 

Sensitivity Analysis - Benthic Immature Instantaneous Mortality 

Because our estimate of the benthic immature instantaneous mortality rate may have been biased 

due to small sample sizes, inappropriate growth rate estimation procedures and the unknown 

relationship between strandings and total mortality, we examined the sensitivity of the 10 years to 

maturity model results to changes in benthic immature mortality. We ran the SolverTM algorithm 

with six different benthic immature mortality rates, using a range of benthic immature mortality 

rates that spanned our estimates from the strandings records. The program chose best fit 

parameters (adult instantaneous mortality, pelagic juvenile instantaneous mortality, and the post1989 decrease in mortality) for each benthic immature mortality rate, with the reproductive 

parameters again held constant (Table 2). Adult survival remained constant at 0.84 for all 

sensitivity runs, primarily because we fixed the number of nests produced in 1966 at 6000. 

Pelagic juvenile mortality compensated for most of the change in benthic immature mortality; the 

negative correlation between these two mortality rates resulted in the same number of females 

recruiting to the nesting population for all models. Thus, the residual pattern remained nearly 

constant for each model, although there was a trend toward a better fit as the benthic immature 

mortality rate increased (Table 2). There was almost no change in time to reach 10,000 nesting 

females (Table 3). However, the estimated number of benthic imrnatures in the population was 

highly dependent on the allocation of mortality to the pelagic and benthic immature stages (Table 

4). Higher annual survival rates for benthic imrnatures gave best fit models with lower survival 

for pelagic imrnatures, so many fewer survivors from a cohort actually recruited to the benthic 

immature stage, resulting in a smaller estimated population size. The opposite was true for lower 

benthic immature annual survival rates. Because we do not have empirical estimates for annual 

survival rates of any life stage, any of these best fit models could be closest to the actual mortality 

schedule of Kemp's ridleys. It is extremely tenuous to use this model to estimate the number of 

benthic irnrnatures in the population. 

Projected recovery trajectories were calculated fiom 1995 forward, based on 1995 standing stock 

sine and mortality estimates obtained fiom the regression rnodel assuming age at sexual maturity 

of 10. Future hatchling production also was modeled in density-independent and 

density-dependent forms. Limitations on human intervention and protection of Kemp's ridley 

nests were imposed on the model. Protection of the nests fiom predators at the Rancho Nuevo 

nesting beach is eEort dependent, and will change with time. The biological desirability and 

logistics of protecting a large number of nests by transporting each nest to a corral (a fenced, 

natural sand hatchery on the beach) for incubation will likely limit the number of nests so 
protected. When some peak number of nests is corral-protected, the remaining nests will be 

protected in situ. ~ 

The INP and FWS have suggested that under current planning a maximum of 

probably be protected in corrals, with an additional 2,000 nests protected in 

Nests deposited beyond these 5,000 would be left completely natural 

predator control may become desirable. Hatch rates were assigned 

regimes as follows: 

The density-independent form of the model was based on assumptions of 2.5 nests per nesting 

female per year, 100 eggs per nest and 75% survival of those eggs. The density-d pendent form 

was based on similar assumptions of nests per female and eggs per nest, but allow d current levels 

of 75% egg survival to apply only to the first 3000 nests, the assumed overall level of protection 

obtainable through corral egg protection programs. The next 2000 nests would be protected with 

in situ techniques, with postulated egg survival rates of 50%. The next 5000 nests would be 

relatively unprotected, with egg survival dropping to a hypothetical 30%, under co ditions of low 

predation mortality. At nest levels above 10,000, higher predation mortality woul be obtained, 

with egg survival of a postulated 20% resulting. Trajectories are presented for the more realistic 

density-dependent case, although results in initial years of the projection are identi al for the two 

mechanisms. d 

# MESTS 

0-3,000 

3,001-5,000 

5,001-10,000 

>10,001 

Trajectories were developed for different future mortality levels relative to the 

Survival of the pelagic immature ages was not varied. Survival of benthic 

components was varied by rescaling the 1995 estimates of instantaneous 

multipliers (+/- 10% and +/- 20%): 

Change in instantaneous mortality rate I 

CATEGORY 

Corrals 

In situ: Screening 

In situ: Low predation 

In situ: High predation 

Proportionate Survival Rate 

Scenario -20% -10% Status +lo% +20% 

QJQ 

Pelagic .44 .44 .44 .44 .44 

Benthic immature .80 .77 .75 .73 .71 

Adult .91 .90 .89 .88 .87 

HATCH K.XS 

75 

50 

30 

20 

(%) 
Under status quo mortality rates, the target level of 10,000 nesting females per year is obtained in 

20 15 (Figure F). As of 20 10, only 6 100 females would be expected to nest. By reducing 

instantaneous mortality on benthic immature and adult turtles by lo%, 10,000 nesting females are 

obtained in 2012. A 20% reduction in instantaneous mortality would be required to obtain 10,000 

nesting females by 20 10. 

Recovery times would be eroded by any increase in mortality of benthic imrnatures and adults. A 

10% increase in instantaneous mortality would delay projected recovery to 2020; a 20% increase 

would delay recovery until 2026. If the 33% decrease in instantaneous mortality rate estimated 

for the 1990-1995 period were lost after 1995, 10,000 nesters would not be obtained until 2046. 

Uncertainty 

f These analyses include no quantification of uncertainty associated with estimates of 1995 standing 

stock size and mortality rates, nor with projected hatchling production. The assumptions required 

to implement the model were so extensive that uncertainty around estimates could not be 

quantified. 

The model works to minimize the annual differences between observed nests and predicted nests. 

Both components involve potentially significant sources of error. In the case of observed nests, 

we felt that although it could have been possible that the nesting range for Kemp's ridieys may 

have extended beyond Rancho Nuevo before the establishment of additional camps in 1989- 1990, 

c no reliable estimates were available for that portion of the range in the early part of thk time 

series. Thus, estimates of nests before 1989 would be biased low to an unknown degree. In the 

case of predicted nests, the quantitative relationship between population size and nudber of nests 

produced annually is imprecisely defined. The relationship depends on remigration rate, nests per 

female, age at sexual maturity and sex ratio. Each of these components has associate4 but 

t currently undefined variance. An approximate coefficient of variation for estimates of) population 

size (from nests) which incorporated only variance in remigration rate and number of nests per 

female is about 33%, based on Monte Carlo simulations incorporating potential pattetns in 

variation. 

Model results are highly sensitive to benthic immature survival rate and age at sexual maturity, 

based on results of initial sensitivity analyses. The benthic immature survival rate shows variation 

not only in terms of the variance of the estimate itself, but also as hnction of the area and size 

range serving as the basis for an estimate. Animals likely mature sexually over a widet age range 

than ranges evaluated here, but the extent of that range and the shape of the maturity schedule is 

unknown. 

The model formulation contains several structural features which may contribute to unreliable 

results. The mortality rate of adults is determined at least in part by the rate of decline of nesting 

females between 1966 and 1978. The mortality rate of pelagic juveniles is critically dependent on 
f stable population dynamics processes ove 

k size and parameter estimates. 
Strandiigs should not increase at a faster rate than population size. Further analysis 

residuals fiom this analysis may be warranted, as well as a statistical analysis of dfle 

the slopes; however, the low sample siie may be prohibitive. Strandiigs in 1996 m 

determine if the low strandings in 1991 and 1992 or the high strandings in 1994 an 

more typical under current management. 

Interim Stranding Limit (ISL) I 

We concluded that estimation of the maximum number of individual Kemp's rid1 

can be taken incidentally to commercial fishing activities without preventing the 

species cannot be made at this time. Estimates of total population size and curre 

were found to be too sensitive to current data deficiencies and estimation model 

reliable. Moreover, any apportioning of those mortality rates to commercial fis 

cannot be made from available data. 

We also concluded that the best proxy for estimation of take at this point should b 

fiom strandings data and other direct empirical observations of stock status. For 

those empirical observations presently consist of hatchling production and nest 1 

represent an imperfect but directly observable quantity, although quantitative re 

between strandings levels, stock size, and mortality rates (including their fishe 

components) remain undeveloped. We anticipate that these relationships wil 

future, and consequently only recommend use of the proposed proxy appro 

year. 

We identified an Interim Stranding Limit (ISL) based on strandings during the re 

population growth assuming these strandings are sustainable in the short term. 

the species will continue to recover at a reasonable rate if stranding levels don' 

that are indexed to increases in nesting population (as a proxy for total popul 

conservative risk-averse manner. As a precautionary measure, stranding lev 

increase faster than one-half the observed rate of increase in nests. Observe 

nests is estimated as the slope of the natural log of nest levels over the past 

transformed to a linear (proportionate) scale. The annual percentage incre 

two. The resulting percentage increase is applied to recent stranding level 

mean + one standard deviation above the mean) of stranding levels over t 

This approach contains substantial risk of not being responsive to chan 

juveniles (especially at early ages) until those individuals have recruite 

Additionally, the extent to which this approach is risk-averse has not be 

manner in which the performance of Potential Biological Removals (PB 

marine mammals). This fUrther supports the use of this proxy for a sin 

(proxy runs Table 5) 
CONCLUSIONS 

1. The Kemp's ridley population appears to be in the early stage of exponential 

the period 1986 to 1995, the rate of increase in the annual number of nests acc 

upward trend should continue with increased hatchling production and protecti 

use of turtle excluder devices. 

2. Si@cantly increased strandigs occurred throughout the southeast and in p 

and Louisiana in 1994 and 1995. Technical and intentional violations in TED 

intense pulse fishing and use of less effective TEDs were all implicated as cau 

mortality. It is also likely that an increase in the number of benthic immature 

resulting fiom increased hatchling production and the use of TEDs may be p 

pool of turtles available for all sources of mortality including bottom trawls 

3. Adults declined fiom a population that produced an estimated 6,000 nests in 

population that produced 924 nests in 1978 and a low of 702 nests in 1985. In 

of nests increased to 1,930. 

4. The estimation of the maximum number of individual Kemp's ridley sea turtl 

taken incidentally to commercial fishing activities without preventing the recov 

cannot be made at this time because of data deficiencies. Strandings represent 

directly observable quantity, although relationships between stranding levels, s 

mortality rates remain unknown. An Interim Stranding Limit (ISL) was devel 

year running averages of strandings which should not increase annually at a r 

population rate of increase. A risk averse approach includes using one half 

annual rate of increase in nests to calculate the ISL. The extent to which t 

averse needs to be evaluated, particularly relative to benthic immature mo 

should only be used as a proxy for one year. 

5. Modeling efforts suggest that survival rates of benthic imrnatures and adults 

post 1990 to balance egg inputs and number of nesting females. This supports 

effect, but we have not yet documented such an effect empirically. 

6. It is currently projected that this species could reach the intermediate recov 

nesting females by the year 2020 given that the assumptions of age to sexual 

specific survival rates are correct. The extent that these assumptions are viol 

this intermediate goal is achieved. However, if the current annual rate of in 

continues, this goal can be achieved. Note that much information is missing 

projections and sensitivity analysis give very different results with different values. 

7. The increase in hatchling production at Rancho Nuevo in recent years is the re 

improvements in turtle camp operations, which produced higher rates of collecti 

still small but increasing number of nesting females, increased protection of thos 
increased hatch rates. If the annual number of nesters continues to increase, and 

continues to expand, with a consequent increase in number of nests, it will take 

efforts at the turtle camps to assure that these positive indications of initial reco 

increased hatchling production. 

8. Continued, and perhaps expanded, human intervention through protection of 

Nuevo nesting beach and other nesting beaches is necessary for recovery during 

hture, otherwise the gains that have been made could be lost. We are concerne 

encroachment, access and potential development that could adversely impact th 

nesting beach, especially the planned expansion of fishing facilities and extension 

Intracoastal Waterway from the U.S. to Tampico, Mexico (USFWS and NMF 

9. The status of this species and ISL needs to be evaluated annually. 

RECOMMENDATIONS 

Many of the following recommendations will produce data necessary to improve th 

of population and assessment models being applied to sea turtles. Reduction of th 

uncertainty in these parameters will significantly increase the value of models and 

model results. 

1. Long-term, in-water indices of abundance are needed to identifjl relative abun 

turtles, particularly Kemp's ridleys, over time and to detect changes in size comp 

implications regarding recruitment. Studies should be conducted in coastal waters 

coast of Florida, in Louisiana, in Texas, in Parnlico Sound in North Carolina, in 

and in one northern embayrnent. All turtles caught should be tagged with PIT t 

flipper tags, and should be scanned for existing PIT tags. Standardized catch p 

(CPUE) methods should be employed to provide site-to-site and long-term co 

Projects should be consistent and long-term, with data collected in a statistically d 

manner. Habitat use and tracking studies conducted as adjuncts to abundance ind 

should not compromise the effectiveness of the primary goal of abundance in 

which provide an assessment of trends in juvenile abundance should be given 

information on trends that will not be reflected immediately on nesting beaches. 

2. Information is needed regarding the location of mortality of stranded sea turtles 

possible causes of strandings. Oceanographic conditions affecting stranding rates 

decomposition rates and oceanic scavenging) should be examined to determine se 

landfall patterns and to identifjl conditions or areas where carcasses would not be 

randiig network index areas within the southeast region should be developed 

adized coverage by the STSSN. The statistical comparability of the index 

considerations in the identification and monitoring of the study sites. 

19 
1 

4. Development of standardized PIT tag techniques and equipment should be developed. A pit 

tag registry should be initiated and maintained. I 

5. Kernp's ridley nesting females should be intercepted on the beach at levels appr aching 100%. 

Saturation tagging should be conducted, hth PIT tags. AU females should be sca ed for 

previously placed PIT tags. 1 

I. 

6. PIT tag scanners should be isstied to representative STSSN participants in eacdstate to 

maximize the mark-recapture components of long-term sea turtle studies. 1 

II 

7. Efforts to expand the shrimp vessel observer program should continue. 

with industry to assess sea turtle catch per unit effort in various areas, 

embayments, nearshore waters and offshore waters to determine 

and seasons in which TED regulations need to be modified. 

8. Age and growth studies should be conducted for Kemp's ridleys in both the Gu f and Atlantic. I 

9. Physiological studies to determine the effects of forced and repeated submersio should be 

continued. t 

10. Aerial surveys of shrimp vessels should be conducted to provide resolution on the distribution 

of fishing effort. 1 

11. The utility of population models could be increased through improved ag 

estimates for ridleys and loggerheads. Life history data collection from nesti 

stranding monitoring, and in-water abundance surveys should support this 

data on sex ratios, maturity, breeding condition and accurate sizes can be 

examination of stranded carcasses. Laparoscopic studies could yield the 

nesting population and more saturation tagging is needed to assess femal 

remigration. 

12. Shrimp fishing effort reduction measures, including closures, gear restrictions, or other 

measures should be evaluated to hrther reduce nearshore shrimping effort for incr ed sea turtle 

protection, while remaining consistent with state management goals and local shri f ping industry 

interests. 

13. Other sources of fisheries-related mortality that have high nearshore effects, s 

crab trawling, long lines, purse seines, recreational hook and line, and bottom 

to be evaluated. 

14. Additional assistance should be provided to Mexico in conducting a thorou 

analysis of the Kemp's ridley nesting beach data. This should include counts o 

hatchlings and nesters, .mark-recapture data, size distributions, and adult femal 

20 
LITERATURE CITED 

Bowen, B. W., A. B. Meylan and J. C. Avise. 1991. Evolutionary c 

distinctiveness of the endangered Kemp's ridley sea turtle. Nature 352:709-711. 

Brongersma, L. D. 1982. Marine turtles of the Eastern Atlantic 

Ocean, p. 407-416. In:Bjorndal, K. A. (Editor), Biology and Conservation 

Smithsonian Institution Press, Washington, D. C. 

Brunenmeister, S. 1984. Standardization of fishing effort and 

production models for brown, white and pink shrimp stocks fished in U.S. 

Gulf of Mexico, p. 187-2 10. In: Gulland, J. A. and B. J. Rothschild 

Shrimps - Their Biology and Management, Fishing News Books 

Surrey, England. 

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Table 1. Time series of nests arid hatchliigs numbers used in these analyses (1995 actual 

hatchlings were 120,03 8) 

Nests Hatchlings Released 

6000 29100 

5500 24100 

5200 15000 

4000 28400 

3000 3 1400 

2000 13100 

1800 14600 

1600 23 500 

1400 23500 

1200 11100 

1100 36100 

1050 30100 

924 48009 

954 63996 

868 37378 

897 53282 

750 48007 

746 32921 

798 58124 

702 51033 

744 488 18 

737 44634 

842 622 18 

878 66752 

992 74339 

1155 7623 8 

1275 921 16 

1184 84210 

1566 107687 

1930 126000 
Table 2. Sensitivity analysis of the 10 year to maturity age-based model for Kemp's ridley sea 

turtles. Benthic immature mortality rate was fixed for each run; adult mortality, pelagic immature 

mortality and the post-1990 decrease in mortality rates were chosen for a least-squares fit of 

predicted nests to observed nests for 1978 - 1995. Annual cohort size defined by actual number 

of hatchlings released from corrals 1966 - 1995. Reproduction parameters fixed for all runs: 

nestslfemale = 2.5, remigration interval = 2 years. Sex ratio = 1: 1. 

benthic immature pelagic immature adult(age lo+) post-1990 sum of 

(age 2 - 9) (age 0 - 1) multipler for squares 

benthic imm. 

and adult z 

2 S z S z S 

Table 3. Estimated year that the population will reach 10,000 nesting females per season, 

assuming that current conditions (survival and growth rates) remain constant, for models with and 

without density-dependent limitation in nest survival (see text for details), and the post-1990 

decrease in instantaneous mortality rates necessary to reach the recovery goal of 10,000 nesting 

females by 2020. 

Year to reach 10,000 nesters post- 1990 change in mortality 

needed to reach 10,000 nesters by 

the year 2020 

no densitydependence 

2014 

2014 

2014 

2013 

2013 

20 13 

with densitydependence 

2015 

20 15 

2015 

2015 

2014 

2014 

no densitydependence 

0.76 

0.79 

0.8 

0.81 

0.83 

0.84 

with densitydependence 

0.72 

0.74 

0.76 

0.78 

0.8 

0.81 
Table 4. Population size estimates from the sensitivity analysis. Assume 1 : 1 sex ratio. 

population sue in 1995 

benthic immature adults benthic immature pelagic immature 

z 

0.36 2,963 36,902 163,360 

0.4 2,968 47,603 169,889 

0.44 2,973 61,590 177,554 

0.48 2,977 79,936 186,553 

0.52 2,98 1 104,075 197,118 

0.56 2,984 135,927 209,520 

Table 5. 3 and 5 Year Running Averages and Interim Stranding Limit. 

Five Year Average Model 

Nests 

Multiplier Strandig 

Of InmwicMean + 1 Std Dev 

0.5 

1986-90 1.0779 0.0390 250.40 

1991 260.16 326.47 

1987-91 1.1 121 0.0560 

1988-92 1.1 167 0.0584 

1989-93 1.0886 0.0443 

1990-94 1.0978 0.0489 

1991-95 1.1309 0.0654 

lime Year Average Model 

Nests 

Multiplier 

OfIncre-ase 

Total ISL Observed 

Mean +l Std DevStrandiigs 

Strandip 

Mean + 1 StdDev 

Total ISL 

Mean 

Observed 

+1 Std DevS&andin@ 
Historical Levels of Kemp's ridley Nest 

Abundance, Rancho Nuevo, 1947 - 1995 

~~~g~Nmoci w gKPF830Sg 2222ZZE-22222222- 

Kemp's ridley Nests Counted, Total Hatchling 

Production, 1966 - 1995 

-+Hatchlings +Nests 

140000 6000 

120000 5000 

p 100000 Y 

4000 1 80000 

C 

60000 3000 

40000 2000 < 

20000 1000 g 

0 0 

WOOONQWWONQWWONt 

WWI-I-~I-~OWWWO~~~~ 

222222222z2z222 

Year 
Appendix 1 

post-1990 mortality multiplier 0.675 

1966-1990 pcst-1990 

~ 

I Age Proporti- Instantaneous Proportionate Instantaneous 

I Survival Mortalii Survival Mastality 

Model Fit 

Observed Predicted 

I Hatchli 

Benthic lmmahlta 

Adult 

Ycar nests nesls 

1966 6000 6000.00 

1967 5500 5045.10 

1968 5200 4242.17 

1969 4000 3567.03 

1970 3000 2999.33 

1971 2000 2521.99 

1972 1800 2120.61 

1973 1600 1783.12 

1974 1400 1499.33 

1975 1200 1260.71 

1976 1100 1183.34 

1977 1050 1097.10 

1978 924 986.03 

1979 954 949.41 

1980 868 931.32 

1981 897 838.59 

1982 750 766.97 

1983 746 744.45 

1984 798 725.52 

1985 702 657.07 

1986 744 705.42 

1987 737 720.65 

1988 842 809.33 

1989 878 951.61 

1990 992 958.49 

1991 1155 1111.42 

1992 1275 1255.97 

1993 1184 1327.41 

1994 1562 1606.09 

1995 1930 1856.77 

r 

0-1 0.442 0.817 0.442 0.817 

2-9 0.657 0.42 0.753 0.283 

10+ 0.841 0.173 0.89 0.1 17 

Residual 

Sum-of-Squares, 1978-1995 

Residual 

0.00 

454.90 

957.83 

432.97 

0.67 

-521.99 

-320.61 

-183.12 

-99.33 

-60.71 

-83.34 

-47.10 

62.03 

4.59 

-63.32 

58.41 

-16.97 

1.55 

72.48 

44.93 

38.58 

16.35 

32.67 

-73.61 

33.51 

43.58 

19.03 

-143.41 

-44.09 

73.23 
Appendix 1 

Figure Al.1 Observed and model predicted Kemp's ridley nests, 

1978-1996 

2000 -- -Observed nests 

- - Predicted nests 

4600 -- 

1000 

Figure A1.2 Residuals: Observed - Expected Nests, 1978-1995 

Figure A1.3 Model estimated adult and benthic immature abundance, 

1966-19915 

12000 60000 

- o- -total benthic juveniles 
Appendix 1 

Figure A1.4 Model estimated benthic immature deaths and observed 

strandings, 1986-1995 

Figure A1.6 Model projections -estimated number of nesting females with 

and without denslty-dependent hatchling production 

1996 2000 2006 2010 2016 2020 2026 

*Density-Independent hatchhg productbn - -A - -Density-dependent hatchling productbn 

Appendix 2 

1966-1990 post-1 990 

Age Proportionate Instantaneous Proportionate Instantaneous 

Survival Mortality Survival Mortality 

Hatchling 

Benthic Immature 

Adult 

0- 1 0.46 0.776 0.46 0.776 

2 - 9 0.657 0.42 0.746 0.293 

101- 0.826 0.192 0.875 0.1 34 

post-1990 mortality multiplier 0.697 

Model Fit 

Observed Predicted 

Year nest0 nests Residual 

1966 6000 7500.00 -1500.00 

1967 5500 6192.13 -692.13 

1968 5200 5 1 12.33 87.67 

1969 4000 4220.83 -220.83 

1970 3000 3484.79 -484.79 

1971 2000 2877.10 -877.10 

1972 1800 2375.39 -575.39 

1973 1600 1961.16 -361.16 

1974 1400 1619.17 -219.17 

1975 1200 1336.81 -136.81 

1976 1100 1237.39 -137.39 

1977 1050 1132.34 -82.34 

1978 924 1003.79 -79.79 

1979 954 959.23 -5.23 

1980 868 936.22 -68.22 

1981 897 833.14 63.86 

1982 750 754.93 -4.93 

1983 746 731.25 14.75 

1984 798 711.70 86.30 

1985 702 638.59 63.41 

1986 744 693.09 50.91 

1987 737 710.52 26.48 

1988 842 807.18 34.82 

1989 878 960.44 -82.44 

1990 992 964.69 27.3 1 

1991 1155 1122.13 32.87 

1992 1275 1266.38 8.62 

1993 1184 1329.69 -145.69 

1994 1562 1607.85 -45.85 

1995 1930 1850.01 79.99 

Residual 

S~m-ofiSquates, 1978-1995 69763.60 
Appendix 

Figure A2.1 Observed and model predicted Kemp's ridley nests, 

1978-1996 

-Observed nests 

- 0- Predicted nests 

Figure A2.2 Residuals: Observed - 

100.00 , 

Figure A2.3 Model estimated adult and benthic immature abundance, 

1966-1996 

total adults 12000 

- -0- .total benthic juveniles 

10000 

V) 

8000 

6000 

4000 

2000 

1966 1970 1974 1978 1982 1986 1990 1994 
Appendix 2 

Figure A2.4 Model estimated benthic immature deaths and observed 

strandings, 1986-1995 

7000 600 

Figure A2.6 Model estimated proportion of neophyte nesting females, 

1966-1996 

Figure A2.6 Model projections -estimated 

and without density-dependent hatchling production 

I -0-Density-Independent hatchling production - -A - -Density-dependent hatchling production I 
Female Standing Stock in 1995 

Age Number 

Appendix 2 
Appendix 3 

1966-1990 post-1990 

Age Propodonate Instantaneous Ro~onate hstaotaneous 

Survival Mortality Survival Mortalii 

Hatchling 

&nthicImmahve 

Adult 

0-1 0.416 0.876 0.416 0.876 

2-9 0.657 0.42 0.764 0.269 

10+ 0.861 0.15 0.909 0.096 

post-1990 mortality multiplier 0.64 

Model Fit 

Observed Predided 

Year nests nests Residual 

1966 6000 4500.00 1500.00 

1967 5500 3873.80 1626.20 

1968 5200 3334.74 1865.26 

1969 4000 2870.69 1129.3 1 

1970 3000 2471.22 528.78 

1971 2000 2127.34 -127.34 

1972 1800 1831.31 -31.31 

1973 1600 1576.47 23.53 

1974 1400 1357,lO 42.90 

1975 1200 1168.25 31.75 

1976 1100 11 15.22 -15.22 

19n loso 1050.74 -0.74 

1978 924 960.99 -36.99 

1979 954 934.16 19.84 

1980 868 922.36 -54.36 

1981 897 843.32 53.68 

1982 750 780.93 -30.93 

1983 746 760.7 1 -14.71 

1984 798 743.3 1 54.69 

1985 702 681.66 20.34 

1986 744 722.69 21.31 

1987 737 735.42 1.58 

1988 842 813.80 28.20 

1989 878 941.44 -63.44 

1990 992 951.13 40.87 

199 1 1155 1097.53 57.47 

1992 1275 1241.80 33.20 

1993 1184 1323.42 -139.42 

1994 1562 1603.26 -4 1.26 

1995 1930 1866.13 63.87 

SdFSguareg 1978-1995 48750.37 
Appendix 3 

Figure A3.1 Observed and model predicted Kemp's ridley nests, 

1978-1 996 

2000 -- -0bse~ed nests 

- 0- Predicted nests 

1600 -- 

1000 

I Figure A3.2 Residuals: Observed - Expected Nests. 1978-1995 

Figure A3.3 Model estimated adul and benthic immature abundance, 

1966-1995 

12000 80000 

10000 ~0000 gdr - -0- -totel benthic juveniles 

8000 40000 $ 

as 

J 6000 3' 

u 30000 3 

% 

4000 20000 z 

UL 

2000 10000 

0 0 

1966 1970 1974 1978 1982 1986 1990 1994 . 
Appendix 3 

Figure A3.4 Model estimated benthic immature deaths and observed 

strandings, 1986-1995 

7000 600 

Figure A3.6 Model estimated proportion of neophyte nesting females, 

1966-1996 

Figure A3.6 Model projections -estimated number of nesting females with 

and without density-dependent hatchling production 

70000 

60000 m 

U 5 60000 1 40000 

OI .E 30000 .. 2 20000 

10000 

0 

1995 2000 2006 2010 2016 2020 2026 

+Density-Independent hatchling productbn - -A- -Density-dependent hatchling production 
Appendix 

Stock in 1995 

Number 

63000.00 

22420.24 

7300.41 

6104.38 

3861.89 

2878.5 1 

1975.78 

1407.71 

663.53 

476.84 

327.52 

1165.38 
Appendix 4 

1966-1990 post-1990 

Age Proportionate lmtanlmeous Proportionate las?antanews 

Survival Mortalitv Survival Mortality r Hgtchling 0-1 0.714 0.337 0.714 0.337 

Benthic Immature 2-11 0.657 0.42 0.738 0.304 

Adult 12+ 0.847 0.166 0.887 0.12 

post-1990 mortality multiplier 0.723 

Model Fit 

Observed Predicted 

Year nests nests Residual 

1966 6000 6000.00 0.00 

1967 5500 5083.29 416.71 

1968 5200 4306.65 893.35 

1969 4000 3648.66 351.34 

1970 3000 3091.20 -91.20 

1971 2000 2618.91 -6 18.9 1 

1972 1800 2218.78 -418.78 

1973 1600 1879.79 -279.79 

1974 1400 1592.59 -192.59 

1975 1200 1349.26 -149.26 

1976 1100 1143.12 -43.12 

1977 1050 968.47 81.53 

1978 924 959.43 -35.43 

1979 954 927.9 1 26.09 

1980 868 857.75 10.25 

1981 897 862.29 34.71 

1982 750 880.46 -130.46 

1983 746 808.49 -62.49 

1984 798 754.67 43.33 

1985 702 751.56 -49.56 

1986 744 748.93 -4.93 

1987 737 687.50 49.50 

1988 842 754.82 87.18 

1989 878 783.20 94.80 

1990 992 892.75 99.25 

1991 1155 1135.01 19.99 

1992 1275 1231.84 43.16 

1993 1184 1453.04 -269.04 

1994 1562 1653.56 -91.56 

1995 1930 1747.59 182.41 

Sumsf-Scpms, 1978-1995 173719.43 
Appendix 4 

Figure A4.1 Observed and model predicted Kemp's ridley nests, 

1978-1 996 

1600 

2000 

0 t +Obsemd nests 

- 0- Predicted nests 

1000 

I Figure A4.2 Residuals: Observed - Expected Nests, 1978-1995 I 

Figure A4.3 Model estimated adult and benthic immature abundance, 

1966-1996 

12000 140000 

- - o - -total benthic juveniles 
Appendix 4 

Figure A4.4 Model estimated benthic immature deaths and observed 

strandings, 1986-1995 

21000 

18000 -- 

15000 -- 

-- 200 

3000 -- -- loo 

Figure A4.6 Model estimated proportion of neophyte nesting females, 

1966-1996 

Flgure A4.6 Model pro]ectlons -estimated number of nestlng females with 

and without density-dependent hatchling production 

...................... 

1996 2000 2006 2010 2016 2020 2026 

-.o-Density-hdepandent hatehllns productkn - -A- -0enrity-dependant hatehung produetkn 
Female Standing Stock in 1995 

Number 

63000.00 

38429.86 

21448.88 

17314.81 

10575.37 

7609.96 

5042.80 

3468.69 

1634.98 

1174.96 

807.03 

603.93 

224.75 

1173.32 

Appendix 4 
ASSESSMENT OF THE LOGGERaEAD SEA TURTLE POPULATION (Caretta caretta) 

IN THE WESTERN NORTH ATLANTIC 

A Report of the Turtle Expert Working Group 

July 1, 1996 

revised April, 1997 

STOCK DEF'INITION AND GEOGRAPHIC RANGE 

ggerhead sea turtle (Caretta carreta) occurs throughout the temperate and tropical regions 

e Atlantic, Pacific, and Indian Oceans @odd, 1988). Its range of habitat includes open 

waters, continental shelves, bays, lagoons, and estuaries. Loggerheads in the Western 

Atlantic nest on high-energy, sandy beaches between the latitudes of 18" and 35" North. 

wide, nesting on warm temperate beaches is much more common than nesting in the 

Loggerhead turtles are best known at their nesting beaches. Nesting females are iteroparous and 

show both philopatry between nesting seasons and site fixity between successive nests within a 

season (Carr, 1967; Dodd, 1988). Also, there is genetic evidence that loggerhead females return 

to nest on the same region of coast where they first entered the water as hatchlings (Bowen et al., 

1993). 

On average, loggerhead adult females in the southeast U.S. make reproductive migrations 

between foraging areas and nesting beaches every 2.5 yrs (Richardson and Richardson, 1982). 

On each migration, females deposit clutches in an average of 4.1 nests (Murphy and Hopkins, 

1984) with each nest averaging approximately 115 eggs @odd, 1988). 

In the western North Atlantic there are at least 4 loggerhead subpopulationsl separated at the 

eparations are based on analysis of mitochondria1 DNA which a turtle inherits fiom its 

. Based on mtDNA, Cheloniid sea turtle species worldwide have highly structured nesting 

lages (Bowen et al., 1992; Bass, 1994; Bowen et al., 1994; Broderick et al., 1994). 

mediated gene flow based upon mtDNA analysis among identified populations is low 

al., 1993; Bowen, 1995). Natal homing is a dominant force shaping this 

ographic structure. The nesting assemblages probably represent distinct genetic entities, 

search, including analysis of nuclear DNA (inherited from both parents) is necessary 

dress stock definition. Analysis of nuclear DNA of green turtles in both the Pacific and in 

ic found moderate rates of male-mediated gene flow (Karl et al., 1992), 

, populations within ocean basins were structured. Nevertheless, given the high site 

ty of nesting females to their natal beach and low gene flow between nesting assemblages, 
nesting beach (Bowen, B. personal communication to Sandy MacPherson, October 28, 1995; 

Bowen ed al., 1993; Bowen, 1995; Encalda et al., in press). Further investigation may delineate 

additional management units. Described by their geographic nesting distribution (Figure I), these 

subpopulations are: 

1. The Northern Nesting Subpopulation, occurring fiom North Carolina to northeast 

Florida, about 29" N. This subpopulation averaged about 6,200 nestdyear, 1989-1995, on 

surveyed beaches (N.C. Dep. Environ., Health, and Nat. Resourc., unpubl. data; S.C. Dep. 

Nat. Resourc., unpubl. data; Ga. Dep. Nat. Resourc., unpubl. data; Fla. Dep. Environ. 

Protect., unpubl. data). 

2. The South Florida Nesting Subpopulation, occurring from 29" N on the east coast to 

Sarasota on the west coast. This subpopulation averaged about 64,000 nestslyear, 1989- 

1995, on surveyed beaches (Fla. Dep. Environ. Protect., unpubl. data). This is the largest 

loggerhead nesting assemblage in the Atlantic and is the second largest in the world (Ross, 

3. The Florida Panhandle Nesting Subpopulation, occurring at Eglin Air Force Base and 

the beaches near Panama City. This subpopulation averaged about 450 nestslyear, 1989- 

1995 (Fla. Dep. Environ. Protect., unpubl. data). 

4. The Yucatan Nesting Subpopulation, occumng on the eastern Yucatan Peninsula, 

Mexico (Marquez, 1990). This subpopulation has about 1,500-2,300 nestslyear (Zurita et 

al., 1993; R. Mkquez, unpubl. data). 

Historically, nesting has been minor elsewhere in the western North Atlantic, except in Honduras, 

Nicaragua, Columbia, Venezuela, Bahamas, Cuba, Jamaica, and Puerto Rico (Sternberg, 198 1). 

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 loggerhead subpopulations 

of the northwest Atlantic account for 35-40% of the global nesting activity of this species (Ross, 

1982). 

uld an assemblage be extirpated, regional dispersal will not be sufficient to replenish the 

eted nesting assemblage within thousands of years. This has been demonstrated amply 

gh the loss of assemblages worldwide; they have not been re-established (e.g. Bermuda, 

an Island, Alto Velo). Thus, both genetic evidence (mtDNA) and the historical record 

e that recolonization of extirpated nesting assemblages does not occur on contemporary 

scales (Bowen et al., 1994). For conservation and management, mtDNA data indicate that 

g 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 
om loggerhead nests show a brief, 1-3 day period of frenzied activity (crawling then 

that gets them to the water and away fiom land. Following this frenzy, loggerheads 

at the surface and become associated with minor and major oceanic convergence 

boundaries (Carr, 1987) and are carried by ocean currents across ocean basins. 

s originating fiom U.S. Atlantic beaches are believed to lead a pelagic existence in 

e for an extended period of time, perhaps as long as 10-12 years, and are best 

fiom the eastern Atlantic near the Azores and Madeira (Maigret, 1983; Carr, 1986, 1987; 

, 1994; Bolten et al., 1994). We will refer to this life history stage as pelagic 

atures. The distribution of pelagic immatures originating fiom Gulf of Mexico beaches is 

wn. It is possible that they are carried by the Loop Current and Florida Current into the 

Atlantic Gyre or are kept within the Gulf by eddies (Collard and Ogren, 1990). 

roximately 40-60 cm straight carapace length (SCL), loggerheads which have inhabited the 

waters of the open ocean begin recruiting to coastal areas where they become benthic 

s (Carr, 1987). Developmental habitat for these loggerheads includes lagoons, estuaries, 

r mauths, and coastal waters typically less than 100 m deep. In U.S. waters small 

immature turtles (<70 cm SCL) are the predominant loggerhead size class found 

here from Cape Cod southward through Texas, except in southern Florida and the eastern 

f Mexico where large benthic immature loggerheads (70-91 cm SCL) are the most common 

(Medon~a and Ehrhart, 1982; Lutcavage and Musick, 1985; Burke et al., 1993; Teas, 

antum Resources, 1994; Epperly et al., 1995b,c; Schroeder et al., 1995) (Table 1, Fig. 

2). Benthic immature turtles occupy coastal feeding grounds for a decade or more before 

maturing and making their first reproductive migration (Carr, 1987). 

Genetic evidence has shown that assemblages of benthic immature loggerheads on foraging 

grounds comprise a mix of subpopulations. Loggerheads from Chesapeake Bay southward to 

Georgia are nearly equally divided in origin between the South Florida subpopulation and the 

Northern subpopulation. The representation of South Florida and Northern subpopulations is: 

Chorgia, 41% and 59% (Sears, 1994); South Carolina, 50% and 50% (Sears et al., 1995); 

Chesapeake Bay, 46% and 54% (Norrgard, 1995). In the Mediterranean, 57% of the immature 

loggerheads were from the South Florida subpopulation and 43% from the Mediterranean 

subpopulation (Laurent et al., 1993; Bowen, 1995). The genetic origins of benthic imrnatures in 

the Gulf of Mexico have not been determined. 

It loggerheads are known to make considerable migrations between foraging areas and 

beaches @odd, 1988). Based upon the returns of tags applied at nesting beaches, 

-nesting adult females from the South Florida subpopulation are distributed throughout the 

as, Greater Antilles, Yucath, eastern Gulf of Mexico, and southern Florida (Bell and 

978; Meylan, 1982; Meylan et al., 1983). Likewise, non-nesting adult females from 

subpopulation occur along the east coast of the U.S. and occasionally in the 

eastern Gulf of Mexico (Meylan, 1982). Limited tagging data suggest those adult females 

g in the Gulf of Mexico which are not part of the South Florida subpopulation remain in the 

of Mexico (Mey lan, 1 9 8 2). 
There is very little information regarding the distribution of adult male loggerheads. Most 

strandings of adult males occur early in the nesting season, suggesting that following mating males 

move away from the nesting beaches. Two exceptions are known. In South Florida, adult males 

are observed year-round (NMFS, unpubl. data) and males involved in mating behavior are 

observed throughout the nesting season in Cape Lookout Bight, N.C. (Cape Lookout National 

Seashore, unpubl. data). 

Both benthic immature and adult loggerheads in U.S. waters make seasonal migrations. Among 

loggerheads inhabiting the Atlantic coast, there is a spring migration of benthic immatures from 

south of Cape Hatteras northward and an autumn migration fiom the Middle Atlantic Bight 

southward (Shoop and Kenney, 1992; Epperly et al., 1995b,c). In this autumn migration some 

animals move south to Florida waters (Keinath, 1993; Epperly et al., 1995b,c) and others remain 

along the western wall of the Gulf Stream as far north as 35'47' north latitude (Ibid.). North of 

central Florida, there is also seasonal movement fiom offshore waters to nearshore and inshore 

waters in the spring and offshore movement during the fall and early winter (Epperly et al., 

1995b). Little is known about seasonal movement of loggerheads in the Gulf of Mexico. 

POPULATION SIZE 

Nesting data collected on index nesting beaches 1989-1995 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 the average number of nests per nesting female, 4.1) and of the number 

of adult females in the entire population (based on the average remigration interval, 2.5). The 

total number of nests laid along the U.S. Atlantic and Gulf coasts, 1989-1995, has ranged fiom 

53,016-85,306 annually, representing, on average, an adult female population of 43,060 (Table 2) 

[(nests/4.1) * 2.51. Generally, about 90.5% of the nests were fiom the South Florida 

subpopulation, 8.8% were from the Northern subpopulation, and 0.6% were from the Florida 

Panhandle subpopulation. 

Two methods were employed to estimate the current population size of all post-pelagic 

loggerheads in the coastal waters of the U.S. Atlantic and Gulf of Mexico coasts. These methods 

are described in detail in Appendices 1-2, along with assumptions and limitations of each method. 

Both employed ratio estimates using numbers of nests, strandings, and, in the second method, 

aerial survey data. A third method, a deterministic model, was deemed an unacceptable approach 

at this time due to the number of untenable assumptions and the lack of information on necessary 

input parameters (Appendix 4). The results of the two ratio methods are similar (Table 3); the 

mean post-pelagic loggerhead population size estimated for the 1989- 1995 period was 224,32 1 

and 234,355 turtles. These assuredly are underestimates for a number of reasons (see 

assumptions and implications in Appendices 1 and 2), especially because the methods only 

estimate the population in nearshore waters from which turtles are likely to strand. Based on 

aerial survey results (see Appendix 3), about 54% of the post-pelagic loggerhead population is in 
utheastern U. S. Atlantic, 29% in the northeastern U. S. Atlantic, 12% in the eastern Gulf of 

o, and 5% in the western Gulf of Mexico. 

ead was listed on July 28, 1978 as a threatened species under the Endangered Species 

(PL93-205). The species is considered to be "Endangered" by the TUCN (the World 

ation Union) and is listed in Appendix I of CITES (Convention on International Trade in 

ered Species of Flora and Fauna). 

stem North Atlantic, most loggerhead tagging, aerial surveys, and ground surveys of 

an in the 1970ts, prior to which there are no comparable data. However, commercial 

indicate that loggerheads were probably much more abundant historically than they 

itzell, 1994). In the first recovery plan for the loggerhead turtle (Hopkins and 

984), loggerhead abundance in the U.S. Atlantic and Gulf was estimated to be 

14,150 nesting females in 1983 (assuming an average of 4.1 nests per nesting 

y). If one assumes an average remigration interval of 2.5 years, an estimate of the 

number of mature females in 1983 is calculated to be 35,375. A more recent (1 989- 1995) 

of approximately 43,06 1 females is higher, in part, due to increased coverage of nesting 

some limitations, yearly numbers of nests remain the most widely available and reliable 

rs of stability in loggerhead subpopulations. Principal limitations are that abundance is 

to reproductive output and that only one sex of the oldest life-history stage in the 

ion is represented. Consequently, nesting numbers respond sluggishly to changes in the 

immature population and may be sensitive to changes in reproductive periodicity. 

ipal advantage of using nesting data to track trends in abundance is that the data have 

widely available in recent years. The most reliable nesting data are from complete 

s of nesting or from large unbiased suivey-samples gathered in a standardized fashion. 

like this came to cover a large majority of southeast U.S. nesting beaches by 1989. 

time series are short for beaches with the most complete nesting-survey coverage, there 

individual beach sites within subpopulation ranges that have been surveyed extensively 

periods of time. 

rends in the Northern Subpopulation 

tes that provide the longest time series for assessing nesting trends in the Northern 

lation are: Cape Island, S.C. (surveyed completely since 1973; U.S. Fish Wildl. Serv., 

. Klinek, pers. comrn.) and Little Cumberland Island, Ga. (surveyed completely since 

; J. Richardson, unplmbl. data). During the period 1989-1995, annual nesting at Cape Island 
was found to predict the total nests for the Northern Subpopulation well (?=0.90, b=0.95, SE of 

b=O. 14, P=0.00 1, n=7). However, a significant relationship could not be found between Little 

Cumberland Island nesting and nesting for the total subpopulation during the same period 

(w.30, b=0.55, SE of b=0.37, P=0.20, n=7). 

At Cape Island, annual numbers of nests from 1973-1995 ranged from 579-2,790 with a 

significant decline averaging 3.2% per year (?=0.47, b=-0.69, SE of k0.16, P=0.0003, n=23). 

Most of this decline appears to have occurred prior to 1979 after which a trend in nesting cannot 

be detected (?=-4.00, b=0.01, SE of b=0.26, P=0.97, n=17). Hopkins-Murphy and Murphy 

(1988) examined statewide nesting in South Carolina and found an average 5% per year decline in 

nesting between the period spanning 1980- 1982 and the period of 1985-1987 (three-year periods 

that are believed to encompass the majority of nesting in the subpopulation). Strandings declined 

at a similar rate during this period (Crowder et al., 1995), which may indicate a decline in the 

total number of benthic loggerheads. 

ough recent nesting at Little Cumberland Island did not mirror nesting of the Northern 

lation as a whole, trends in the long time-series of nesting data there may reflect 

nces that shorter time series do not. During the period 1964-1995, nesting at Little 

rland Island ranged fiom 34-278 with a significant decline averaging 2.6% per year 

7, b=-0.82, SE of b=0.10, P<0.0001, n=32). A significant decline, 2.5% per year, also is 

for the period 1964-1988 (r250.47, b=-0.69, SE of b=O.15, P=0.0002, n=25) but cannot be 

onstrated for the period 1989-1995 (?=0.21, b=-0.46, SE of b=0.40, P=0.30, n=7). 

Extensive standardized aerial, ground, and tagging surveys conducted throughout the Northern 

subpopulation's nesting range have begun only recently around 1989 (Meylan et al., 1995; N.C. 

Dep. Environ., Health, Nat. Resourc, unpubl. data; S.C. Dep. Nat. Resourc., unpubl. data; Ga. 

Dep. Nat. Resourc., unpubl. data; Fla. Dep. Environ. Protect., unpubl. data). During the period 

1989-1 995, annual numbers of nests ranged 4,3 72-7,899 with no detectable trend (?=0.017, 

b=O. 13, SE of b=0.44, P=0.78, n=7). 

Trends in the South Florida Subpopulation. 

In the South Florida Subpopulation range, there are eight beaches that have been surveyed 

completely since 1983 (Meylan et al., 1995; Fla. Dep. Environ. Protect., unpubl. data). One of 

these beaches, Hutchinson Island, was surveyed biannually, 197 1 - 198 1, and annually, 198 1- 1994 

(Quantum Resources, 1994). Annual nesting fiom a composite of the eight Florida beaches 

predicted annual nesting at all Florida index beaches well (?=0.96, b=0.98, SE of b=0.10, 

P<0.001, n=7) as did annual nesting only at Hutchinson Island (?=0.96, b=0.98, SE of b-0.10, 

P<0.001, n=7). This indicates that for the period prior to complete coverage of the nestirig range, 

both Hutchinson Island and the composite of beaches may accurately reflect nesting trends for the 

total South Florida Subpopulation. 
At Hutchinson Island during the period 1971-1994, nesting ranged from 2,757-7,126 with a 

significant trend of 4.0% per year (1-?=0.62, b=0.79, SE of b=0.15, P<0.0001, n=19). Trends for 

Hutchinson Island were also significant for the period 1971-1988 (2.2% per year, M.32, 

b4.56, SE of b4.25, P=0.04, n=13) and for the period 1983-1994 (4.5% per year, w.64, 

ba0.80, SE of ba0.19, P=0.002, n=12). The composite of eight beaches also showed a significant 

in~rease in nesting (5.7% per year; range, 14,627-26,776 nests per year) for the 1983-1994 period 

(i2t0.63, b4. 80, SE of b-0.19, P4.002, n=12). 

As in the Northern subpopulation range, extensive standardized nesting surveys in Florida have 

begun only recently. The most complete and unbiased indication of annual nesting numbers are 

for the period 1989-1995 and come fkom censuses on a set of index beaches that account for 

approximately 80% of the State's nesting (Schroeder, 1994; Fla. Dep. Environ. Protect., unpubl. 

data). The number of nests on index beaches for this period ranged 39,092-59,147 with no 

detectable trend (+0.33, b=0.58, SE of b=0.37, P=O. 18, n=7). 

One of very few unbiased indices of abundance for benthic immature and adult loggerheads can be 

found on the south-central Atlantic coast of Florida: the Florida Power and Light Company, St. 

Lucie Nuclear Power Plant. There, turtles are drawn into a cooling water intake that has operated 

almost continously since 1976 (Quantum Resources, 1994). Captures at the plant from 

1977- 1995 reveal cycles in abundance for all sue classes (Fig. 3) (Quantum Resources, unpubl. 

data). During this period, numbers of adult loggerheads 2 92 cm SCL increased significantly 

(&0.52, b=0.72, SE of b=O. 17, P=0.0005, n=19), but a trend cannot be demonstrated in numbers 

of small benthic immatures < 70 cm SCL (?=0.10, b=0.3 1 SE of b=0.23, P=0.20, n=19) nor in 

numbers of large benthic irnrnatures 70-91 cm SCL (+o. 18, b=0.43, SE of b=0.22, P=0.07, 

11-19). As in the NMFS stranding data for this region few adult loggerheads occur outside the 

nesting season. 

Trends in the Florida Panhandle and Yucatan Subpopulations 

Although nesting surveys in the Florida panhandle and in the Yucatan Peninsula have been too 

irregular to allow a meaningfbl analysis of trends, recent surveys do provide estimates of current 

nesting levels. Annual nesting in the Florida panhandle 1989 to 1995 ranged from 113-928 nests 

(Meylan et al., 1995; Fla. Dep. Environ. Protect., unpubl. data) and annual nesting in the states of 

Campeche, Yucath, and Quintana Roo, Mexico is estimated to be roughly 50, 100, and 1,331- 

2,166 nests, respectively (Zurita et al., 1993; Rene Marquez, unpubl. data). Nesting in Quintana 

Roo appears to be declining (R. Mbquez, unpubl. data). Elsewhere, Juventud Island, Cuba, is 

believed to have nesting on the order of several hundred nests per year (Rene Marquez, unpubl. 

data). 

pulation models developed for loggerheads in the Southeast (Crouse et al., 

Crowder et al., 1994) have used sensitivity analyses to evaluate the relative contribution of 

s and parameters (age to maturity, fecundity, stage-specific mortality rate, 

n growth. They found that the model population growth rate was 

ly insensitive to changes in fecundity and survival of the eggJhatchling stage. However, 

pulation was quite sensitive to changes in annual su~val ofthe 

larly 50 - 80 cm SCL benthic immatures. This is also the 

resented in U. S. loggerhead strandigs (Crouse et al., 1987; Teas, 

(1987) concluded that reduction in human-induced mortality of 

loggerheads was likely the most effective ways to effect recovery of 

on these and other findings, NMFS began requiring TEDs in some shrimp trawls in late 

Ds were required in most southeastern shrimp trawls with the 

in December 1994 (57 FR 57348, December 4, 1992). The use of 

reduced trawling-related mortality of loggerheads. Crowder et al. (1995) reported a 

overall reduction in loggerhead strandings in South Carolina coincident with the 

quirements. Recently, the National Marine Fisheries Service estimated 

ed on observer data (36 takes observed with 0.2% observer coverage over 4 years), the 

e of the U.S. shrimp fishery has been approximately 4,500 loggerhead sea turtles annually 

ne 1996 Biological Opinion). If one assumes all these takes were lethal and compares that 

to Henwood and Stuntz's estimate of mortality for the 1980ts, TEDs have reduced lethal 

about 54% (85-95% if Henwood and Stuntz's estimates were low by a factor of 3-4). 

a turtle nesting in South Carolina, which had been experiencing a decline of 5% annually 

the mid-1980's (Hopkins-Murphy and Murphy, 1988), appears to have stabilized 

ntation of TEDs (Hopkins-Murphy and Murphy, 1994; Hopkinsmortality sources that are being addressed also include: 

o The summer flounder bottom-trawl fishery off southern Virginia and North Carolina. 

Observer data indicated catch-per-unit effort rates (CPUE) equal to or exceeding those of 

the shrimp fishery (Epperly et al., 1995c), although the fishing effort was much lower, 

resulting in fewer overall takes. TEDs have been required in this fishery since 1992. 

Additionally, observers are currently placed on summer flounder trawlers as far north as 

New York to ascertain whether TEDs should be required elsewhere in this fishery. 

Dredging. The use of hopper dredges is seasonally restricted in southeast U.S. Atlantic 

waters to minimize potential impacts on sea turtles. All hopper dredges used in this area 

must have full-time observer coverage. 
o Oil-platform removals. The Minerals Management Service requires a precise set of 

procedures to ensure that sea turtles are not impacted by detonation of charges used in 

platform removal. 

o Beach nourishment. Seasonal restrictions, sand compatibility and even sand tilling are 

required for various beach nourishment projects throughout the southeast. 

o Nest protection: a whole host of measures, including restricting vehicular access, nest 

screening (against predators), nest relocation (erosion), regulating beachfront lighting 

(nesting female and hatchling disorientation) and others are pursued to various degrees at 

various beaches to increase nest success. 

o Longlime fisheries. The Crouse et al. (1987) loggerhead population model was sensitive to 

changes in mortality of the pelagic immature stage loggerheads. It is known that pelagic 

longline fisheries do take turtles, sometimes in large numbers (Aquilar et al., 1992; Brady, 

1994; Witzell, 1984). While protection of this stage is much more diicult, NMFS has 

begun consultation regarding U.S.-based longline fisheries in both Atlantic and Pacific 

waters. 

o Gill net fisheries. There are a number of more or less "local" gill net fisheries (e.g., pompano 

in east central Florida, black drum at the mouth of the Chesapeake Bay, inshore fl 

in North Carolina, etc.), ranging the length of the U.S. Atlantic and Gulf coasts w 

turtles to a larger or smaller degree. NMFS has been working with the states involved in 

an attempt to identifjl and address these gill net takes, but progress has been slow to date. 

o Power plant intakes. Several power plants have been found to entrain sea turtles through 

their cooling water intake tubes. Screens have been helpfbl at some; others monitor 

regularly to release entrained turtles. Most turtles are tagged and released unharmed. 

These locations have been usefbl to obtain live turtles for a number of studies. 

o Recently NMFS has recommended that the U.S. Coast Guard consult with NMFS regarding 

the potential impacts of generic permits for speed boat races, regattas, etc., and their 

potential impacts on endangered and threatened sea turtles. 

INTERIM STRANDING LIMITS 

While undertaking to reduce mortality from all sources, NMFS has been unable thus far to 

systematically assess cumulative human-induced mortalities and their effects on sea turtle species' 

continued existence and recovery. Understanding the population response to changes in various 

life history parameters is prerequisite to establishing Interim Strandings Limits for human-induced 
factors. Consequently, this was one of the primary tasks the Turtle Expert Working 

asked to undertake. 

ly, the primary index of turtle mortality comes f3om sea turtle strandiigs (turtles that wash 

nitored beaches and are reported to the Sea Turtle Stranding and Salvage Network) 

. The vast majority (90%) of these strandiigs are dead animals. Two shortcomings of 

ure of mortality are that: (1) it is most sensitive to nearshore mortality and completely 

ive to mortalities far offshore; and (2) the percentage of carcasses that wash ashore 

y is dependent on wind and ocean currents, etc. The majority of strandings show no 

cause of death, such as injuries from a boat propeller or shark attack, entanglement in 

et, emaciation suggesting illness, etc. What can be said, is that most stranded turtles 

althy at the time of death, many were actively feeding and have full stomachs; their 

and lack of obvious injuries is consistent 9th drowning in shrimp trawls although 

factors may contribute to some of these deaths, also. 

difficult to determine what proportion of at-sea mortality is reflected in strandings. Murphy 

opkins-Murphy (1 989) released marked carcasses offshore South Carolina, of which only 

ere later recorded as strandings. Epperly et al. (1996) reported that, at most, 7-13% of the 

attributable to the winter trawl fishery for summer flounder stranded. Another rough 

n can be made by noting that the National Academy of Sciences panel estimated that 

- 55,000 (most likely around 33,000-44,000) loggerheads and Kemp's ridleys drowned 

in shrimp trawls in the late 1980s prior to TED requirements (National Research 

, 1990). In the years 1986-1989 (when the stranding network monitoring effort was both 

consistent), total strandings for all species and all mortality factors reported by the 

network ranged fiom 1,991 - 2,373. This suggests that approximately 5-6% (range 3.6 

of the total mortality due to shrimp trawls was reflected in strandings over this period. 

er words, while strandings in the presence of shrimp trawling can be used as a rough index 

g mortality, they likely underestimate trawling mortality, severely at times. 

er of studies have shown that, at least at certain places and times (primarily Texas and 

Carolina), strandings, in number, have been significantly associated with offshore shrimp 

effort (Caillouet et al., 199 1 ; Caillouet et al., 1996; Crowder et al., 1995; Gallaway et 

, National Research Council, 1990). The National Research Council (1990) concluded 

in times of the year in South Carolina and Texas, 70 - 80% of strandiigs were due to 

ng mortality. Crowder et al. (1995) assessed a 14 year time period of one of the most 

and reliable strandings datasets available (South Carolina) and found a statistically 

t correlation between shrimp trawling activity and strandings. 

rtle Expert Working Group (TEWG) concludes that estimation of the maximum number of 

a1 loggerheads that can be taken incidentally to commercial fishing cannot be made at this 

The estimation of current total population size, mortality rates and population growth rates, 

d be unreliable at this time, due to deficiencies in data and model sructure. Furthermore, 

ause of the long pelagic stage and delayed age of maturity of loggerheads, using nesting trends 
to determine population status is much riskier than with a faster maturing species, such as Kempt 

ridleys. 

Despite the shortcoming of using strandiigs as an index of turtle mortality, the 

strandiungs remain the best data currently available for estimating nearshore 

the next year, the TEWG recommends use of a proxy Interim Strandings Lit (ISL) 

previous five-yr running average of strandings, plus one standard deviation to accoun 

variation (Table 6). While the TEWG recommended indexing a similar cal 

increase of nesting for Kemp's ridleyts, the group does not recommend such an 

loggerheads for several reasons: 

(1) The rate of increase of nests for the South Florida subpopulation, about 4%, is much 

lower than that for Kemp's. The rate is not statistically different from zero for the peri 

1989- 1994. 

(2) There has been no detectable increase in nesting in the Northern subpopulation. 

(3) Due to the slow maturation rate in loggerheads, the period over which pelagic and benthic 

itrunatures are exposed to mortality may be as long as 15 or more years, making it risky to 

use annual numbers of nests as indices of annual population size or changes in populatio 

size. For example, it is known that pelagic juvenile loggerheads are being taken, 

potentially in large numbers, in several high seas longline fisheries (Aquilar et al., 1992; 

Brady, 1994; Witzell, 1984). It is believed also that fishing effort in these fisheries has 

increased considerably since the United Nations high seas driftnet ban went into effect in 

December 1992. The impacts of such losses will not be reflected in nesting numbers fo 

many as 20yr or more. This latter reason also makes it imperative that a better method 

deriving and indexing an ISL be developed over the next year. 

Once an ISL has been calculated, the takes must then be allocated among potential "user" groups. 

This task is not in the charge of the TEWG, but the group has some recommendations with regard 

to this: 

(1) When an obvious natural mortality cause, such as cold-stunning, is identified, the Group 

recommends removing those individuals fiom the ISL tally. The ISL is derived in such a 

way as to identify potential strandings above natural mortalities, and if these were 

included, they would inappropriately increase the 5-year running average used for 

calculating &re years' ISLs; 

(2) For management purposes, it likely would be beneficial to calculate and monitor the IS 

on a regional and seasonal basis, allowing better assessment of whether current mortali 

is unusual and excessive. 
be seen (Table 6) that the ISLs for loggerheads calculated in this manner were exceeded in 

m Gulf of Mexico in 1994 and in the northeast U. S. Atlantic in 1992, 1993, and 1995, 

y equalled it in 1994 in this latter area. These excesses in mortality relative to historical 

along with the information that the combined subpopulation have shown no signs of 

overall, strongly suggests the need to reduce mortality for this species. 

rscores the temporary nature of this interim proxy (ISL): as noted above, 

underassessment of actual nearshore mortality and do not measure 

e mortality at all; the long time to maturity for loggerheads makes basing any take limits on 

numbers very risky. Furthermore, this interim proxy (ISL) may be to high, which could be 

al if used beyond 1996 (i.e., in the northeast, where ISLs have been exceeded in three of 

r years, those years of excessive strandings now form the basis for calculating the 

herefore, next year's ISL). Both additional information on juvenile population 

ce and population trends, and additional work on deriving a more suitable, less risky, ISL 

on are necessary. 

TEWG did examine the approach codified under the 1994 reauthorization of the Marine 

al Protection Act, where takes by fisheries and other human activities are now managed by 

lating a Potential Biological Removal (PBR) (Barlow et al., 1995). PBR is that number of 

als that can be removed from the population by human activities while still allowing the 

ion to recover, and move toward Optimum Sustainable population size. PBR is calculated 

e minimum population size and the intrinsic rate of growth of the population, with 

ative factors built into the calculation in several ways thus reducing the potential for 

mating allowable removals. The PBR calculation was validated by a series of simulations 

ssible marine mammal population trajectories, to determine if they would be adequately 

lating a minimum population size for the western North Atlantic loggerhead assemblages 

ed a number of assumptions, many of which were violated. Furthermore, the particular 

n for calculating PBRs was validated for populations with a marine mammal-like life 

ch might be inappropriate for marine turtles. The TEWG plans to pursue development 

approach to management of human mortality factors for endangered and threatened 

rtles over the next year by attempting to derive and validate an ISL calculation based on 

ar principles for sea turtles. 

CONCLUSIONS 

I. There is evidence supporting the separation of Western North Atlantic loggerhead turtles into 

at least four distinct subpopulations, three within the continental U.S. They cannot yet be 

managed separately because we do not know the in-water distribution of each subpopulation. 
h suggests that at least two of the subpopulations intermingle on the foraging grounds of 

Atlantic coast. 

n the recovery goal of more than 12,800 nests for the Northern Subpopulation (inferred 

e recovery plan) and given that currently nests number about 6,200 and have not increased 

eptibly, and that the number of nests declined during the 1980ts, it appears unlikely that the 

very goal for this subpopulation will be achieved. 

Given a recovery goal of measurable increases in the South Florida Subpopulation, and gi 

served increase throughout the 1980's at eight Florida nesting beaches, it appears that 

ulation is stable and may be increasing. However, the time frame for nesting surveys on 

eaches utilized by this subpopoulation is too small to evaluate trends throughout the 

n. Thus, recovery rates for the entire subpopulation cannot be determined with certainty at 

e. Further caution is warranted with this subpopulation because, although nesting activi 

een increasing, catches of benthic immature turtles at the St. Lucie Nuclear Power Plant 

e canal have not been increasing. 

4. Given that a recovery goal was not set for the Florida Panhandle Subpopulation and that it is 

extremely small, beaches within this region need to be established as index nesting beaches to 

assess trends in this subpopulation. 

5. Given that recovery goals were not set for the Yucatan Subpopulation nor for any other nesting 

assemblage in the Caribbean, steps need to- be taken to assess the status of these groups. The 

Yucatan subpopulation appears to be declining. 

oncludes that estimation of the maximum number of individual loggerheads that 

aken incidentally to commercial fishing operations cannot be made at this time. The 

was not able to model the population due to data and model structure deficiencies. The 

recommends use of a 1-yr proxy Interim Strandings Limit based on the previous five-yr 

of strandings, plus one standard deviation. During the interim, research to 

the recognized data and model structure deficiencies should be initiated and a turtlePBR approach should be examined to be used until sufficient data are available to model 

S needs to assess all human-induced mortality sources and reduce their impact. In 

NMFS should address problems in the northeast U.S. Atlantic, where ISLs have been 

ed 3 of the last 4 yr and were nearly equal during the fourth year. 

There are a number of Priority 1 recovery plan tasks that are NMFS' responsibility that have yet 

e klly implemented. Several would address research recommendations listed in the next 

ion. 
MMENDATIONS 

the following recommendations will produce data necessary to improve the estimates of 

d assessment models being generated and analyzed by the Turtle Expert Working 

uction of the level of uncertainty in these parameters will significantly increase the 

f models and the accuracy of model results. NMFS needs to facilitate the completion of 

recommendations. 

ch Recommendations 

erm, in-water indices of loggerhead abundance in coastal waters are needed to identifjl 

of sea turtles over time; and to detect changes in size composition with 

ding recruitment. Ongoing studies in St. Lucie and Indian River, Fla.and 

, N.C. should be continued. Additional studies should be started in other parts of 

ge, including one in each Sea Turtle Conservation Zone . All turtles caught should be 

with PIT tags as well as flipper tags, and should be scanned for existing PIT tags. CPUE 

e standardized by year to provide long-term comparability. The projects should 

stent and long-term, with data collected in a statistically defensible manner; this implicitly 

NMFS oversight and funding. Sea turtles encountered during these studies should be 

etic material to support mtDNA and nuclear DNA studies. Habitat use and 

dies conducted as adjuncts to abundance index studies should not compromise the 

effectiveness of the primary goal of indexing abundance. Projects which provide an assessment of 

trends in juvenile abundance should be given priority, to provide information on trends that will 

not be reflected immediately on nesting beaches. 

rmation is needed regarding the location and source of mortality of stranded sea turtle 

ses, as well as predictions regarding whether sources can be monitored by documenting 

ings. Trajectory modeling of oceanographic conditions (considering decomposition rates 

scavenging) should be developed to determine sea turtle carcass landfall patterns or to 

ifL conditions under which carcasses would not be likely to strand. 

Age and growth studies need to be completed. 

4. Physiological studies to determine the effects of forced and repeated submersion should be 

continued. 

netic studies of benthic immature and adult loggerheads stranded and collected in the 

ern North Atlantic should be conducted to determine natal origin. 

e utility of population models could be increased through improved age-class survival 

ates, improved information on the distribution of each subpopulation, and improved 

information on basic life history, such as (a) the number and distribution of number of nests laid 

6 1 
esting female during a season and the distribution of re-migration intervals, (b) age and 

h data to determine length in pelagia and age of maturity, etc., (c) in-water abundance 

ds, and (d) sex ratios. Data collection fiom nesting beach activities, stranding monitoring, and 

in-water abundance surveys should support this goal. 

Management Recommendations 

1. Stranding network index areas within the southeast region should be developed for lon8-term 

standardized coverage by the STSSN. The statistical utility and defensibility of the index areas 

should be primary considerations in the identification and monitoring of stranding index areas. 

dized PIT tag techniques and equipment needs to be developed. A centralized sea 

tag database should be developed. PIT tag scanners should be issued to representative 

SN participants in each state to maximize the mark-recapture components of long-term sea 

le studies, including nesting beach surveys. 

3. WS' efforts to obtain fbnding to expand the observer program should continue. NMFS 

should work with industry to assess sea turtle CPUE in various areas throughout the Western 

North Atlantic, including inshore embayments, nearshore waters and offshore waters, and in the 

Eastern North Atlantic. 

4. Attempts to reduce effort should continue. Limited time closures, gear restrictions, or other 

measures could be used effectively to reduce nearshore shrimping effort for increased sea turtle 

protection. Based on recommendations fiom these analysis, proposed rules implementing these 

recommendations should be promulgated in the event current protection measures are insufficient 

as evidenced by stranding levels exceeding the proposed ISLs, and strandings are determined to 

be caused primarily by shrimping activities. 

S should continue to evaluate the impacts of other sources of mortality, including gillnets, 

rawling, and bottom trawl fisheries, especially those that have high nearshore effort, and 

e all human-induced sources of mortality. 

A loggerhead recovery team should convene and consider new information, especially the 

etic data that indicate there are at least 4 subpopulations in the western North Atlantic. 

overy goals for each need to be defined. 

7. The TEWG or another similar group should meet annually to identifjl and review the status of 

this species to ensure recovery. 
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58. 

11, W.N. 1994. The U.S. commercial sea turtle landings. U.S. Dep. Commer. NOAA Tech 

Memo. NMFS-SEFSC-350, 130 p. 

, G.R., G.H. Balazs, and J.A. Wetherall. 1995. Growth in juvenile loggerhead sea turtles 

(Caretta cmetta) in the northern Pacific pelagic habitat. Copeia 1995:484-487. 

, G.R., A.H. Wynn, and C. Ruckdeschel. 1986. Age determination of loggerhead sea turtles, 

Caretta clmetta, by incremental growth marks in the skeleton. Smithson. Contrib. 2001. 

427,34 p. 

ta-G., J.C., R.. Herrera and B. Prezas. 1993. Tortugas Marinas del Caribe. 1n:Biodiversidad 

Marina y Costera de Mexico. (Eds.: S.I. Salazar and N.F. Gonzalez), CONABIO, 

CIQRO, lvIexico:73 5-75 1. 
le 1. Sizs frequency of post-pelagic loggerhead sea turtles stranding along the U.S. Atlantic and Ciulf 

of Mexico coasts, by zone and size class, 1986-1994. Note that an additional number of turtles 

stranded that either were not meamred or measurements were not stand=-. 

5llushm zsuxuhm 2i!ashn Bhl 

8 50 13 7 1 

1 1 

1 7 1 9 

17 165 108 290 

14 207 124 345 

4 11 5 20 

2 8 4 14 

20 48 17 85 

6 11 6 23 

9 17 10 36 

20 49 15 84 

5 3 8 

2 2 1 5 

5 2 7 

0 

0 

20 15 1 36 

139 54 7 200 

97 57 6 160 

220 124 13 357 

90 65 4 159 

6 28 7 41 

22 35 8 65 

36 75 57 168 

16 1 229 88 478 

234 235 133 602 

402 202 79 683 

977 245 82 1304 

462 187 115 764 

456 165 80 70 1 

215 113 62 390 

323 107 55 485 

256 97 44 397 

64 57 24 145 

30 2 2 34 

27 24 10 61 

51 25 15 9 1 

93 14 5 112 

75 2 77 

5 5 

0 

0 

7 1 

I 



ggerhaad strandings reported t o the Sea Turtle Stranding and Salvage Network (STSSN) by region, 1 
able 61 Strandings rtnd Incidental Strandings Limits (ISL), by year, by region. western Gulf of Mexico, eastern Gulf of 

Mexico, southeast U.S. Atlantic, and northeast U.S. Atlantic. Ihe ISL is the 5 year mean + 1 stadad deviation 

(SD). For e~~mple, the ISL for the western Gulf of Mexico in 1996 is 163 sbmded loggerhead turtles >40 cm that 

are not cold stun kills. This is calculated as the 5 year mean, ending with 1995. Asterisks indicate years when ISL 

was exceedd 

NMFS statistical zones used to report sea turtle strandngs. 
Figure 3. Captures of post-pelagic loggerhead sea turtles, 1977- 1995, at the Florida Power and 

Light Company, St. Lucie Nuclear Power Plant. Small benthic immatures are <70 cm, large 

benthic immatures are 70-91 cm, and adults are 292 cm (Quantum Resources, 1995, unpubl. 

ADULTS B44 LARGE IMMATURES SMALL IMMATURES 

ma 

leo $ la 

I 140 

a 

(3 120 sJ 1, 

LL 

0 80 

Ot 

UI 

I" 2 40 

20 

0 

YEAR 
APPENDIX 1 

or Nts = (Nf x St) I Sf 
=Nts + [(Nf x 2.5) - Nf] 

ion of loggerhead turtles in the benthic near 

assumptions and limitations. These are listed below: 

at it does not estimate the size oft 

but only the number of turtles in nearshore where 

ccurately represent th 

at the mortality causi 

on is estimated during the May-August ti 

rs can be calculat 

sive at this time of year, and has been consistent 

he ratio of adults of undetermined sex was 1 : 1 is valid o 

sex ratios. Analysis of the actual data suggested that 

t particularly biased in either direction. 

es that subadults and males remigrate annually. 

e number of nestslne 

s. The effect of underestimating the number of 

to overestimate the number of females in the popul 

the remigration interval would be to underestimate 

nesting data were available or included in the analysis for the 

Cuba, or other Caribbean nesting beaches. No data 

se areas and it is not known if turtles which nest in these areas 

8 1 
inhabit 'U.S. waters. The effect of excluding turtles &om these areas would be to 

underaimate the population. 

No estimates of variance were available for the population estimates produ 

method. These would be expected to be large given that variance would b 

with tht: numerous data inputs for the method including: counts of nests, nest 

per seawn, remigration interval, adult female percentage in the strandings. 
APPENDIX 2 

RATIO METHOD - SOUTHEAST U.S. ZONES WITH SPATIAL EXPANSION 

-pelagic loggerhead sea turtles ' 

of stranded adult females and 

males, and adults of unknown sex. This 

ons of the U.S. coastal waters based on 

during aerial surveys in the other regions compared 

f this method is to use the number of nesting females loc 

a during the nesting season (an accurate e 

er of mature females (r 92 cm) that str 

sting season to obtain a stranding rati 

e classes and both sexes on U.S. b 

r of turtles present during that 

bstantial nesting occurs (s.e. 

d the northeast U.S.), aerial 

ance throughout the range. Th 

total number of loggerheads in t 

ast, and Gulf of Mexico. Adult females not present duri 

the estimate for loggerheads availabl 

of the method are outlined below. 

ate the number of post-pelagic loggerheads in the U.S. Atlantic population: 

imate the Number of Loggerheads (Ntse) in southeast U.S. waters 

A. Compile NMFS nesting data, 1989-1994 by area: NMFS statistical zone 

FL) and zones 29-35 (n.e. FLY GA, SC, NC), corresponding to the South Flo 

Subpopula~tion, in part, and the Northern Nesting Subpopulation, respectively 

Divide the number of nests by 4.1 nestsfseason (from Murphy and Hopkins, 198 

determine the number of adult females in each zone available to strand during 

B. Compile southeast U.S. stranding data by area (zones 24-28 and zones 29- 

period May - August, 1989-94 by siie (C92 cm and, if r 92 cm, by sex) (App 

1). Adults (r 92 cm) of unknown sex were assumed to be 1 : 1 males and females. 
C. Solve for Ntsez, the total number of loggerheads available to strand by area (zone 

24-28 and zones 29-35), based on the proportion: 

Sf2 Nfi 

-- - - -- or NTsez = (Nfk * 

Stz Ntsez 

where Sfz is the number of stranded loggerhead adult females near 

is the totd number of stranded loggerheads near the nesting areas; Nfi is t 

adult females available to strand in the area during a single nesting season 

s have shown that juvenile turtles off Georgia and So 

Northern and South Florida nesting subpopulations, 

ratio method for Zones 29-35 was apportioned to the 

orida and 58% Northern (data from Sears, 1994). 

D. Add Ntsez from each of the two areas to obtain Ntse. 

e the number of loggerheads in the waters of the U.S. (Nts) 

A. Multiply Ntse by spatial expansion factors derived for aerial pelag 

Cape Hatteras to Key West (zones 24-35) in 1982 (Thompson, 1984); 

Pamlioo/&lbermarle Sounds in 1989-9 1 (Epperly et al. 1995a), Ca 

Scotia 1979-8 1 (Shoop and Kenney, 1992), Chesapeake Bay 1982-85 

Gulf of Mexico 1983-86 (Thompson et al., 199 1) to determine the 

U.S. coastal waters, but outside the southeast U.S. Atlantic (Nts) ( 

number of adult females not nesting during a given year (Nfi x 2.5 

obtain the number of post-pelagic loggerheads in U.S. coastal waters (Nt) 

contains the results. 

Nt =Nfix 2.5 

ations for the ratio method are listed in Appendix 1, and the 

for the aerial expansion estimates are listed in Appendix 3. 

and aerial data, all of those assumptions and limitations app 

eads sea turtles are distributed as reported from the variou 

84 
Appendix 2 

addition, because the ratio calculations are computed by area, the problem of sd 

s to determine the percentage of adult females in the strandiigs is greater than in the 

for. all zones combined. 
APPENDM 3 

SPATIAL EXPANSION FACTORS TO DETERMINE 

THE NUMBER OF LOGGERHEADS OUTSIDE 

SOUTEIEAST U.S. WATERS 

e the relative population sizes of post-pelagic 

. Only surface estimates of densitylnumbers 

the time period when turtles are maximall 

sting activity in the southeast U.S. were comp 

on SETS ziurveys (Thompson, 1984, Table lo), the estimated number of lo 

Cape Hatteras and Key West, Fla., was 1 

e is for offshore waters (seaward of CO 

waters. The largest inshore water body in 

-Albemarle Sound, N.C. The surface estimate for all hard-shell turtles i 

1989 was 3 13 turtles (95% C.I. 92-523) (Epperly et al., 1995% Table I; den 

Sa, Table 6) estimated that loggerheads account for an avera 

Thus, the estimate for loggerheads alone is 250 (72-418). 

s were flown at 152 m. The offshore surveys were fl 

with downward visibility, but the inshore surveys were flown from side-viewing 

he estimated number of loggerhead turtles on the surface in summers 

atteras and Nova Scotia, was 7,702 turtles (95% C.I.=2,220-11,000) 

Table 3). Again, this estimate is for offshore waters, only, and does 

nt in inshore waters. The largest inshore water body in the region is the 

e estimate for all hard-shelled turtles in the lower portion of this 

, was 372 (95% C.I.=133-611) (Byles, 1988, Table 3). 

live turtles in the Bay are loggerheads (Keinath et al., 1987, Table 1). 

is 3 12 (1 12-5 13). There is not an estimate of numbers 

but it represent a smaller area and turtle density in the upper B 

, about half (47%) of that of the lower Bay (Keinath et al., 

surveys were flown at an altitude of 229 m fiom a platform 

shore surveys were flown at 152 m from a platform with sideGoMe~c surveys in the Gulf of Mexico offshore and inshore waters and bays during the 

983-1!386, the mean number of loggerhead turtles sighted per linear km flown in the 

Gulf was 0.005 and the estimate for the northeast Gulf was 0.0085 (Thompson et al., 

3). ]Ewe assume the effective strip width = 0.556 km [l/f(O)=half-swath; we 

the KO) calculated fiom other Gulf surveys using the same platform, observers, altitude, 

ohoefener et al., .1990)], then the densities in the northwest and northeast Gulf, respectively 

86 
Appendix 3 

km. Given areas, respectively, of 144,056 sq. km . 2), the estimated number of turtles on the surfb 

. and 3,359 in n.e.). Note that these surveys were 

a platform with downward visibility. 

estimate crudely the relative proportion of the post-pelagic loggerhead 

t U.S. Atlantic waters during the summer. Based on the above 

ation sizes, by area, and given an absolute number for southeast U.S. 

pulation sizes for each area can be obtained by multiplying the southeast 

by these expansion factors: 

0.017 

0.516 

0.021 

0.3 12 

oughout U.S. Atlantic and Gulf waters 1979-1 989. If this was 

was declining, the effect would be to overestimate the proportion of 

own prior to the SETS surveys (the North Atlantic) and to 

on in the areas flown later (Chesapeake Bay, N.C., Gulf of Mexico). 

f turtles across these areas remained constant over the time period. 

relative distribution of loggerhead sea turtles is the strandings 

e must assume that turtle mortality rates are equal everywhere 

Mexico waters. The relative distribution of 

, 21.6%; southeast U.S. Atlantic, 59.1%; eastern 

o, 9.2%) is similar to the results of aerial surveys 

. Atlantic, 54.5%; eastern Gulf of Mexico, 12.0%; 

ighted, or that even if the number of small turtles are 

istributed similarly throughout the coastal waters. 

e maximally distributed during the summer and that their sufiacing behavior (e.g. 

on the surface) was the same in all areas during these months. 

rms and altitudes flown provide equivalent results. 

87 
methods of' determining density (e.g. strip vs line-transect methodology) yield the 

data are not available to partition out loggerheads fio 

ighted were loggerheads (leatherbacks were invariably s 

there are a multitude of assumptions inherent in aerial survey methodology. See 

1. (1995a) for a discussion of these assumptions. 
APPENDIX 4 

DETERMINISTIC MODEL 

le for loggerheads based on adult females 

f growth and survival 

e annual survival 

anagement. The TEWG suggested th 

assume a stable age distribution for that year, multiplyi 

989, adult annual survival and the conversion of nests 

rthern subpopulation was clearly decli 

y and Murphy, 1988), some Florida beaches (see 

nests over the same time period. Third, it was di 

e model outputs. The number of benthic i 

immature survival, is also high, suggesti 

recruit to the benthic feeding stage. Gi 

modeling exercises for loggerheads 

eter estimates and 
of Input Parameters for Deterministic Model 

of length of the pelagic stage is 10 years (Bolte 

juvenile studies indicate that turtles 

d 60 cm. Although they may begin recruiti 

1993; Bolten et al., 19 

e assumed that 50 cm turtles were approxim 

oster (1 994) predicted 5 1 cm turtles 

was determined by examining data fiom 

mg the average size of first-time nesters; this size was 

.54, N=106) (carapace curved lengths were converted 

of Teas (1993)). The average size of ne 

ion (North Carolina: Cape Lookout National S 

; South Carolina: South Island; Georgia: Wassaw 

SD=5.00, N=441). There are no estimates of th 

lorida subpopulation for comparison. The avera 

subpopulation (Merritt Island a 

.63, N=119). None of th 

was used as the best estimate 

interval of the mean calculated for 

nd 2) it does not to exceed the average size of nesters of the 

s, and reports were reviewed to determi 

loggerhead turtles (Frazer and Ehrh 

1995; Frazer, 1987; Medonqa, 198 

nwood, 1987). Estimates of average 

izes of several studies were small or i 

ased on growth rates of b 

ust require at least 10 years to complete, perhaps 20 years. 

age at sd maturity we used 20,25 and 30 years in the mo 
curves were fit to two data points corresponding to the be 

f the benthic immature 

maturity=20 yr, a=4.966 

and b=-12.4588. For an age of 

1s obtained from the growth curve analyses were used to convert the 

a from size distribution stranded to the age distribution stranded. The 

nding limb and a descending limb of a 

e descending limb were determined and a 

on age for the selected age interval yielded 2, an estimate of total 

.0202) over the period of 12-16 years of 

D=0.0307) over the period of 14-19 

.2006 (SD=0.0572) over the period of 

outh Florida, and Florida Panhandle 

age number of nests laidnesting female; 

mber of females nesting each year, 

o estimate the total number of adult 

chlings which enter the ocean. 

ode1 was ev'aluated with several primary sex ratios including 3 : 1, 2: 1, and 1 : 1 F:M. 
Table I 1. Post-pelagic loggerhead strandings, May-August 1989-1994, by zone groups, si 

liSex.