STATUS REVIEWS OF SEA TURTLES LISTED UNDER

THE ENDANGERED SPECIES ACT OF 1973:

Loggerhead Turtle, Caretta caretta

East Pacific Green Turtle, Chelonia mydas

Leatherback Turtle, Dermochelys coriacea

Hawksbill Turtle, Eretmochelys imbricata

Kemp's Ridley Turtle, Lepidochelys kempii

Olive Ridley Turtle, Lepidochelys olivacea

United States Department of Commerce

 National Oceanic and Atmospheric Administration 

 National Marine Fisheries Service

Silver Spring, Maryland

 

United States Department of the Interior

U.S. Fish and Wildlife Service

Washington D.C.
ii

Literature citations should read as follows:

Plotkin, P.T. (Editor). 1995. National Marine Fisheries Service and U. S.

Fish and Wildlife Service Status Reviews for Sea Turtles Listed under the

Endangered Species Act of 1973. National Marine Fisheries Service, Silver

Spring, Maryland

 

Pamela T. Plotkin, Ph.D., Editor, National Marine Fisheries Service, Office of Protected

Species, 1315 East-West Highway, Silver Spring, MD 20910

Current Address: Pamela T. Plotkin, Ph.D. University of Delaware, Department of

Entomology and Applied Ecology, 240 Townsend Hall, Newark, DE. 

 

Additional copies of this plan may be purchased from:

U.S. Fish and Wildlife Reference Service

5430 Grosvenor Lane

Suite 110

Bethesda, Maryland 20814

(301)492-6403 or

1-800-582-3421

The fee for the Status Review will vary depending on the current page charges.
iii

TABLE OF CONTENTS

Table of Contents ......................................... iii

Executive Summary ........................................ iv

Loggerhead Sea Turtle ....................................... 1

East Pacific Green Sea Turtle ................................. 24

Leatherback Sea Turtle ...................................... 37

Hawksbill .............................................. 76

Kemp's Ridley Sea Turtle ................................... 109

Olive Ridley Sea Turtle .................................... 123
iv

EXECUTIVE SUMMARY

Introduction.

Section 4 of the Endangered Species Act requires that National Marine Fisheries Service

(NMFS) and the U.S. Fish and Wildlife Service (USFWS) review the status of listed species

at least once every five years. The Services use these status reviews to determine whether a

designation as threatened or endangered accurately reflects the current status of a listed

species. If the status of the species has either improved or deteriorated, appropriate action

will be taken to ensure that the species is listed accurately. The status reviews for the

individual species of sea turtles were compiled by independent experts for NMFS and

USFWS from 1990 through May 1992 and were updated by Pamela Plotkin, Editor, in 1995. 

The status review for the green turtle (Chelonia mydas) is incomplete and will be published

separately. These status reviews represent the best biological information available to the

expert authors.

Summary of Findings:

Loggerhead turtle (Caretta caretta). It is recommended that the loggerhead turtle remain

listed as threatened throughout its range. Information indicates that its status has worsened in

recent years. The numbers of nesting females have declined in Georgia and South Carolina,

while Florida's nesting aggregations have remained intact. However, increased human

presence in areas of heavy nesting is likely to impact dense nesting populations in Florida in

the near future. Further study is needed on U.S. loggerhead population stock structure. 

Recent evidence suggests that the adult loggerheads nesting in Georgia may represent a

population that is distinct from those turtles nesting in Florida. If there are two separate

populations along the southeast Atlantic coast, then the northern one (i.e., those nesting in
v

Georgia, South Carolina, North Carolina, and northward) is in severe decline and should this

continue, consideration must be given to listing them as endangered.

Leatherback turtle, (Dermochelys coriacea). It is recommended that the leatherback turtle

remain listed as endangered throughout its range. While the leatherback turtle enjoys

complete protection in the United States, the species is often inadequately protected

elsewhere. Leatherback turtles foraging in the United States include juveniles which will

eventually recruit into the breeding aggregations of the western Atlantic and eastern Pacific,

where they will face considerable threats at the nesting grounds. In Mexico, the largest

breeding colony of leatherback turtles in the western hemisphere has suffered enormous

losses of both adults and eggs in recent decades. Similarly nesting leatherbacks were killed

and eggs collected in Malaysia and the British Virgin Islands where commercial and

subsistence take has extirpated once thriving populations. Threats to leatherbacks in the

marine environment have not been satisfactorily quantified but appear serious. Large

juveniles and breeding age adults are at increasing risk due to incidental capture and

drowning in fishing gear and entanglement and ingestion of persistent marine debris. 

Hawksbill turtle, (Eretmochelys imbricata). It is recommended that the hawksbill turtle

remain listed as endangered. Depleted U.S. populations are not currently considered to be

declining, but neither are there indications of recovery despite more than a decade of legal

protection. In addition, habitat destruction and clear evidence of over-exploitation continue

on a global scale. With the exception of the Kemp's ridley (Lepidochelys kempii), the

hawksbill is considered by many to be the most endangered of all the marine turtles. There

is little doubt that U.S. Caribbean and Pacific nesting colonies have been severely depleted

during the twentieth century. The illegal domestic harvest of eggs and turtles continues in

the United States, especially in Caribbean and Pacific island territories. 
vi

Kemp's ridley turtle, (Lepidochelys kempii). It is recommended that the Kemp's ridley

turtle remain listed as endangered. The Kemp's ridley is the most endangered of all sea

turtle species and shows little sign of recovery. A minor increase in nesting may be an

artifact of greater nesting beach coverage or may indicate that the nesting population has

increased. The lower average clutch size (95 eggs) that has been reported in recent years

may indicate an influx of novice nesters. Efforts to restore the species must concentrate upon

protecting subadult and adult animals and must take place over a long period of time because

of the species' slow maturation. The key to this task is the use of turtle excluder devices

throughout the U.S. and Mexican shrimp trawl fisheries in adult and developmental habitats,

and continued protection of the species' nesting beach.

Olive Ridley Sea Turtle, (Lepidochelys olivacea). It is recommended that the olive ridley

turtle in the eastern Pacific continue as threatened, the breeding population in Mexico

continue as endangered, and the western Atlantic population be reclassified to endangered.

The status of the olive ridley in the western Atlantic is unambiguous. Populations are very

low and continue to decline, almost certainly as a result of long-standing incidental capture in

shrimp trawls. This population is in worse condition than that of the endangered Kemp's

ridley and should be listed as endangered. Available data are too few to assess the status of

the species in the eastern Atlantic and northern Indian Oceans. However, probable

widespread egg collection throughout West Africa and known problems with incidental

capture and mortality in shrimp trawlers and gill nets near the arribada beaches of Orissa,

India seem to indicate that the populations there are stressed and should continue to be listed

as threatened. In the eastern Pacific, data indicate that some nesting aggregations are in

decline, while others appear relatively stable. There currently are no controls at all on

incidental capture by shrimp trawl and long-line fisheries throughout this vast region, which

appears significant throughout much of the region. 
1

Loggerhead Sea Turtle, Caretta caretta

Prepared by Nat B. Frazer, Ph.D.

Biological Background

Nesting adult female loggerheads in the southeastern U.S average about 92 cm in straight-line

carapace length, with a mean mass of about 115 kg (NRC 1990, NMFS 1990). Thus, they

tend to be larger than ridleys (Lepidochelys olivacea and L. kempii) and hawksbills

(Eretmochelys imbricata) but smaller than green turtles (Chelonia mydas) and leatherbacks

(Dermochelys coriacea). They have a carapace that is covered with horny scutes, as do the

other members of the family Cheloniidae. Loggerheads typically have five vertebral scutes,

with five costal scutes on either side of these, and eleven or twelve marginals lateral to the

costal scutes. However, there is much individual variation in the number of scutes, and it is

not unusual to find individuals displaying asymmetrical scute arrangements (e.g., eleven

marginals on one side and twelve on the other). The nuchal scute is in contact with the first

costal on either side, and the head is much larger in relation to the carapace than is the case

with other members of the Cheloniidae. The tendency of many loggerheads to have an

extensive community of epibionts covering the carapace (Caine 1986) often obscures their

carapacial markings. Ventrally, there are almost invariably three pairs of pore-less

inframarginal scutes on the "bridge" between the carapace and plastron. Adults and larger

juveniles have a orange or reddish-brown carapace and a yellowish or light brown plastron. 

Hatchlings range in color from light brown to almost black, with a three-keeled carapace

about 4.2 cm long and a somewhat lighter two-keeled plastron. The keels disappear as the

turtles increase in size (age). More complete descriptions and a review of taxonomic and

descriptive literature may be found in Dodd (1988).
2

Like other turtles in general (Wilbur and Morin 1988), and other sea turtles in particular

(NRC 1990), loggerheads are relatively long-lived and late-maturing animals (Dodd 1988). 

Growth rates of loggerheads reared in captivity for fourteen years in semi-natural conditions

indicate that maturity may be reached in about 16 years (Frazer and Schwartz 1984). 

However, additional captive studies indicate that growth rates are dependent upon

temperature, food quantity and food quality (Stickney et al. 1976, Nuitja and Uchida 1982),

casting some doubt as to the applicability of information obtained from studies of captive

loggerheads in providing accurate estimates of growth rates and age at maturity in natural

populations.

Based on studies of growth rates for wild loggerheads in Florida and Georgia waters, age at

maturity is estimated to be from 10-30 years (Frazer and Ehrhart 1985, Zug et al. 1986). 

The maximum lifespan for females in the wild would then be about 62 years, assuming that it

takes 30 years to mature, with a maximum of an additional 32 years of reproductive life-span

thereafter (Frazer 1983). lt is likely, however, that age at maturity and growth rates vary

widely among, and even within, local populations of loggerheads (Dodd 1988), as has been

demonstrated in long-term studies of freshwater turtles (Gibbons et al. 1981, Zweifel 1989,

Frazer et al. 1991). Recent estimates on age at maturity of Australian loggerheads suggests

that females mature between 34.3 and 37.4 years (Frazer et al. 1994).

Adult female loggerheads are iteroparous both within and among years, typically displaying

high fecundity. Most females usually nest from three to five times in a season, although

numerous individuals have been recorded laying six nests (Frazer and Richardson 1985a,

LeBuff 1990) and seven nests have been recorded for one individual (Lenarz et al. 1981). 

Average clutch size differs according to location, but usually falls between 95 and 150 eggs

per nest (Frazer and Richardson 1985b, Dodd 1988, LeBuff 1990). The estimated survival
3

rate from the time eggs are laid until the hatchlings reach adulthood is only about 0.1 - 0.2 %

(i.e., only one or two per thousand: Frazer 1986a).

Although loggerheads display iteroparity among years, most adult females do not return to

nest in successive years (Dodd 1988). The most frequently observed interseasonal nesting

interval for loggerheads in the U.S. is 2 years, although the intervals are not well understood

(Frazer 1989a) and individuals are known to increase or decrease their interseasonal nesting

intervals (Frazer 1984).

It is not unusual to observe large fluctuations in numbers of nesting loggerheads from year to

year in a given locality (NRC 1990), perhaps due to individuals shifting their interseasonal

intervals. This is similar to the behavior of freshwater turtle populations in which the

proportion of reproductively active females may fluctuate widely from year to year (Frazer et

al. 1990). The causes for such fluctuations are not understood for any turtle species, but are

presumed to be environmentally induced, perhaps involving the accumulation of resources

(e.g. nutrient or energy reserves) necessary for reproduction (Wilbur and Morin 1988).

Loggerheads are circumglobal in their distribution, inhabiting coastal, estuarine and

continental shelf waters in subtropical and tropical areas as well as in the temperate zones as

far north as Newfoundland and as far south as Argentina (NRC 1990, NMFS 1990). Their

major nesting areas are described as being "antitropical" (Pritchard 1979), indicating their

tendency to nest in subtropical areas immediately north of the Tropic of Capricorn or south

of the Tropic of Cancer. Some nesting occurs even further up into temperate areas such as

Japan and the Mediterranean between 30E and 40E N latitude (Dodd 1988). However, there

also is some scattered nesting of loggerheads in tropical areas such as the Caribbean, and

what is perhaps the world's largest nesting aggregation (ca. 30,000 nesting females per year)
4

is on Masirah Island, Oman (Dodd 1988). Additional important nesting areas (> 200

females per year) are found in Australia, Brazil, Colombia, Greece, Jamaica, Mexico,

Madagascar, Malagasy Republic, Mozambique, South Africa and the U.S. (Groombridge

1982, Mager 1985, Ehrhart 1989). 

Loggerheads occur in Atlantic, Gulf of Mexico, and Caribbean waters under the jurisdiction

of the United States (Ehrhart 1989). Nesting occurs from New Jersey (Brandner 1983)

around the Gulf of Mexico to Texas (Shaver 1988) and to Puerto Rico in the Caribbean

(Ehrhart 1989). The major U.S. nesting area is in eastern Florida, which may be the secondmost important worldwide nesting concentration after Oman (Dodd 1988, Ehrhart 1989,

NRC 1990, NMFS 1990). The best estimate of annual nesting on southeastern U.S. beaches

is just over 14,000 nesting females in an "average year" (Ehrhart 1989), with over 90 % of

the activity occurring on eastern Florida beaches. Within this subregion, the greatest density

of nesting activity occurs in south Brevard County. Over a seven year period, nest densities

averaged 441 clutches per kilometer per year (Redfoot and Ehrhart 1989). Much smaller but

important regular nesting aggregations occur in South Carolina, Georgia, and North Carolina

(NRC 1990, NMFS 1990).

Population Size

It is difficult to obtain accurate information on population size for any loggerhead population,

given the difficulties of finding and following the turtles at sea. This is especially true for

the younger age classes that may spend several years adrift in Sargassum in the north Atlantic

gyre (Carr 1986, 1987, Schwartz 1988). However, it is generally agreed that the most

suitable index to population stability results from monitoring the numbers of females nesting

at a given rookery from year to year (Ehrhart 1989, NRC 1990, NMFS 1990). The annual
5

fluctuations in the numbers (or proportions) of reproductively active female turtles (Frazer

1984, 1989, Frazer et al. 1990) make it difficult to assess trends in population size based on

numbers of nesting females for any species, and this is recognized even by those who suggest

that numbers of nesting females provide the best indicator available (Ehrhart 1989, NRC

1990, NMFS 1990). lt is widely recognized that such annual fluctuations may mask general

trends in population size for loggerheads unless studies are carried out over several decades

(Richardson 1982, Frazer 1983, NRC 1990). Nonetheless, most estimates of population

increase, stability, or decline, currently rely on monitoring numbers of nesting females. 

Detailed information on population sizes and trends not only is difficult to obtain, but often is

even more difficult to interpret once obtained. The problem of assessing worldwide

population levels is exacerbated by the fact that few countries support the level of research

necessary to monitor long-term fluctuations in sea turtle populations. It is generally believed

that there has been a historical decline in the numbers of loggerheads in several areas outside

the U.S. (Table 1). 

One area for which long-term data are available is on the Bundaberg coast in south

Queensland, Australia. The nesting beaches of the Bundaberg coast, as well as those of the

islands of the Capricorn Bunker Groups of the southern Great Barrier Reef, have shown a

decline in numbers of nesting loggerheads in recent years (Limpus and Reimer 1994). 

Numbers of nesting loggerheads on the Bundaberg coast have fallen from an average of well

over 400 during the late 1970's, to fewer than 300 each year of the late 1980's. ln fact,

there has been a steady monotonic decline from just under 300 in 1985 to well under 200 in

1989. This general decline in numbers of nesting loggerheads is corroborated by additional

studies at the turtles' feeding grounds (Limpus and Reimer 1994). 
6

Long-term studies of nesting females have recently provided information that was not

available at the time of the last status review of loggerheads (Mager 1985). Numbers of

nesting females in South Carolina declined by over 26 % in a five year period

(Hopkins-Murphy and Murphy 1988). Nesting activity on Little Cumberland Island, Georgia

during the late 1980's had decreased by 50 % from levels seen there throughout the

mid-1970's and was only 25 % of that observed there in the mid-1960's through early 1970's

(NRC 1990). Thus, long-term study has documented a decrease of more than 5 % a year in

South Carolina over a 5-year period, and of more than 3 % a year in Georgia over a 28-year

period. The rate of decline in numbers of nesting loggerhead females in these areas is similar

to the rate of decline of the most severely endangered sea turtles, the Kemp's ridley

(Lepidochelys kempii) in Mexico (Frazer 1986b).

As mentioned earlier, over 90 % of the loggerhead nesting activity in the U.S. occurs in

eastern Florida. No similar decline has been demonstrated for the Florida nesting

aggregations. However, the National Research Council's (1990) suggestion that the number

of loggerheads nesting on Hutchinson Island, Florida might be increasing may be overly

optimistic. Their results (NRC 1990) do not indicate any significant trend towards either an

increase or a decrease (e.g., p > 0.06). Martin et al. (1989), who study loggerheads in the

area, also reported no significant (p > 0.05) increase or decrease in numbers of females

nesting at Hutchinson Island over the 16-year period from 1973 to 1989. In addition, Ernest

et al. (1989) reported no significant increase or decrease in 13 years of monitoring captures

of loggerheads in the intake canal for the St. Lucie Power Plant on Hutchinson Island,

indicating that abundance of loggerheads in the waters off the island has not increased in the

past several years. Thus, despite suggestions to the contrary (NRC 1990), there appears to

be no indication of any statistically significant increase of loggerheads in the Hutchinson

Island area, and it seems best to assume (guardedly) that the numbers have been stable in
7

recent years rather than increasing. As Martin et al. (1989) suggested, the predicted

increased in human traffic on the beach, as well as attempts at erosion control, might

negatively impact nesting on Hutchinson Island in the near future. 

As mentioned above, south Brevard County supports the highest density of loggerhead nests

of any U.S. beach. The linear regression performed on numbers of nests deposited in south

Brevard County, Florida, over a seven year period (Table 2) also indicates no significant (p

> 0.30) increasing or decreasing trend in the numbers of nests recorded by Redfoot and

Ehrhart (1989) for the 21 km study area. Although nesting in this extremely important area

appears to be relatively stable over the past few years, those who study the loggerheads there

urge caution concerning the impact of the anticipated increase in the human population in the

area (Redfoot and Ehrhart 1989). They also call attention to the fact that raccoon

depredation of loggerhead nests almost tripled in that area between 1985 and 1988 (Redfoot

and Ehrhart 1989), indicating that although numbers of nests may be remaining fairly stable,

the numbers of hatchlings emerging from those nests may be declining.

Listing Factors

Numerous authors have called attention to the current and impending degradation of

loggerhead nesting beaches, particularly in Florida (e.g., Martin et al. 1989, Redfoot and

Ehrhart 1989). Such degradation may result from any number of causes, including

"nourishment" or augmentation of eroded beaches (Nelson and Dickerson 1989a, NRC 1990,

NMFS 1990), beach "armoring" or attempts at erosion control (NRC 1990, NMFS 1990),

increased human traffic on the beach (Mortimer 1989), attempts to clean the beaches of

debris with heavy equipment (NRC 1990, NMFS 1990), photopollution (Nelson and
8

Dickerson 1989b, Witherington 1986) or even the unintentional introduction of exotic

vegetation onto nesting beaches (LeBuff 1990, Schmelz and Mezich 1988). 

Quantitative estimates of human-induced mortality of nesting females, eggs and hatchlings

are not readily available. However, during 1988 alone, the Florida Department of Natural

Resources received reports of over 10,000 hatchlings that had been disoriented by artificial

lighting (NMFS 1988). With the anticipated continued human population growth and

development of beach front property in Atlantic and Gulf coast states in general, and in

Florida in particular, the degradation of nesting environments seems likely to increase in the

immediate future.

Efforts at providing nesting refuges in areas of heavy nesting (Jackson et al. 1988, Possardt

and Jackson 1989, Possardt 1991) will doubtlessly help to mitigate some of these problems. 

However, if the degradation of nesting habitat outside the refuge results in loggerheads

increasing the density of their nests inside the refuge, this could result in a degradation of

nesting habitat inside the refuge. It is anticipated that there would be an increase in the

number of nests destroyed as turtles nesting later may dig up nests that were laid earlier

(Bustard and Tognetti 1969). In addition, we are not able to predict the effect on the beach

ecosystem should loggerhead nesting decline dramatically in areas outside the refuge. As

Conway (1988:265) warned at the National Forum on Biodiversity "...sustaining species...in

small fragmented refuges provides little to the Earth in the way of basic ecological services"

except in those isolated areas.

Degradation of loggerhead habitat is not restricted to the terrestrial phase of their life cycle

on beaches. As Carr (1984:231) reminded us, "Turtle food comes mainly from the bottom in

the shallow fringes of the sea..." These near-shore and continental shelf waters, where adult
9

and large juvenile loggerheads spend much of their time, are also the areas on which human

activities potentially have their largest impact. Dredging, sea floor mining, oil spills,

industrial and agricultural runoff, and dumping of debris all can potentially degrade the

marine habitat occupied by loggerheads (Witzell 1989, NRC 1990). Several studies provide

quantitative information on loggerheads insofar as the ingestion of plastics or entanglement in

marine debris are concerned (e.g. Stanley et al. 1988, Plotkin and Amos 1988, Plotkin

1989), but quantification of most other aspects of the degradation of the marine habitat is not

yet available (Witzell 1989). Nevertheless, it is expected that increased human population

pressures will exacerbate the degradation of the near-shore and continental shelf areas of the

marine environment and thus negatively impact loggerhead foraging and breeding areas in the

immediate future.

There is presently no legal commercial or recreational take of loggerheads in waters under

U.S. jurisdiction. Although they may be subject to exploitation in other countries (see Table

3 in Mager 1985), most countries have at least some regulations affecting the take of sea

turtles (NRC 1990). The National Research Council's Committee on Sea Turtle

Conservation does not consider the illegal take of sea turtles or their eggs to be a significant

problem in the United States (NRC 1990).

Dodd (1988), provides extensive review lists of predators, commensals and parasites of

loggerheads. Stancyk (1982) earlier provided a review of much of the published information

on predators of sea turtles and their eggs. However, little quantitative data are available that

would give an accurate indication of the overall effect of disease or depredation on

population levels, although depredation is known to be especially heavy on hatchlings and

eggs. For example, up to 80% of loggerhead nests may be destroyed by the raccoon,

Procyon lotor, on some southeastern U.S. beaches (NRC 1990; NMFS 1990) in some years. 
10

Most turtle species display high mortality rates for eggs and young juveniles but low

mortality rates for adults and larger juveniles (see review by Iverson 1991a). Hence, heavy

depredation of eggs and hatchlings may be the norm for most turtle species, and the selective

pressure resulting from such heavy depredation of early life stages is thought to have been

responsible for the evolution of longevity and iteroparity in the Order Testudines (Wilbur and

Morin 1988, Congdon and Gibbons 1990). The pattern holds up for the all turtle species for

which complete life tables have been constructed (e.g., Tinkle et al. 1981, Crouse et al.

1987, Frazer et al. 1990, Iverson 1991b). Thus, it is difficult to ascertain whether occasional

years of heavy depredation of loggerhead eggs and hatchlings is a normal or abnormal

occurrence in the life of the species in any particular area. lt is clear, however, that

sustained levels of heavy predation on these early life stages can severely threaten loggerhead

populations if adults and larger juveniles are not experiencing their typically high natural

survival rates due to human-induced mortality in shrimp trawl nets (Crouse et al. 1987). 

Although Mager (1985) suggested in the last status review that existing regulatory

mechanisms were sufficient for the protection of loggerheads, the existing regulations have

not prevented the decline of nesting aggregations in some areas such as South Carolina and

Georgia (NRC 1990, Hopkins-Murphy and Murphy 1988, Frazer 1986b). Even where

numbers of nesting females are relatively constant, there is some question as to whether

existing regulations will prevent the continued degradation of beach and near-shore marine

habitats (Martin et al. 1989, Redfoot and Ehrhart 1989).

Recent demographic models of loggerhead population dynamics indicate that protection of

large juveniles and adults may be the most critical management need in the immediate future,

and it is these very life stages that are at greatest risk of capture in shrimp trawl nets (Crouse

et al. 1987, Frazer 1989b, NRC 1990). In areas where numbers of nesting females have
11

been declining, there is some evidence that the use of turtle excluder devices (TEDs) on

shrimp trawl nets has resulted in a decrease of stranded loggerheads. Crowder et al. (1994a)

reported an overall trend which showed loggerhead strandings declining at about 5 - 6 % per

year. However, given the resistance of the shrimping industry, confusion and equivocation

in terms of the TED regulations in some areas earlier (Rudloe and Rudloe 1989, Williams

1990), and recent increases in sea turtle strandings in the Gulf of Mexico (Shaver 1994),

presumably a result of incorrect installation of TEDs in some shrimp trawl nets, it remains to

be seen whether enforcement can ensure that TEDs will be deployed consistently and

effectively. Early estimates of annual survival rates for adult females (0.81) and larger

juveniles (0.70) were obtained for turtles in U.S. Atlantic waters before the introduction of

TEDs (Frazer 1983, 1987). There is a great need for similar estimates of annual survival

rates of large juvenile and adult loggerheads in these waters once an effective history has

been established over several years with TEDs in place, as well as for the Gulf of Mexico. A

recent model of loggerhead population dynamics (Crowder et al. 1994b) has shown that it

may take up to 70 years or more before the deployment of TEDs on shrimp trawls results in

any substantial observable increase in the numbers of nesting females.

Henwood and Stuntz (1987) estimated that between 9,000 and 10,000 loggerheads were

killed each year in U.S. Atlantic and Gulf of Mexico waters as a result of drowning in

shrimp trawl nets. The National Research Council (1990) provided convincing arguments

that these previous estimates might be low. Their own order of magnitude numerical

estimates of human-induced mortality of juvenile and adult sea turtles in U.S waters (Table

3) indicate that incidental capture in fishing gear is the most important known source of

human-induced mortality (NRC 1990), and may be as much as five times the level estimated

by Henwood and Stuntz (1987). Inundation of eggs by heavy rains or tides and erosion or

accretion of sand on nesting beaches are among the most important abiotic factors affecting
12

survival of eggs and hatchlings. However, none of these is considered to be an important

source of mortality threatening loggerhead populations (NRC 1990, NMFS 1990).

Conclusions

ln the last status review of loggerheads, Mager (1985) concluded that there was inadequate

information available to assess whether its status had changed since the initial listing as

threatened wherever it occurs (1978). Subsequent information presented above clearly

demonstrates that its status has changed for the worse in recent years. There can be little

doubt that numbers of nesting females have declined in Georgia and South Carolina (Frazer

1986b, Hopkins-Murphy and Murphy 1988, NRC 1990), while Florida's nesting

aggregations have held fairly constant (Table 2, Martin et al. 1989, NRC 1990). Increased

human presence in areas of heavy nesting is likely to impact dense nesting populations in

Florida in the near future (Redfoot and Ehrhart 1989, Martin et al. 1989), even if nesting

beach refuges are provided. There certainly can be no case for delisting loggerheads, given

the evidence presented herein. Therefore, recommendation is that loggerheads retain the

listing as threatened wherever they occur. Although the current recommendation is to retain

the "threatened" designation, further study is needed before the next status review concerning

U.S. loggerhead population stock structure. Stoneburner (1980) presented morphological

data suggesting possible differences in carapace shape between loggerheads nesting in

Georgia and those nesting in Florida. Recent biochemical evidence indicates that the adult

loggerheads nesting in Georgia may represent a population that is distinct from those turtles

nesting in Florida (Bowen et al. 1993), despite earlier findings to the contrary (Smith et al.

1977). The possibility that loggerheads along the southeast Atlantic coast represent at least

two distinct populations deserves close attention and further research. If there are, indeed,

two separate demes, then the northern one (i.e., those nesting in GA, SC, NC, and
13

northward) is in great difficulty. Its calculated rate of decline is as great or greater than that

of the Kemp's ridley (Frazer 1986b, Hopkins-Murphy and Murphy 1988). Although the

Florida nesting aggregation has been stabilized in numbers over the past several years, it is

threatened with additional detrimental effects because of an increase in the coastal human

population. Should additional research confirm the existence of the two demes, and should

the decline of the northern nesting population continue its decline, careful consideration must

then be given to listing them as endangered when their status is next reviewed.
14

Literature Cited

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Brandner, R.L. 1983. A sea turtle nesting at Island Beach State Park, Ocean County, New

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Bustard, H.R. and K.P. Tognetti. 1969. Green sea turtles: a discrete simulation of density

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Caine, E.A. 1986. Carapace epibionts of nesting loggerhead sea turtles: Atlantic coast of

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Carr, A.F., Jr. 1986. Rips, FADS, and little loggerheads. BioScience 36:92-100.

Carr, A.F., Jr. 1987. New perspectives on the pelagic stage of sea turtle development. 

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21

Table 1. Areas for which there has been a historical decline in numbers of loggerhead sea

turtles, Caretta caretta, or for which there is a suspected decline due to heavy disturbance or

exploitation (Demetropoulis and Hadjichristophoru 1989, Ehrhart 1989, Gil and Leon 1988,

Groombridge 1982, Groombridge and Whitmore 1989, Hewavisenthi 1990, Ross 1982,

Veniselos 1986).

Evidence of Historical Decline Suspected Decline

Bahamas Azores

Colombia Cape Verde Islands

Cuba Dominican Republic

Cyprus Guatemala

Greece Italy

Honduras Madagascar

Israel Madeira

Japan Mozambique

Mexico West Africa

Panama

San Andreas Archipelago

Sri Lanka
22

Table 2. Number of nests deposited annually in a 21 km study area of the beach in south

Brevard County, Florida by loggerhead sea turtles, Caretta caretta (Redfoot and Ehrhart

1989, Owen et al. 1991, Owen et al. 1993, Owen et al. 1994).

Year Number of Nests

1982 7,995

1983 9,423

1984 7,753

1985 10,240

1986 10,745

1987 9,780

1988 8,838

1989 9,381

1990 14,328

1991 13,194

1992 12,754

1993 10,591
23

Table 3. Estimated annual numbers of deaths and relative importance of sources of humaninduced mortality of juvenile and adult loggerhead sea turtles (Caretta caretta) in U.S.

waters. Relative Importance: 1 = high; 2 = moderate; 3 = low; 4 = unimportant. 

(Adapted from NRC 1990).

Cause Mortalities/year Relative Importance

Shrimp Trawling 5,000-50,000 1.0

Other Fisheries 500-5,000 2.5

Beach Development ? 3.0

Dredging 50-500 3.0

Entanglement ? 3.0

Oil Platform Removal 10-100 3.0

Collisions with Boats 50-500 3.0

Directed Take 5-50 3.0

Power Plant Entrainment 5-50 3.0

Recreational Fishing ? 3.0

Beach Vehicles ? 3.0

Beach Lighting ? 3.5

Beach Replenishment ? 4.0

Toxins ? ?

Ingestion of Plastics ? ?

or Other Debris
24

East Pacific Green Sea Turtle, Chelonia mydas

Prepared by Javier Alvarado and Alfredo Figueroa

Biological Background

Historically the black sea turtle has also been known as the East Pacific green turtle in

English, and as tortuga negra, tortuga prieta, caguama prieta, tora and sacasillo in Spanish

(Cliffton et al. 1982, Cornelius 1986). The East Pacific green turtle and the green turtle

(Chelonia mydas) share the following features that distinguish them from other sea turtles: 

four pairs of lateral scutes on the carapace; one pair of elongated prefrontal scales between

shell scutes that do not overlap; and the tomium of the lower jaw with a sharply serrated

cutting rim that corresponds with strong ridges on the inner surface of the upper tomium

(Carr 1952, Marquez 1990).

In 1868, Bocourt proposed the binomial form Chelonia agassizii based on the examination of

East Pacific green turtle specimens from the Pacific coast of Guatemala. The usage of this

binomial, however did not gain general acceptance and in the present century, the East

Pacific green turtle often has been referred to as Chelonia mydas agassizii even though there

has not been a detailed published argument to reduce it to subspecific rank. There has been

recent controversy over the taxonomic status of the East Pacific green turtle and whether or

not it is distinct from the Pacific green turtle Chelonia mydas. In distribution the East Pacific

green turtle is restricted to the east Pacific, where according to Pritchard (1983) there exists

occasional sympatry between black and green turtles without apparent intergradation. 

Morphological differences (Cornelius 1986, Groombridge and Luxmoore 1989, Figeroa and

Alvarado 1990) and osteological features (Kamezaki and Matusi 1995) show a distinction

between Chelonia agassizii and other Chelonia. The East Pacific green turtle is distinguished

from the green turtle mainly by size, coloration and carapace shape. The carapace of adult

East Pacific green is narrower, more strongly vaulted and more indented over the rear

flippers than in the green turtle (Cornelius 1986, Marquez 1990). In size and weight the East

Pacific green is conspicuously smaller and lighter than the green turtle. In the rookeries of

Mexico, the mean size for nesting East Pacific green is 77.0 cm in straight carapace length

(SCL) (Alvarado and Figueroa 1989a). On the Galapagos Islands, the mean SCL for nesting

East Pacific green is 75.0 cm (Marquez 1990). Adult female East Pacific greens weigh
25

between 65 - 125 kilograms (Cornelius 1986). On the other hand, the typical adult female

green turtle has an average shell length of 96.6 cm (SCL) and weighs between 65 - 160

kilograms (Groombridge 1982). East Pacific green males in the rookeries of Mexico are

smaller than females with an average SCL of 75.0 cm (Figueroa 1989). The mean cite for

East Pacific green hatchlings is 4.5 cm in SCL (Zamora 1990). Morphological data are

supported by nuclear DNA analysis which showed that Pacific Mexico and Galapagos Islands

Chelonia were distinct from other Chelonia populations (Karl et al. 1992). However,

mtDNA analyses do not show distinctness of the Eastern Pacific Chelonia from Chelonia

populations in other regions of the world (Bowen et al. 1992, Dutton et al. In press). 

From a dorsal view, true color of adult East Pacific green turtle is characteristically dark. 

The carapace and upper surfaces of the head and flippers is slate grey to black whereas the

plastron varies from bluish to dark grey (Caldwell 1962). Hatchlings are black to dark gray

above and white below with a white border around the dorsal edge of the carapace and

flippers (Cornelius 1986) while young individuals are usually brightly colored with a mottled

or radiating pattern of brown, olive or yellow (Caldwell 1962).

There is no sexual dimorphism in subadult male and female turtles, however, adult males

have longer tails that reach well beyond the rear edge of the shell, while adult females have

shorter tails that barely reach beyond the rear edge. Mature males have recurved long claws

for grasping the female and their plastrons become softened during breeding seasons. 

Carapace length at sexual maturity in females is about 55.0 cm (SCL) (Alvarado and Figuero

1989a). An estimate of age at sexual maturity for the Galapagos population is > 46 years

(MacFarland 1984).

Adult East Pacific green turtles are primarily herbivorous, eating sea grasses and algae

(Mortimer 1982) and in some areas they also feed on shellfish, jelIyfish and other marine

organisms (Brown and Brown 1982, Cornelius 1986). Predators of adult sea turtles are

primarily man and sharks (Stancyk 1982). Eggs are eaten by skunks, boars, domestic and

feral dogs, pigs, coatis, ghost crabs, ants, a beetle (Trox suberosus) and a scavenger fly

(Sarcophagidrle) (Green and Ortiz-Crespo 1982, Alvarado and Figueroa 1985, Marquez

1990). Hatchlings are eaten by coatis, shorebirds, lizards, ghost crabs, vultures, bonyfishes

and sharks (Marquez 199O).
26

East Pacific green turtles occur along the west coast of America from central Baja California

and the Gulf of California (Alvarado and Figueroa 1990a) to southern Peru (Marquez 199O). 

Small numbers or isolated individuals have been reported from California (Dutton and

McDonald 199O), British Colombia, Canada and Chile (Marquez 1990). The main nesting

sites are two neighboring beaches at Maruata Bay and Colola on the coast of Michoacan,

Mexico (Cliffton et al. 1982), and on the Galapagos Islands in Central America (Hurtado

1984). There are also less important nesting grounds in Mexico, namely the Pacific coast of

Guerrero, Jalisco, Oaxaca and Chiapas, and the islands of Clarion and Socorro (Marquez

199O) and along the Central American Pacific coastline (Cornelius 1982).

The nesting season varies with location. Nesting occurs in Michoacan between August and

January, with a peak in October-November (Alvarado and Figueroa 1985), between March

and July at the Socorro and Clarion Islands (Marquez 1990), and between December and

May with a peak in February on the Galapagos Islands (Green and Ortiz-Crespo 1982). 

Nesting cycles show differences between northern and southern populations. In Michoacan

females typically nest in two or three year cycles and deposit between one and seven clutches

per season at about 12-14 day intervals (Alvarado and Figueroa 1990a). In the South, on the

Galapagos Islands, females nest in two to five year cycles and deposit between one to five

clutches per season at about 14 day intervals (Marquez 1990). Average number of eggs per

clutch varies between 65 in Michoacan (Alvarado and Figueroa 1990a) to 80 in the

Galapagos Islands (MacFarland 1984). After 42 to 62 days of incubation (Marquez 1990)

hatchlings emerge mostly at night and travel quickly to the sea. After some time in a pelagic

existence, they enter shallow inshore and nearshore waters as juveniles.

Population Size

The status of sea turtle species is perhaps best indicated by long-term changes or trends in the

sizes of individual populations. Because females repeatedly return to the same beaches to

nest and because this is the time in their life cycle in which they are most available for direct

counting, counts of nesting females or nests provide the best available long-term index of the

status of populations.

The Michoacan nesting population consists of about one-third of the entire mainland-nesting

East Pacific green turtle population in the east Pacific (Cliffton 1982). Nesting females in
27

Michoacan have exhibited a clear decline in numbers over the last 40 years. In the early

1970's it was estimated that about 25,000 female turtles nested in Michoacan (Cliffton et al.

1982). Earlier observations tend to corroborate this figure; Peters (1962) reported tracks of

around 250 turtles on a 0.8 km of beach at Maruata Bay in August 1950, some two months

before peak nesting. Cliffton et al. (l982) take this to indicate about 900 turtles nesting in

Maruata Bay within several days of Peters' observations. This figure is 50 percent higher

than the number estimated nesting ln the bay for the entire 1979 season. By the middle

1970s nesting turtles were no longer common in August. In the past 10 years the estimated

nesting population in Michoacan has ranged from a high of 5,586 females in 1982 to a low of

940 in 1984 (Alvarado and Figueroa 1990a). Recent estimates show that the number of East

Pacific green turtles in the Michoacan nesting population is now extremely low. An

estimated 470 female turtles nested in the 1994-1995 season (Alvarado et al. 1995). In the

Galapagos Islands, East Pacific green turtles occur in moderate numbers. Between 1,200 -

3,250 females nest in the archipelago annually (Green 1994).

Although traditionally used as a food source by many coastal inhabitants, the primary cause

of decline in East Pacific green turtle numbers has been the systematic exploitation of eggs

and adults for marketing. Harvest of adults and eggs in Pacific Mexico increased steadily in

the present century, especially in the 1960's and 1970's (Alvarado and Figueroa 1990a). 

Although largely undocumented, incidental catch by shrimp trawlers is probably another

major mortality factor in Mexico (Groombridge 1982) as well as in Central America

(Cornelius 1982).

Listing Factors

1. The Present or Threatened Destruction, Modification or Curtailment of its Habitat or

Range.

Since 1980, the East Pacific turtle nesting beaches in Colola and Maruata,

Michoacan, Mexico, have been protected by armed patrols. Since 1986, they have

been declared Natural Reserves, and have been protected from development

(Alvarado and Figueroa 1989b). The Galapagos Islands are a National Park and all

nesting beaches are completely protected (Groombridge 1982). Accordingly the main

nesting beaches are not yet threatened by destruction or modification. The many
28

smaller nesting beaches in Mexico and Central America are largely unprotected from

habitat destruction.

In its oceanic environment East Pacific green turtle should be adversely affected by

habitat alterations and substances to the same degree as other sea turtles. The

following list includes some of the activities and substances that may adversely affect

or degrade the habitat of these turtles.

A. Pollutants from industrial and residential development. Discharges of

agricultural and industrial chemicals, petroleum products, and domestic

sewage may have indirect effects on herbivorous turtles by reducing food

sources through degradation of grass pastures, and direct effects such as a

possible reduction in health and fitness of individual animals that may be

manifested by mortality and the disruption of physiological functions

(Coston-Clements and Hoss 1983). However, levels of pollutant discharges

for much of the east Pacific are unknown and the direct effects of these

pollutants on East Pacific green turtles have not yet been investigated.

B. Disposal of garbage at sea. The growing presence of trash in the marine

habitat is an increasing threat to endangered populations of sea turtles and

other marine life. Marine debris poses a variety of threats to sea turtles from

ingestion to entanglement. Carr (1987) pointed out that sea turtles of all

species eat plastic scraps and other buoyant material and become entangled in

lines and netting discarded by fishermen. Plastic can become impacted in an

animal's digestive system, causing problems ranging from reduced nutrient

absorption, intestinal damage, or false feeling of satiation to blockages

resulting in death from starvation (Balazs 1985). Turtles entangled in lines or

netting may be subjected to injury, strangulation, starvation, and drowning

(Carr 1987).

Although traditionally used as a food source by many coastal inhabitants, the

primary cause of decline in East Pacific green turtle numbers has been the

systematic exploitation of eggs and adults for marketing. Harvest of adults

and eggs in Pacific Mexico increased steadily in the present century, especially

in the 1960's and 1970's (Alvarado and Figueroa 1990a). Although largely
29

undocumented, incidental catch by shrimp trawlers is probably another major

mortality factor in Mexico (Groombridge 1982) as well as in Central America

(Cornelius 1982).

C. Power boats. Power boats can injure or kill species of sea turtles

(Coston-Clements and Hoss 1983). During 1989 in San Diego Bay, several

East Pacific green turtles were found dead by injuries caused by power boats

(P. Dutton, pers. comm.).

2. Overutilization for Commercial, Recreational, Scientific and Educational Purposes.

The use of East Pacific green turtles for scientific and educational purposes, while

unquantified, is undoubtedly little and not a contributing factor in the decline of these

populations. In Mexico, the scientific take of East Pacific green turtles is controlled

by a permit program designed to protect the species.

Excessive exploitation is the overwhelming threat to East Pacific green turtle

populations throughout most of their distribution range, and has evidently caused

dramatic declines in numbers over the present century.

Prior to the 1950s the coastal breeding sites of the East Pacific green turtle in

Michoacan were relatively undisturbed as the coastline was virtually uninhabited and

of difficult access. During the 1950s coastal areas were increasingly cleared and the

coastal settlements of Maruata and Colola were established. Access to the area

remained difficult by land and, at that time, use of turtles by the local Nahuatl Indians

was at a low-subsistence level (Alvarado and Figueroa 1990b). During the late

1960s commercial markets for sea turtle products opened up in the area, involving

both leather and eggs. During the early 1970s about 70,000 eggs were collected each

night of the breeding season at Colola and a further 10,000 - 20,000 from Maruata

Bay (Cliffton et al. 1982). This total harvest of eggs continued until 1980 when

armed protection of Colola and Maruata was begun and hatcheries for the relocation

of nests were established (Alvarado and Figueroa 1985). 
30

Until the end of 1963 processing of turtle leather in Mexico was on a local, artisanal

level (Groombridge and Luxmoore 1989). During 1964 with the increase in

international demand for the product, larger scale, industrial processing began

(Alvarado and Figueroa 1989b). From the mid-1960s to mid-1970s East Pacific

green turtles in Michoacan were harvested mainly for their skins (Alvarado and

Figuero 1989b). From 1965 to 1977 about 165,000 turtles were harvested in the

Mexican Pacific (Groombridge and Luxmoore 1989). In 1973 a new coastal highway

reached the area and smugglers from the northern states of Sinaloa and Sonora gained

easy access to the Michoacan turtles and since then most poached turtles have been

smuggled to the north for meat consumption (Alvarado and Figueroa 1990a).

Cliffton et al. (1982) note that historically the East Pacific green turtle was abundant

within the Gulf of California and along the Pacific coast of Baja California. 

Large-scale commercial exploitation of turtles in the northern Mexican feeding

grounds began in the 19th century, considerably earlier than at the Michoacan nesting

grounds. At the turn of the century an estimated 1,000 live turtles a month were sent

to San Diego from Laguna Ojo de Liebre, Baja California (0'Donnell 1974). From

1956 to 1963 East Pacific green turtles harvested at the northern feeding grounds were

the most important component of the Mexican turtle fishery with a live weight

production of 3,430 metric tonnes (Groombridge and Luxmoore 1989). Cliffton et al.

(1982) noted that by the late 1970s, the remaining sea turtles had been virtually

extirpated from the Gulf of California.

In the early 1970s large numbers of overwintering East Pacific green turtles were

discovered off the south shore of Tiburon Island in the Gulf of California off the

Sonoran coast (Felger et al. 1976). Exploitation of these turtles began in 1975 and by

1980 they had been drastically reduced (Cliffton et al. 1982).

Overall it is certain that since the late 1950s many hundreds of thousands of East

Pacific green turtles, quite probably over one million, have been landed on the Pacific

coast of Mexico (Groombridge and Luxmoore 1989). This fishery has been by far the

most important factor in the collapse of East Pacific green turtle populations.
31

In Central America as well as in the feeding grounds in Colombia and Panama the

East Pacific green turtle is captured as bycatch of the olive ridley fishery (Marquez

1990). In 1982, Brown and Brown reported the existence of a East Pacific green

turtle fishery in Peru with an estimated annual catch of about 1,500 - 2,000

individuals, mainly subadults. The current situation of this fishery in Peru is

unknown.

3. Disease or Predation.

Cutaneous fibropapillomatosis has been recorded in several green turtle populations

(Balazs 1986). These tumors can indirectly cause mortality. Turtles whose vision is

blocked by tumors are unable to feed normally and turtles with fibropapillomatosis are

more prone to entanglement in monofilament line and other debris (Balazs 1986). 

There have not been any reports of this disease in East Pacific green turtles from

Mexico, Middle America or South America. However, McDonald and Dutton (1990)

found fibropapillomas in several juvenile East Pacific green turtles from San Diego

Bay, California.

4. Inadequacy of Existing Regulatory Mechanisms.

Taking of sea turtle eggs was prohibited by law in Mexico in 1966 (Marquez et al.

1982) and the capture of East Pacific green turtles was prohibited in 1984 (Alvarado

and Figueroa 1990b). However, laws are often ignored and poaching of both adults

and eggs has been common practice. The establishment of turtle camps at the East

Pacific green turtle nesting grounds in Michoacan since 1980 has significantly

diminished the poaching of eggs and breeding adults (Alvarado and Figueroa 1990a). 

In 1990 a presidential decree prohibited capture of all sea turtles and trade in sea

turtle products (Aridjis 1990). This should make clandestine trade in turtle products

much more difficult in the future.

In Central and South American countries where East Pacific green turtles breed or

forage, regulatory mechanisms exist for the protection of sea turtles. As in Mexico,

however, laws are often ignored and there is a brisk commerce of eggs in those

countries where nesting is common (Cornelius 1982). In the Galapagos Islands only
32

local inhabitants are legally allowed to fish for turtles and only on a subsistence basis

and egg poaching is practically nonexistent (Green and Ortiz-Crespo 1982).

5. Other Natural or Manmade Factors Affecting its Continued Existence.

The effects of natural phenomena on the continued existence of the East Pacific green

turtle are not known. However, natural forces that affect the production of eggs and

hatchlings, include El Nino, storms, rain and wave surge. These forces can produce

beach erosion or accretion, prevent turtles from nesting and destroy eggs and

hatchlings.

East Pacific green turtles are incidentally taken by shrimp trawlers as well as in

pelagic fisheries such as the longline fishery in the eastern Pacific Ocean. However,

the extent of this incidental take and the effect on the population remains unknown

(Cornelius 1982, Groombridge 1982). With the population as depauperate as it is,

any level of take of large juveniles or adults should be considered a problem for the

species.

Conclusion

Excessive exploitation throughout its distributional range has caused a dramatic decline in

numbers of the East Pacific green turtle in the present century. Efforts to recover the species

must concentrate on enforcing national and international regulatory mechanisms so that nests,

hatchlings, adults, subadults and juveniles are given effective protection in both breeding and

foraging habitats.

Utilization of adequate gear to reduce the number of East Pacific green turtles caught

incidentally in other fisheries, particularly the shrimp trawl fishery, should be required

throughout the East Pacific green turtle distributional range. Effective public education to

promote conservation of the East Pacific green turtle in coastal communities should be

undertaken.
33

The best available information indicates that the East Pacific green turtle is severely depleted

and in danger of extinction. Therefore, this species should be listed as an endangered species

throughout its range.
34

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Alvarado, J. and A. Figueroa. 1989a. Madurez sexual tardia de las tortugas marinas. 

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Michoacan, Mexico II. Cuadernos de Investigacion. U.M.S.N.H. No.7. 72 pp.

Alvarado, J. and A. Figueroa. 1990a. The ecological recovery of sea turtles of Michoacan,

Mexico. Special attention: the black turtle, Chelonia agassizii. Report to U.S. Fish

and Wildlife Service and World Wildlife Fund-U.S. August 1990. 88 pp.

Alvarado, J. and A. Figueroa. 1990b. Promoting community participation in the

conservation of Mexico's sea turtles. WWF Reports. June/July 1990. 

Alvarado, J., C. Delgado, A. Figeroa, F. Hernandez, T. Zubieta, I. Suazo. 1995. 

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Aridjis, H. 1990. Mexico proclaims total ban on harvest of turtles and eggs. Mar. Turt. 

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Balazs, G.H. 1985. Impacts of ocean debris on marine turtles: Entanglement and ingestion. 

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of Commerce, Honolulu, HI. pp 387-429. 

Balazs, G.H. 1986. Fibropapillomas in Hawaiian green turtles. Mar. Turt. Newsl.

39:1-33.

Bowen, B.W., A.B. Meylan, J.P. Ross, C.J. Limpus, G.H. Balazs, and J.C. Avise. 1992. 

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terms of matriarchal phylogeny. Evolution 46(4):865-881.

Brown, C.H. and W.M. Brown. 1982. Status of sea turtles in the Southeastern Pacific: 

emphasis on Peru. Pp.235-240. In K.A. Bjorndal (ed.). Conservation and Biology

of Sea Turtles. Smithsonian Institution Press, Washington, DC. 

Caldwell, D.K. 1962. Sea turtles in Baja Californian waters (with special reference to those

of the Gulf of California), and the description of a new subspecies of Northeastern

Pacific green turtle. Contrib. Sci. Los Angeles County Museum. 61:1

Carr, A.F. 1952. Handbook of turtles: the turtles of the United States, Canada and Baja

California. Comstock Publishing Associates, Ithaca, NY. 542 pp.

Carr, A.F. 1987. Impact on non-degradable marine debris on the ecology and survival

outlook of sea turtles. Mar. Poll. Bull. 18:352-356.
35

Cliffton, K. 1982. The ecological recovery of Chelonia agassizii, the East Pacific turtle of

Mexico. Report to World Wildlife Fund-U.S. 76 pp.

Cliffton, K., D.O. Cornejo, and R.S. Felger. 1982. Pp. 199-210. Sea turtles of the Pacific

coast of Mexico. In K.A. Bjorndal (ed.). Biology and Conservation of Sea Turtles. 

Smithsonian Institution Press, Washington, DC. 

Cornelius, S.E. 1982. Status of sea turtles along the Pacific coast of middle America. Pp.

211-219. In K.A. Bjorndal (ed.). Biology and Conservation of Sea Turtles. 

Smithsonian Institution Press, Washington, DC. 

Cornelius, S.E. 1986. The Sea turtles of Santa Rosa National Park. Fundacion de Parques

Nacionales, Costa Rica. 64 pp.

Dutton, P.H., S.K. Davis, T. Guerra, and D.W. Owens. In press. Molecular phylogeny 

for marine turtles based on sequences of the ND4-leucine tRNA and control regions

of mitochondrial DNA. Molecular Phylogenetics and Evolution.

Dutton, P. and D. McDonald. 1990. Status of sea turtles in San Diego Bay. Sea World

Research Institute. Technical Report No. 90-225. 17 pp.

Felger, R., K. Cliffton, and P. Regal. 1976. Winter dormancy in sea turtles: independent

discovery and exploitation by two cultures. Science 191:283-285.

Figueroa, A. 1989. Contribucion a la determinacion del status taxonomico de la tortuga

negra (Chelonia agassizii) Bocourt, 1868 de Michoacan, Mexico. U.M.S.N.H. 

Escuela de Biologia, (Tesis Prof.). 73 pp.

Green, D. and F. Ortiz-Crespo. 1982. Pp.221-233. Status of sea turtle populations in the

Central Eastern Pacific. In K.A. Bjorndal (ed.). Biology and Conservation of Sea

Turtles. Smithsonian Institution Press, Washington, DC. 

Green, D. 1994. Galapagos sea turtles:An overview. Pp. 65-68. In B.A. Schroeder and 

B.E. Witherington (compilers). Proceedings of the Thirteenth Annual Symposium on

Sea Turtle Biology and Conservation. NOAA Tech. Memo. NMFS-SEFSC-341. 

U.S. Dept. Commerce, Miami, FL. 278 pp. 

Groombridge, B. 1982. The IUCN Amphibia-Reptilia Red Data Book Part 1, Testudines, 

Crocodylia and Rhyncocephalia. IUCN, Gland, Switzerland. 426 pp.

Groombridge, B. and R. Luxmoore. 1989. The green turtle and hawksbill (Reptilia

Cheloniidae) world status, exploitation and trade. IUCN Conservation Monitoring

Center. Cambridge, U.K.

Hurtado, M. 1984. Registros de anidacion de la tortuga negra, Chelonia mydas en las Islas

Galapagos. Bol. Cientifico y Tecnico, Instituto Nasional de la Pesca. Guayaquil,

Ecuador. Vol. VI(3):77-106.

Karl, S.A., B.W. Bowen, and J.C. Avise. 1992. Global population structure and malemediated gene flow in the green turtle (Chelonia mydas): RFLP analyses of

anonymous nuclear DNA regions. Genetics 131:163-173.
36

McFarland, C. 1984. Population ecology of the East Pacific green turtle (Chelonia mydas

agassizii) in the Galapagos Islands. Nat. Geogr. Res. Rep. 16:463-476.

Marquez, R. 1990. Sea turtles of the world. An annotated and illustrated catalogue of sea

turtle species known to date. FAO Fisheries Synopsis. Vol.II(125). 8l pp.

Marquez, M.R., A. Villanueva 0., and M. Sanchez Perez. 1982. The population of Kemp's

ridley sea turtle in the Gulf of Mexico-Lepidochelys kempii. Pp. 159-164. In K.A.

Bjorndal (ed.). Biology and Conservation of Sea Turtles. Smithsonian Institution

Press, Washington, DC. 

McDonald, D. and P. Dutton. 1990. Fibropapillomas on sea turtles in San Diego Bay, 

California. Mar. Turt. Newsl. 51:9-10.

Mortimer, J.A. 1982. Feeding ecology of sea turtles. Pp. 103-109. In K.A. Bjorndal

(ed.). Biology and Conservation of Sea Turtles. Smithsonian Institution Press,

Washington DC.

 

O'Donell, J. 1974. Green turtle fishery in Baja California waters: history and prospect. 

M.S. Thesis. California State University, Northridge, CA.

Pritchard, P.C.H. 1983. Conserving sea turtles: review and comments. Copeia

4:1108-1111.

Stancyk, S. 1982. Non-human predators of sea turtles and their control. Pp.139-152. In

K.A. Bjorndal (ed.). Biology and Conservation of Sea Turtles. Smithsonian

Institution Press, Washington, DC. 

Zamora, R. 1990. Relacion de la humedad del ambiente incubatorio en el avivamiento de

Chelonia agassizii en la playa de Colola, Michoacan, Mexico. U.M.S.N.H. Escuela

de Biologia, (Tesis Prof.). 43 pp.
37

Leatherback Sea Turtle, Dermochelys coriacea

Prepared by Karen L. Eckert, Ph.D.

Biological Background

The leatherback is the largest living turtle and is so distinctive that it is placed in a separate

taxonomic family, the Dermochelyidae. Whereas other sea turtles have bony plates covered

with horny scutes on the carapace, the slightly flexible carapace of the leatherback is

distinguished by a rubber-like texture. The carapace is about four cm thick and is largely

constituted of tough, oil-saturated connective tissue raised into seven prominent ridges and

tapered to a blunt point posteriorly. A nearly continuous layer of small dermal bones lies just

below the leathery outer skin of the carapace. The narrow ribs lack pleural flanges and

remain widely separated throughout life. No sharp angle is formed between the carapace and

the plastron, resulting in the animal being somewhat barrel-shaped. Adult males are

distinguished by a long, thick tail that extends well beyond the posterior carapace margin.

The epidermis is black, with varying degrees of pale spotting, and is scaleless. The

underside is mottled pinkish-white and black; the proportion of light to dark pigment is

variable. The front flippers are proportionally longer than in other sea turtles and may span

270 cm in an adult. Hatchlings are likewise predominately black with mottled undersides,

but differ in being covered with tiny polygonal or bead-like scales. The flippers have white

margins, and rows of white scales appear as stripes along the length of the back. Front and

rear flippers lack claws. In both adults and hatchlings, the upper jaw bears two tooth-like

projections, each flanked by deep cusps, at the premaxillary-maxillary sutures. For a
38

detailed discussion of anatomy, including embryonic development, see Deraniyagala (1932,

1936), Dunlap (1955), Pritchard (1971), Pritchard and Trebbau (1984), and Renous et al.

(1989).

The population of leatherback turtles nesting at the Sandy Point National Wildlife Refuge

(NWR) on St. Croix, U. S. Virgin Islands (USVI), has been intensively studied since 1981

under the auspices of the USVI Division of Fish and Wildlife. A similar project was initiated

on Resaca and Brava beaches (Culebra Island, Puerto Rico) in 1984 by the U. S. Fish and

Wildlife Service. The average curved carapace length (CCL) of nesting females on these

beaches is 155 cm (range 137-176 cm) and weights of 262-506 kg have been recorded

(Eckert and Eckert 1983, 1985, Basford et al. 1986, 1988, Brandner et al. 1987, 1990,

Tucker 1988, Tallevast et al. 1990). Hatchlings average 61.3 mm (n=398) straightline

carapace length and 45.8 g (n = 282) in weight (Eckert et al. 1984). Similar data are

reported from Australia (Limpus and McLachlan 1979), French Guiana (Fretey and Girondot

1989), Costa Rica (Hirth and Ogren 1987), and Trinidad (Bacon 1970). Nesting females are

larger and smaller, respectively, in South Africa (averaging 164 cm CCL; Hughes et al.

1967) and Pacific Mexico (averaging 144 cm CCL; Sarti et al. 1987). The largest

leatherback on record is a male that stranded on the coast of Wales in 1988 and weighed 916

kg (Morgan 1989). 

Leatherbacks have the most extensive range of any reptile (71EN to 47ES; Pritchard and

Trebbau 1984) and exhibit broad thermal tolerances. The species is commonly reported in

New England waters and northward into eastern Canada, as well as occasionally sighted in

British Columbia and northward into Alaska. In Trinity Bay, Newfoundland, an adult was

observed by fishermen throughout an entire day swimming in open water amongst ice (Goff

and Lien 1988). The core body temperature for adults in cold water has been shown to be
39

several degrees C above ambient (Frair et al. 1972). This may be due to several features,

including the thermal inertia of a large body mass, an insulating layer of subepidermal fat,

counter-current heat exchangers in the flippers, and potentially heat-generating brown adipose

tissue (Mrosovsky and Pritchard 1971, Frair et al. 1972, Greer et al. 1973, Neill and Stevens

1974, Goff and Stenson 1988). The skeleton remains extensively cartilaginous, even in adult

animals, and the species is unique among turtles in showing an extensive cartilage canal

vascular system in the long bones (Rhodin et al. 1981).

With the exception of very preliminary data on growth in mature females (nesting females

returning to Sandy Point NWR gain, on average, about 0.5 cm CCL per year; R. Boulon,

unpubl. data), information concerning the growth rate of wild individuals is entirely lacking. 

Based on the unusual features of the leatherback skeletal system, as well as evidence that

some individuals have grown quite rapidly in captivity, Rhodin (1985) has speculated that

leatherbacks may reach sexual maturity in 2-3 years. Bels et al. (1988) challenge this

hypothesis in their report of a healthy captive leatherback 1200 days (3.28 yrs) old with a

carapace length of 82 cm and a weight of 28.5 kg, dimensions which are an order of

magnitude smaller than mature leatherbacks measured in the wild. Leatherbacks may indeed

reach sexual maturity at an earlier age than other sea turtles; however, information on the

growth of juveniles in the wild is necessary before growth rates can be accurately predicted.

Analysis of stomach contents has shown that leatherbacks feed primarily on cnidarians

(medusae, siphonophores) and tunicates (salps, pyrosomas) in temperate and boreal latitudes

(Bleakney 1965, Brongersma 1969, Den Hartog and Van Nierop 1984, Davenport and Balazs

1991). The turtles are sometimes observed in association with jellyfish, such as Stomolophus

(Leary 1957, Lohoefener et al. 1988), Aurelia or Chrysaora (Musick 1988, Collard 1990),

but actual feeding behavior only occasionally has been documented. Foraging has been
40

observed at the surface (e.g., on Aurelia, as reported by Eisenberg and Frazier 1983), but is

also likely to occur at depth. Based on offshore studies of diving by adult females nesting on

St. Croix, Eckert et al. (1989a) proposed that the observed inter-nesting dive behavior

reflected nocturnal feeding on vertically migrating zooplankton, chiefly siphonophore and

salp colonies. Den Hartog (1980) speculated that foraging may occur at depth after finding

nematocysts from deep water siphonophores in leatherback stomach samples.

The evidence currently available from tag returns and strandings in the western Atlantic

suggests that adults engage in routine migrations between temperate and tropical waters,

presumably to optimize both foraging and nesting opportunities. Leatherbacks tagged while

nesting in the Caribbean sometimes strand on New England shores (Pritchard 1973, 1976,

Lambie 1983, Boulon et al. 1988). Rhodin and Schoelkopf (1982) reported a gravid female

stranded in August on the shores of New Jersey, the eggs possibly residual from her last

clutch of the season in more southerly latitudes. Conversely, an adult female tagged in

Chesapeake Bay in May 1985 was killed in the Gulf of Guacanayabo, Cuba, in July 1986

(Keinath and Musick 1990). The species composition of epibiotic barnacle communities on

Caribbean-nesting leatherback provides indirect evidence that gravid females embark from

and subsequently return to temperate latitudes (Eckert and Eckert 1988). The longest

documented movement is that of an adult female who traveled 5,900 km to Ghana, West

Africa, after nesting in Surinam (Pritchard 1973). Recent studies by Morreale et al. (1993)

using satellite telemetry to monitor post-nesting movements of leatherbacks from the Pacific

and Caribbean coasts of Costa Rica have found that nesting cohorts travel along similar

migration routes and that these routes tend to be located along deepwater bathymetric

contours.
41

Long distance (> 100 km) intra-seasonal movement sometimes occurs among nesting

beaches; for example, between French Guiana and Surinam (Pritchard 1973) and between St.

Croix and Puerto Rico (Eckert et al. 1989b). Keinath and Musick (1993) used satellite

telemetry to monitor the inter-nesting movements of a female leatherback tagged while

nesting on St. Croix. After leaving St. Croix, this leatherback traveled a minimum distance

of 515.2 km and nested twice more on Vieques and Culebra Islands, Puerto Rico. 

An observation of courtship and/or mating in Puerto Rican waters was described by Carr and

Carr (1986); however, there is recent evidence that mating typically occurs prior to (or

during) migration to the nesting ground (Eckert and Eckert 1988). Nesting is generally

nocturnal. Preferred nesting beaches have deep and unobstructed offshore access (Hirth

1980, Mrosovsky 1983). When nesting occurs on high energy, spatially unpredictable

beaches, it is not uncommon for large numbers of eggs to be lost to erosion (Mrosovsky

1983, Eckert 1987). The most detailed reproductive data available for entire nesting colonies

have been collected at Sandy Point NWR, St. Croix, and on Culebra Island, Puerto Rico. In

these areas, nesting commences in March (rarely in February) and continues into July. 

Females renest on average every 9-10 days, deposit an average of 5-7 nests per annum

(observed range 1-11), and remigrate at 2-3+ year intervals. Annual project reports record

that clutch size averages 116 eggs, including 80 yolked eggs, on Sandy Point (1982-1989)

and 103 eggs, including 70 yolked eggs, on Culebra Island (1984-1989). Similar data have

recently been reported elsewhere in the Western Atlantic; for example, in French Guiana

(Fretey and Girondot 1989) and Costa Rica (Hirth and Ogren 1987). Data from Culebra

suggest that remigrants (returning turtles tagged during a previous nesting year) lay more

clutches per annum than do turtles recorded nesting at this rookery for the first time (Tucker

and Frazer 1991).
42

The temperature in the nest during the 55-75 days of incubation influences the sex of

hatchlings. The "pivotal temperature" (= 1:1 sex ratio) for leatherback eggs is estimated to

be 29.25EC-29.50EC in Surinam and French Guiana (Mrosovsky et al. 1984, Rimblot-Baly et

al. 1986-1987). Pivotal temperatures have not been defined elsewhere for this species. 

Several authors have cautioned against artificial incubation techniques which potentially bias

sex ratios (e.g., Mrosovsky and Yntema 1980, Morreale et al. 1982, Dutton et al. 1985). 

Under natural circumstances hatchlings emerge from the nest in the early evening and, to the

best of our knowledge, orient toward the ocean using light; that is, using the brightness of

the seaward horizon as their primary cue. Once in the water there is some evidence that

hatchlings maintain their orientation by swimming at a fixed angle relative to waves

(Lohmann et al. 1990). Locomotion is described by Davenport (1987). Nothing is known of

the behavior or survivorship of post-hatchlings. In his review of the pelagic stage of

post-hatchling sea turtle development, Carr (1987) found no evidence that young leatherbacks

associate with Sargassum or epipelagic debris. On 19 May 1989, an injured juvenile (29 cm

CCL) was found in the surf off Playa Humacao, Puerto Rico. It was transferred to Coral

World, St. Thomas, for rehabilitation and subsequently died there (Johnson 1989). Beyond a

few such chance encounters, there are no data on immature leatherbacks under U. S.

jurisdiction.

Population Size

Nesting grounds are distributed circumglobally (ca. 40EN to 35ES), with the Pacific coast of

Mexico supporting the world's largest known colony of nesting leatherbacks. The only

available estimates of the world population of leatherback turtles are confined to breeding

females. From nest counts and a knowledge of intra- and inter-annual nesting periodicities,

the number of breeding females can be approximated. Pritchard (1982) estimated the number
43

of adult female leatherbacks worldwide to be 115,000, with about half of them nesting along

the Pacific coast of Mexico. Recent declines in the number of leatherbacks nesting in

Mexico and elsewhere have occurred and Spotila et al. (In press) estimate that only 20,000 to

30,000 female leatherbacks exist worldwide. During the 1986-1987 breeding season, an

estimated 4796 nests were laid on 4.5 km of beach at Mexiquillo, Michoacan, Mexico (Sarti

et al. 1989). Nesting at this site has been steadily dropping, however. An estimated 1200

nests were laid in 1990-1991, and a mere 70 nests were laid in 1993-1994 (Laura Sarti M.,

UNAM, unpubl. data). It is not uncertain whether the decline reflects a natural fluctuation

(perhaps related to El Nino), or is a warning that the population is in serious jeopardy. The

other large colony in the western hemisphere is Yalimapo-Les Hattes, French Guiana, where

the total number of adult females is estimated at 14,700-15,300 (Fretey and Girondot 1989). 

As erosion has degraded nesting beaches in French Guiana, the large colony there has spilled

over into Surinam. There were fewer than 100 leatherback nests laid in Surinam in 1967,

but annual numbers have risen (amid natural year-to-year fluctuation) to 5565 nests in 1977

and 9816 nests in 1987 (Reichart and Fretey 1993).

Until recently, large colonies also flourished in the eastern hemisphere. At least 13,000

leatherback nests were reported in 1984 on 17.8 km of coast extending eastward from

Tanjung Jamursba on the Bird's Head Peninsula, Irian Jaya, making this Indonesian site one

of the most important in the world (Bhaskar 1985). In 1991, Betz and Welch (1992) visited

the Jamursba-Medi rookery and reported that nesting levels had declined to a mere 25% of

those reported by Bhaskar (1985); the near total collection of eggs has most certainly

contributed to the population's demise. Similarly, the beach at Rantau Abang (Terengganu,

Malaysia), once considered a major nesting area, has experienced a dramatic decline as a

result of intensive egg collection. The data show a steady drop in nesting activity from 6721

landings in 1968 to 372 in 1988 (Chua 1988) to 207 in 1991 (Sukarno Wagiman, Marine
44

Fisheries Resource Research Center, Terengganu, in litt., December 1991). In addition,

about 400 leatherbacks are caught annually off the coast of Terengganu (Brahim bin Salleh et

al. 1987); trawl nets are responsible for some 60% of the total number of turtles caught, with

the balance attributed to drift nets (Chan et al 1988). Population declines also have been

reported in India, Sri Lanka, and Thailand (see Ross 1982).

Pritchard (1971), Bjorndal (1982), Groombridge (1982), Carr et al. (1982), Bacon et al.

(1984a), and Pritchard and Trebbau (1984) offer the most complete discussions of the present

distribution of leatherback nesting grounds. Based on summaries presented by these authors,

nesting in the western Atlantic occurs in Brazil, the Guianas (French Guiana, Surinam,

Guyana), Venezuela, Central America (notably Panama and Costa Rica), and Mexico. In the

West Indies, Trinidad and the Dominican Republic are relatively important (supporting

perhaps 200-300 leatherbacks per annum), while nesting elsewhere in the eastern Caribbean

is predictable but occurs nowhere in large numbers. There is considerable anecdotal

information to suggest that nesting has declined at many Caribbean rookeries over the last

several decades. However, with few exceptions (see Cambers and Lima 1990, Eckert 1992)

documentation is lacking. Information concerning the historical status of western Atlantic

populations is currently being compiled by the Wider Caribbean Sea Turtle Recovery Team

and Conservation Network (WIDECAST), as part of an effort to draft sea turtle recovery

action plans for each of 39 Caribbean governments.

In the eastern Atlantic, direct and/or indirect evidence of nesting is available from Angola,

Sénégal, Liberia, Côte d'Ivoire, Ghana, Gabon, Togo, and Zaire. Fretey and Girardin

(1988) report nesting between Pongara Point and Ndindi in Gabon, near the Congo border,

and suggest a population size of 1276-2553 breeding females. In the Mediterranean, rare

nestings are reported from Israel and the southern coast of Sicily. In the Indian Ocean, the
45

best known nesting grounds are in Sri Lanka and in Tongaland, South Africa. Small-scale

nesting occurs in India, and larger aggregations are known in the Union territories of the

Andaman and Nicobar Islands. In Indonesia, Sumatra and Java reportedly receive some

nesting; Irian Jaya is still important. Thailand (especially Phuket) is important, Burma less

so, and there is still some nesting in Malaysia. In the western Pacific, Papua New Guinea,

the Solomon Islands, and Australia have nesting (see Eckert 1993 for review). The largest

nesting colonies in the eastern Pacific are located in Mexico and Costa Rica. Nesting occurs

on a smaller scale along the Central American coastline, Colombia, Ecuador and, possibly,

Peru.

In the U. S., St. Croix (USVI) and Culebra (Puerto Rico) support the largest nesting colonies

of leatherback turtles. Between 1981-1993, the total number of nests laid per year at Sandy

Point NWR ranged from 82 (1986) to 345 (1992), or 18 to 55 females (McDonald et al.

1993). Sandy Point NWR receives 50-70 % of all leatherback nests in the USVI. Between

1983-1988, 11-52 nests (=3-11 females) were laid each year at Manchenil Beach, St. Croix

(Adams 1988). Shoy's Beach on the north shore of St. Croix receives ± 20 nests per year

and another 15 beaches receive a lower level of nesting (Eckert 1992). Between 1982-1988,

41 leatherback activities (both successful and unsuccessful nests) were reported from Buck

Island off the northeast coast of St. Croix. One and nine crawls were recorded on West

Beach, Buck Island, in 1989 and 1990, respectively; there was no leatherback nesting in 1991

(Zandy Hillis, USNPS, in litt., 11 May 1992). A half dozen beaches on St. John and St.

Thomas rarely exceed one nesting leatherback each per year (Boulon 1987).

On Culebra Island between 1984-1990, 88-184 nests (=12-27 females) were laid per annum

(Tallevast et al. 1990). Resaca and Brava beaches receive 91-100% of the nesting on

Culebra (Tucker 1988); nesting occurs elsewhere in Puerto Rico, but nowhere in large
46

numbers (Cintron and Cintron 1987). In 1991, 18 nests were reported at Playa Añasco,

three nests at Mayagüez, and one nest at Cabo Rojo; between 1986-1991, 149 nests were

reported at Luquillo (Benito Pinto, PR DNR, pers. comm.). Between 1988-1991, 9-26 nests

per year were laid at Playa Humacao (Manuel Corbet, PR DNR, unpubl. data). An

April-June 1991 survey of Vieques Island revealed 30 leatherback nests (B. Pinto, pers.

comm.). Kontos (1986) documented 11 nests on Mona Island between 17 April-19 June

1985; there were no nests in 1986, five crawls in 1987, and none in 1988 or 1989

(Richardson 1990). Nesting on the Atlantic coast of Florida may sometimes approach that

reported in the U. S. Caribbean, but nest density is considerably lower. The number of nests

reported per year in Florida has risen with increasing survey effort from 18 nests in 1979 to

177 nests in 1992 (Meylan et al. 1995). Nesting north of Florida is very rare (see Schwartz

1977, Ruckdeschel et al. 1982). No nesting is reported from areas under U. S. jurisdiction

in the Pacific Ocean (Eckert 1993).

Estimates of foraging population sizes are not available. Sightings and incidental catch data

indicate that leatherbacks (adult or near adult size) are found in Alaska as far north as

60.34EN, 145.38EW and as far west as the Aleutian Islands (Hodge 1979, Stinson 1984). 

Documented encounters extend southward through the waters of British Columbia (Logier

and Toner 1961, MacAskie and Forrester 1962), Washington and Oregon (Eisenberg and

Frazier 1983, Brueggeman 1991), and California (Stinson 1984; Starbird et al. 1993). After

analyzing some 363 records of sea turtles sighted along the Pacific coast of North America

(from 29E45'N northward), Stinson (1984) concluded that the leatherback was the most

common sea turtle in U. S. eastern Pacific waters. The species is also present in U. S.

temperate Atlantic waters where recent efforts to calculate sea turtle population size using

aerial survey techniques have been largely unsuccessful with regard to leatherbacks

(Thompson 1988, Epperly et al. 1990), presumably because of a statistically inadequate
47

number of sightings. In the one published study that made an attempt to quantify density,

Lohoefener et al. (1988) reported 0.027 leatherbacks per 100 km2 during periods of peak

density offshore of Louisiana in October 1987.

A survey of the Cape Canaveral area conducted during March 1982-August 1984 reported

that 94.5 % of all leatherback sightings (n = 128 total) occurred east of the 20 m isobath and

90.6 % occurred during the summer (Schroeder and Thompson 1987). An earlier (1979)

aerial survey of the mid- and north-Atlantic areas of the U. S. Outer Continental Shelf

between North Carolina and Nova Scotia showed leatherbacks to be present April to

November throughout the study area (but most likely to be observed from the Gulf of Maine

south to Long Island); peak estimates of relative abundances during the summer were in the

hundreds (Shoop et al. 1981). The same study concluded that leatherbacks were observed

more frequently in colder waters at higher latitudes during the summer than were other sea

turtle species. In Cape Cod Bay, Massachusetts, sightings peak in August and September

(Prescott 1988). Sightings peak progressively earlier in the year moving south along the

eastern seaboard. In North Carolina public sightings of leatherbacks inshore and offshore are

highest in May (Epperly et al. 1995). In South Carolina abundance peaks in mid-May, based

on annual aerial offshore survey data reporting mixed groups (male, female; adults,

subadults) moving north at this time of year (Sally Murphy, SC WMRD, pers. comm.).

It is not known at the present time whether leatherback populations under U. S. jurisdiction

are stable, increasing, or declining, but there can be no doubt that some nesting populations

(e.g., St. John and St. Thomas, USVI) have been virtually extirpated (Eckert 1992). 

Historical levels of nesting are not known at either Sandy Point NWR or Culebra Island, but

data from continuing studies at these sites show annual fluctuations which do not appear

abnormal and do not project a long-term decline. Even as these important nesting beaches
48

are protected, however, egg poaching remains a serious concern in both the USVI, and

especially Puerto Rico. Furthermore, incidental catch and entanglement at sea are

unquantified and potentially important sources of mortality.

Listing Factors

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

Nesting beaches around the world have been and are being degraded by residential and

commercial coastal construction (particularly high density development), beach-front

lighting, beach armoring (including stabilization structures such as seawalls, groins, and

erosion barriers), mechanized beach cleaning, upland deforestation, sand mining,

uncontrolled recreational use, vehicle traffic, litter, and the removal of stabilizing vegetation. 

General overviews of habitat alteration and its observed consequences are provided by

Bjorndal (1982), Groombridge (1982), Coston-Clements and Hoss (1983), Bacon et al.

(1984b), Hopkins and Richardson (1984), Raymond (1984), National Research Council

(1990), and NMFS\FWS (1992). The potentially damaging effects on leatherbacks of

chemical pollutants in the marine environment, such as industrial effluents, crude oil and

pesticides, have yet to be determined.

At Sandy Point NWR, St. Croix, contemporary problems include egg poaching, domestic

animals (dogs, horses), four-wheel drive vehicles, umbrellas and volleyball nets, cooking and

bonfires, indiscriminate garbage disposal, rape, robbery, loud music, drinking, and drug use

(Greg Hughes, USFWS, in litt., 12 May 1992). Litter such as discarded glass and metal has

caused nesting attempts to be aborted at the Refuge (K. Eckert, pers. observ.) and vehicle

traffic has resulted in death to 20 % - 25 % of hatchlings in nests which have been driven
49

over (Eckert et al. 1984). Many of these problems were eliminated when a policy of dusk to

dawn closure was enacted in 1993, and later further reduced to weekend visitation only. 

Still, problems remain. For example, nesting females and particularly hatchling turtles are

disoriented by coastal lighting and are often attracted, particularly on moonless nights, to the

intense glare of elevated baseball field lighting in Frederiksted several kilometers to the north

(K. Eckert, pers. observ.).

At the other major U. S. nesting ground, the development of Culebra's coastal zone poses

direct threats to nesting and offshore habitats. Beach sand mining, nest compaction and

hatchling death due to vehicular and pedestrian traffic, photopollution, sedimentation from

upland clearing, and anchor damage have been documented at Brava, Resaca, and Zoni

beaches (Teresa Tallevast, USFWS, in litt., 2 June 1992). On the main island of Puerto

Rico, coastal zone development is a chronic threat, bringing increased traffic, litter, and

artificial lighting. The recent construction of a pipeline at Playa Humacao, a beach regularly

visited by leatherbacks, involved bright lights maintained on the beach while construction

proceeded during the nesting season; not a single leatherback ventured ashore during this

time (B. Pinto, pers. comm.). Similar problems plague leatherback nesting colonies

throughout the Caribbean region; for example leatherback nesting beaches have been lost to

sand mining in the BVI (Cambers and Lima 1990), St. Kitts and Nevis (Eckert and

Honebrick 1992), St. Lucia (d'Auvergne and Eckert 1993) and Grenada (Eckert and Eckert

1990a), and elsewhere. Beach-front development is everywhere increasing, and brings with

it a host of threats related to coastal construction and recreation.

Oceanic foraging grounds are also being degraded, primarily as a result of anthropogenic

waste disposed at sea (e.g., UNEP 1984, CEE 1987). There is mounting evidence that the

ingestion of plastic bags, presumably mistaken for jellyfish, is commonplace for leatherback
50

turtles and can be fatal. Eleven of 15 leatherbacks that washed ashore on Long Island, New

York, during the summer of 1982 had plastic bags "totally blocking their stomach openings"

(Anon. 1983). Ten of 33 dead leatherbacks washed ashore on Long Island between 1979 and

1988 had ingested plastic bags, plastic sheets, or monofilament (Sadove and Morreale 1989). 

Plastic bags and film were "common" in the intestinal tracts of leatherback carcasses

discarded by fishermen in Peru (Fritts 1982). Mrosovsky (1981) reviewed stomach content

data and found that in seven of 16 cases (44%), the leatherbacks examined had ingested

plastic. Hughes (1974) reported that an adult female stranded in Natal, South Africa, "had

its duodenal tract completely filled by a sheet of heavy plastic" measuring 3 m x 4 m. Near

Rochelle, France, seven of eight leatherbacks examined had swallowed plastic (Duron and

Duron 1980). These are not isolated data, and suggest that the ability of leatherback

populations to recover depends not only on protection from direct harvest and the

conservation of important nesting habitat, but also on the general health and cleanliness of the

oceans.

Oil spills are a continuing source of concern and the extent of their effect on leatherbacks or

their eggs has not been assessed. In September 1989, following Hurricane Hugo, a 42,000

gallon spill of #6 fuel oil (heavy crude oil) at the Water and Power Authority facility in

Christiansted, St. Croix, left south coast beaches heavily oiled. Between March 1991 and

March 1992, two more spills, both outside of U. S. waters, threatened U. S. nesting

beaches. The first occurred on 6 March 1991, 13 nm north of Nevis, when the

Trinidad-registered barge Vestabella, loaded with about 560,000 gallons of #6 fuel oil, sank

in 600 m of water after a towing cable snapped; the initial oil slick was more than 30 miles

long (Simmonds 1991). According to The Daily News (30 March 1991), a USVI newspaper,

tar balls began appearing on St. John on 21 March and soon thereafter were reported from

St. Thomas, St. Croix, Culebra, Vieques, and the main island of Puerto Rico. One year
51

later, on 15 March 1992, a pipe ruptured during ship-to-shore pumping of #6 fuel oil to a

transfer station at St. Eustatius Terminal, Netherlands Antilles. One hundred barrels of crude

oil were released to the sea in a slick that headed northwest out across the rich fishing

grounds of the Saba Bank. Heavy seas broke up the slick before it entered U. S. waters, but

tar balls eventually fouled the coast of Puerto Rico (Z. Hillis, pers. comm., 1992).

2. Over-utilization for Commercial, Recreational, Scientific, and Educational Purposes

There is virtually no international commerce in leatherback turtle products. Nonetheless,

local commercial and subsistence exploitation is heavy in many parts of the world. The

literature is replete with accounts of the harvest of eggs and/or adults in all major breeding

areas. In the Western Atlantic, leatherbacks are widely taken for both meat and oil, the latter

consumed for general vitality or employed medicinally for respiratory congestion. Tag

returns indicate that this take occurs at least in part in populations that migrate between

foraging grounds under U. S. jurisdiction and tropical nesting beaches outside U. S.

jurisdiction. Prior to recent conservation programs involving indigenous peoples (Tambiah

1992), an estimated 80 % of nesting females were slaughtered during their attempts to nest

on beaches in Guyana (Pritchard 1986). Ross and Ottenwalder (1983) reported that nearly

100 % of gravid females were harvested by local people for food in the Dominican Republic. 

The harvest of adults continues in the BVI (Cambers and Lima 1990), Grenada (Eckert and

Eckert 1990a), St. Lucia (d'Auvergne 1992), Trinidad (Chu Cheong 1990), and elsewhere. 

In addition to harvest on tropical nesting beaches, there are several reports of leatherbacks

shot or harpooned in temperate Atlantic (Brongersma 1972) and Pacific (Stinson 1984)

latitudes.
52

The harvest of eggs also persists throughout the turtle's breeding range, including much of

the Western North Atlantic. In Puerto Rico, where the collection of sea turtle eggs has been

illegal for many years, Matos (1986) reported a "flourishing black market offering high

prices for meat and eggs" and concluded that the illegal hunting of nesting females and eggs

was a threat to population recovery. As recently as 1992 there was evidence of egg poaching

at major U. S. rookery sites; i.e., Sandy Point NWR and Shoy's Beach on St. Croix (G.

Hughes, pers. comm., 1992) and on Culebra (T. Tallevast, pers. comm., 1992). In Pacific

Mexico, where leatherbacks feeding off the west coast of the U. S. presumably go to

reproduce, egg collection is widespread despite laws prohibiting it (L. Sarti M., pers.

comm., 1994). Further south, the harvest of eggs by local people for food or sale is so

intensive that few nests laid from Costa Rica to Colombia can be expected to survive

(Pritchard 1989). In Malaysia, a predominantly Islamic nation where the consumption of sea

turtle meat is forbidden by religious custom, virtually 100% of the eggs were collected

(Chan et al. 1985) prior to a 1989 moratorium on egg harvest. This exploitation is

implicated in the steady decline of that population. Between 1967-1976 and 1984-1988,

based on mean values, there was an 86% drop in the number of turtles visiting the

well-studied Terengganu rookery (Chua 1988).

The effects of utilizing leatherback turtles for scientific purposes are not likely to negatively

affect remaining populations. Leatherback research consists primarily of population surveys,

hatchery programs, and other activities which do not involve the death or debilitation of

individuals. In the United States, the use or take of leatherbacks for scientific purposes (e.g.,

captive diet studies, blood sampling, deploying remote equipment) is controlled by a permit

system designed to protect endangered and threatened species. Live leatherbacks beyond the

hatchling stage are difficult to maintain in captivity and are not intentionally captured for

recreational or educational purposes.
53

3. Disease or Predation

The extent to which disease or parasites influence the survival prospects of wild leatherback

turtles is unknown, but these are not considered serious threats. Wolke (1981) reported a

case of enteritis. Ogden et al. (1981) diagnosed hematogenous septic arthritis and

osteomyelitis involving the elbow, distal humerus, and proximal radius and ulna in a 135 cm

adult stranded on the New England coast after being struck in the head by a boat propeller. 

Rothschild (1987) documented avascular necrosis of bone in Oligocene and Eocene members

of the Dermochelyidae; the extent to which the phenomenon debilitates modern dermochelids

is unknown. Analysis of adipose tissue, as well as samples of liver and pectoral muscle,

from an adult male leatherback stranded in Wales in 1988 indicated that heavy metal values

were not elevated above 'normal' background levels seen in other organisms (Davenport et

al. 1990, Davenport and Wrench 1990).

Internal parasites include intestinal nematodes (species unidentified) and trematodes

(Pyelosomum [=Astrorchis] renicapite, Calycodes anthos), amoebae (cf. Entamoeba

histolytica), and gall bladder flukes (Cymatocarpus sp.) (Dunlap 1955, Yerger 1965,

Brongersma 1972, Threlfall 1979). An external parasitic isopod (Excorallana antillensis)

was reported on gravid leatherback nesting on St. Croix by Eckert and Eckert (1988). 

Several barnacle species appear to colonize leatherback turtles, including Stomatolepas

dermochelys [=elegans], Platylepas hexastylos, Balanus trigonus, Conchoderma auritum, C.

virgatum, and Lepas anatifera (see Eckert and Eckert 1988 for review).

Ghost crabs (Ocypode quadratus), vultures (Coragyps atratus), monitor lizards (Varanus

sp.), genet cats (Genetta sp.), water mongooses (Ictonyx sp.), night herons (Nyctanassa

violacea), gray foxes (Urocyon sp.), raccoons (Procyon lotor, P. cancrivorus), coatis (Nasua
54

nasua), fly larvae (Megaselia scalaris), locust larvae (family Acrididae), ants, and a variety

of bacteria and fungi prey on eggs and hatchlings in the nest (e.g., Pritchard 1971, Fretey

1981, Whitmore and Dutton 1985). Exotic species, such as the wild pig (Sus scrofa

sulawensis) in New Guinea (Bhaskar 1985) and the mongoose (Herpestes auropunctatus) in

the Caribbean (Eckert et al. 1984), also consume eggs and hatchlings. Rates of depredation

do not appear to pose an important threat to U. S.-nesting populations. Losses to non-human

predators at Sandy Point NWR are estimated to be < 0.5% of annual productivity (Eckert et

al. 1984). In contrast, at Les Hattes, French Guiana, the loss to domestic dogs of nearly

1000 hatchlings per night has been documented (Fretey 1981).

Crocodiles periodically consume nesting adults in Papua New Guinea (Lockhart 1989). 

Jaguars and tigers may have been historical predators on nesting adults mainly in the

Guianas, but are less important today (Pritchard 1971, Henri Reichart, STINASU, pers.

comm., 1994). At sea leatherback hatchlings are consumed by fishes, having been found in

the stomachs of carnivorous fishes in the waters of French Guiana (Fretey 1981). Juveniles

and adults are preyed upon by killer whales (Orcinus orca) (Caldwell and Caldwell 1969,

Sarti et al. 1994), crocodiles (Pritchard 1981), and, presumably, sharks. Rates of

depredation on any life stage in the open ocean are unknown.

4. Inadequacy of Existing Regulatory Mechanisms

In the United States, both national and international laws protect leatherback turtles from

harassment, harvest, and commerce. According to the Code of Federal Regulations (50 CFR

17.11), the leatherback sea turtle is listed as Endangered throughout its entire range under the

U. S. Endangered Species Act (ESA) of 1973, as amended (35 FR 8495; June 2, 1970). 

Similarly, the species is classified as Endangered in the International Union for Conservation
55

of Nature and Natural Resources' (IUCN) Red Data Book, where taxa so classified are

considered to be "in danger of extinction and whose survival is unlikely if the causal factors

continue operating" (Groombridge 1982). Leatherbacks are included on Appendix I of the

Convention on International Trade in Endangered Species of Wild Fauna and Flora (CITES). 

The U. S. ratified CITES in 1974.

All federal agencies must ensure that actions authorized, funded, or carried out by them do

not result in the destruction or adverse modification of habitat designated as Critical Habitat

for marine turtles pursuant to Section 7 of the ESA. Critical Habitat has been designated by

both the U. S. Fish and Wildlife Service (USFWS) and the National Marine Fisheries Service

(NMFS). The beach at Sandy Point, St. Croix, became the first nesting beach for any

marine turtle to be proposed as Critical Habitat (Federal Register, 23 March 1978; 43 FR

12050-12051) (Dodd 1978). In September 1978, FWS determined the perimeter of the

Sandy Point peninsula (3.06 km in length and 0.32 km wide from MHT inland) to be Critical

Habitat for leatherback turtles. In March 1979, NMFS designated the surrounding waters,

following the 100 fathom contour from 17E42'12"N around the peninsula to 64E50'00"W, as

Critical Habitat (see USFWS 1981). In September 1984, FWS purchased 2.4 km of this

nesting beach, establishing Sandy Point NWR. With three species of sea turtle nesting at this

site, Sandy Point is one of the most unique endangered species Refuges in the U.S.. Yet a

manager was not hired until 1990, and the Refuge still lacks an informational kiosk and other

interpretive materials. 

The most important leatherback nesting beaches in Puerto Rico, namely Resaca and Brava on

Culebra Island, were designated Critical Habitat in 1982 (effective 26 July), but the

designation was for the protection of hawksbill, not leatherback, turtles (Dodd 1978). There

is a National Wildlife Refuge on the island which includes a portion of the upland areas
56

behind the major leatherback nesting beaches (Playa Resaca, Playa Brava), but Culebra NWR

does not encompass the beaches themselves. These two beaches are managed by the

Commonwealth of Puerto Rico, with technical assistance and enforcement back-up provided

by the USFWS (U. S. Department of the Interior). A paucity of proper land use planning

and zoning, as well as a lack of rigorous or consistent enforcement of existing zoning and

natural resource regulations by PR DNR and a corresponding inability or unwillingness on

the part of the USFWS to enforce environmental policies, has had calamitous results at some

nesting beaches, including uncontrolled coastal development, sand mining, erosion, and egg

poaching.

Broad legal protection has been effective in eliminating egg and turtle poaching in the

continental USA, but not so in Puerto Rico or the USVI. On the island of Puerto Rico, the

theft of eggs for local consumption is nearly ubiquitous (Matos 1986, Cintron and Cintron

1987). Inadequate law enforcement is a widely recognized problem. Some progress has

been made in Culebra. Egg poaching on this satellite island was once described as "extensive

and unrelenting" (Carr 1978), but has noticeably declined on some beaches as a result of

nightly patrol and nest protection programs initiated in 1984 (Tucker 1988). Egg theft also

continues in the USVI, principally on St. Croix. The Virgin Islands Department of Planning

and Natural Resources' Law Enforcement Division lacks sufficient funds and manpower to

fully cover the beaches and phones are unmanned on weekends when most people observe

violations (Z. Hillis, in litt., 11 May 1992). An ongoing egg protection program at Sandy

Point NWR has greatly reduced (but not eliminated) egg poaching and has resulted in

substantial increases in annual productivity (Eckert and Eckert 1990b). Regulations for the

protection of leatherbacks and their eggs are adequate, but vigilant enforcement and more

visible public awareness campaigns are clearly needed.
57

At sea, leatherbacks are vulnerable to incidental capture. In addition to the take of

leatherbacks by the shrimp trawling industry (e.g., Henwood and Stuntz 1987), NMFS has

information that leatherbacks are incidentally taken in other types of commercial fishing gear

in U. S. waters, particularly pelagic longlines and drift/gillnets (Wetherall et al. 1993). The

agency has issued a Biological Opinion on this subject, evaluating a variety of fisheries with

regard to the incidental catch problem (NMFS 1989). The use of turtle excluder devices

(TEDs) is mandatory in the U.S. (Crouse 1993) and federal law also requires that foreign

countries exporting shrimp to the U.S. enact and enforce sea turtle conservation measures,

including the use of TEDs (Ple 1990). Not all TEDs will release leatherbacks because of

their large size. There is, however, a provision in the U. S.-prepared regulations to close

areas to trawling or greatly restrict tow times to protect turtles such as leatherbacks (C.

Oravetz, pers. comm., 1992).

Recent progress has been made on the issue of incidental capture in waters beyond national

jurisdiction. On 20 December 1991, the United Nations General Assembly passed a driftnet

resolution co-sponsored by the United States and Japan that called upon all nations to reduce

fishing effort in existing large-scale pelagic driftnet fisheries by 50 % by 30 June 1992 and to

fully implement a global moratorium on all such driftnets by 31 December 1992. The

resolution passed by consensus. In addition to incidental capture, leatherbacks are also

threatened by marine pollution. A variety of legal regimes exist to address the problem of

ocean dumping both regionally and internationally (see Bean 1987, Lentz 1987). These

include the Law of the Sea Convention, The London Dumping Convention, the International

Convention for the Prevention of Pollution from Ships (MARPOL), and a host of regional

agreements, such as those sponsored by the UNEP Regional Seas Programme. Despite

achievements in these areas, the problems persist on a grand scale. International support for
58

stringent regulations governing indiscriminate fishing and ocean dumping is needed, as is a

mechanism for effective enforcement.

5. Other Natural or Manmade Factors

Incidental capture and entanglement at sea in active and abandoned fishing gear present

serious problems for leatherback turtles worldwide (e.g., Brongersma 1972, Balazs 1985,

Wetherall et al. 1993). Strandings reported from U. S. Atlantic, Gulf of Mexico, and

Caribbean coasts increased 123 % in 1987 (n = 134) over 1986 ( n = 60) (Schroeder 1988). 

Strandings declined in 1988 (n=79), but have since risen again (n=104 in 1989, n=102 in

1990, n=118 in 1991, n=123 in 1992) (Wendy Teas, NMFS, pers. comm.). In April 1991,

30 leatherbacks washed ashore dead in Georgia, nearly equally the total of 39 during the

previous 11 years; shrimp trawling was implicated (Shoop 1991). As of 1 June 1992, a

record number of 28 adult and subadult leatherbacks had stranded on the beaches of South

Carolina; gillnets and trawlers were the most likely causal agents. Six of the South Carolina

turtles were fresh enough to necropsy. None of these had obvious injuries and all had a gut

full of Stomolophus jellyfish. In three cases, froth in the trachea suggested recent drowning. 

In addition, some stranding network reports noted post-mortem and/or ante-mortem injuries,

including skull fractures (Sally Murphy, SC WMRD, pers. comm.).

Summarizing a decade (1977-1987) of data, Prescott (1988) implicated entanglement (mainly

in lobster pot lines) in 51 of 57 (89 %) adult leatherback strandings in Cape Cod Bay,

Massachusetts. Goff and Lien (1988) reported that of 20 leatherbacks encountered off the

coasts of Newfoundland and Labrador (1976-1985), 14 (70 %) were entangled in fishing

gear, including salmon nets, herring nets, gill nets, trawl lines, and crab pot lines. Gill net

entanglement is also suspected in the waters of New England, based on observations of
59

leatherbacks with cuts, severed limbs, or chafing marks (National Research Council 1990). 

Further south the species is a rare but regular casualty of shrimp trawls, and some individuals

have become "entangled in the gangion or caught on the hook" of longlines set in the

northern Gulf of Mexico (Hildebrand 1987), southeastern U. S. (Witzell 1984), and northern

Caribbean (Cambers and Lima 1990, Tobias 1991). On 21-22 March 1992, five 300-foot

gill nets were legally set as close as 10 m from shore in Critical Habitat off Sandy Point

NWR. There is no evidence to indicate that gravid females were caught during this incident,

but the fishermen involved readily admitted that leatherbacks had been ensnared in the past

and killed in order to save the net (G. Hughes, pers. comm., 1992).

In the eastern North Pacific leatherbacks become entangled in gill nets off the coasts of

California, Oregon, and Washington (Stick and Hreha 1989, Scott Eckert, HSWRI, pers.

comm., 1994). Eleven leatherbacks were captured in gill nets by a single fishermen from

Bahia de la Paz, Baja California, between 1985 and May 1987 (Alvarado and Figueroa

1990). Both adults and juveniles are occasionally caught in tuna purse-seines operating in the

eastern tropical Pacific (S. Eckert, pers. comm., 1994). A very young individual (about 15

cm carapace length) was captured in a purse-seine in April 1976 about 180 nm west of San

José, Guatemala (Robert Pitman, NMFS, pers. comm.). Balazs (1982) documented the

death of leatherbacks in pelagic drift nets. Further south, Frazier and Brito Montero (1990)

reported that "at least several hundreds" are captured per annum incidental to the swordfish

industry in Chile. Finally, leatherbacks are struck, sometimes fatally, by propellers

(Brongersma 1972, Stinson 1984). Leatherbacks are also incidentally captured in the water

intakes of industrial facilities, such as power plants. Three were trapped in the St. Lucie

Power Plant (Hutchinson Island, Florida) in 1978 and two more in 1981 (Roithmayr and

Henwood 1982); a total of eight were reported between 1976 and 1988 (Ernest et al. 1989). 

Subsequent captures occurred in 1989 and 1992, one turtle in each case. The size range of
60

leatherbacks involved has been 112.5-150.0 cm SCL; all were released unharmed (Eric

Martin, Applied Biology, pers. comm., 1992).

On the nesting beach, storm events and seasonal erosion degrade or destroy nesting beaches

and can result in egg losses ranging from < 2.5% in Malaysia to 50% or more in the

Guianas and the U. S. Virgin Islands (Mrosovsky 1983, Eckert 1987). Eggs may also be

destroyed by females digging into nests constructed earlier in the season (Fretey 1986). 

Shimmering inland lagoons, or debris that blocks the ocean's bright horizon, can disorient

hatchlings and preclude their passage to the sea (Mrosovsky 1983). Adult females in French

Guiana are sometimes unable to reach the sea after nesting, having become trapped behind

accumulations of dead trees, impaled against a root or branch, or held prisoner between

stumps deeply embedded in the sand (Pritchard 1971, Fretey 1977, 1981). Females

disoriented by an inland lagoon or swamp may mistake its reflection for that of the sea and

orient toward it after nesting. Turtles mired in the soft mud of a mangrove swamp or river

delta can die from exposure, exertion or asphyxiation.

Conclusion

While the leatherback sea turtle enjoys complete protection in the U. S., the species is often

inadequately protected elsewhere. The direct and/or indirect harvest of adults continues in all

major and most minor breeding grounds. This despite the fact that the IUCN considers the

species Endangered throughout its global range; that is, "in danger of extinction and whose

survival is unlikely if the causal factors continue operating" (Groombridge 1982). 

In Mexico, the largest breeding colony of leatherback turtles in the western hemisphere has

suffered enormous losses of both adults and eggs in recent decades (Pritchard 1982, Eckert
61

1993). The extent to which the recent ban on sea turtle harvest in Mexico (Aridjis 1990) will

ultimately mitigate this situation remains unknown. Referring to Irian Jaya, host to the

largest breeding colony of leatherback turtles in the eastern hemisphere, Bhaskar (1985)

wrote, "The proliferation of outboard engines in the area in the last decade has made the once

remote nesting beaches easily accessible to egg collectors who remove clutches by the

boatload. On less favored beaches ... to the southeast, nesting leatherbacks are killed for

food by inhabitants newly settled on the coast. In years to come, the turtle population will

undoubtedly be under even greater human pressure as transmigration of Indonesian families

into Irian Jaya continues." This prediction appears to have come to pass. Recent reports

indicate an alarming decline in the nesting population at the Jamursba-Medi site (Betz and

Welch 1992).

Pritchard (1981) reviewed the status of leatherback turtles in the Solomon Islands and

concluded that there had been a gradual decline over the years, ominous but difficult to

document. While not all leatherback populations are declining, Pritchard's words accurately

describe the situation throughout much of the species' range. Perhaps it is fortunate in a

backhanded way that in some areas these seemingly innocuous gradual declines have matured

into population collapse. By examining these disasters, by evaluating cause and effect, we

may be able to avert similar catastrophes elsewhere. One case in point is Terengganu,

Malaysia, where nesters arrived by the low thousands per annum 25 years ago and now

arrive in the tens. The collapse was precipitated not by the harvest of turtles, but of eggs. 

As a generation of breeders has died out, there has been no significant recruitment. Another

illustrative example is the British Virgin Islands, where a subsistence harvest of gravid

females for meat and oil has extirpated a once thriving population. The nature of sea turtle

life history allows nesting turtles and/or eggs to be over-harvested for decades from large

populations, seemingly without consequence (Mortimer 1995). By the time declining trends
62

are statistically perceived amid the background noise of normal annual fluctuations, a

population may be committed to a downward spiral from which it is not capable of

recovering. Malaysia recently implemented a moratorium on the harvest of leatherbacks and

their eggs and the BVI is attempting to do the same. In both cases it is likely that the

decision has come too late. Many leatherback populations have been seriously

over-harvested in past decades and will need wise stewardship into the next century if they

are to avoid potentially irreversible declines.

Causal factors operative in the marine environment have not been satisfactorily quantified but

appear serious. Thompson et al. (1990) concluded that intensifying human utilization of

deeper offshore waters is placing all life history stages of the leatherback turtle at "increasing

risk". All indications are that debilitation, particularly of large juveniles and breeding age

adults, due to entanglement in and ingestion of persistent marine debris is high. In addition,

this status review has summarized incidental capture and drowning in a wide variety of

fishing gear. On the basis of these data it would not be unreasonable to conclude that the

status of foraging populations occurring in U. S. jurisdiction has worsened since protection

was first afforded the species under the Endangered Species Act nearly 20 years ago. It is

noteworthy that animals foraging in the U. S. include juveniles which will eventually recruit

into the breeding aggregations of the western Atlantic and eastern Pacific, where they will

(should present circumstances prevail) face considerable threat at the nesting grounds. The

best available commercial and scientific data indicate that the leatherback sea turtle

(Dermochelys coriacea) should remain listed as Endangered throughout its range pursuant to

Section 4 of the Endangered Species Act of 1973, as amended.
63

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76

Hawksbill Sea Turtle, Eretmochelys imbricata

Prepared by Karen L. Eckert, Ph.D.

Biological Background

Hawksbill sea turtles (Eretmochelys imbricata) are distinguished from other sea turtles by two

pairs of prefrontal scales; thick, posteriorly overlapping carapace scutes; four pairs of costal

scutes, the anterior-most not in contact with the nuchal scute; and two claws on each flipper. 

Some scute variation occurs in both adults and hatchlings (Limpus et al. 1983). The carapace

is typically serrated along the posterior margin (becoming less so with age) and is

"tortoiseshell" in color and pattern, the dorsal laminae showing radiating streaks of brown,

black and amber. Carapace color is geographically variable and may also change with age

(see Witzell 1983 for review). The head is relatively narrow, the beak tapers to a point, and

the maxilla projects slightly beyond the mandible. The scales of the head are dark brown

with pale yellow margins. Both adults and hatchlings have a normal tetrapod gait while on

land, with alternating movements of opposing flippers (Pritchard 1979). Two subspecies, E.

i. imbricata in the Atlantic Ocean and E. imbricata squamata in the Pacific Ocean, have been

described on the basis of differences in coloration and carapace shape. The criteria have

proven unreliable in distinguishing the two forms, however, and subspecific designations are

rarely used (Meylan 1984a). A general synopsis of the biology of the species is provided by

Witzell (1983) and more recent, in-depth discussions can be found in Pritchard and Trebbau

(1984) and Groombridge and Luxmoore (1989).

Mean straightline carapace length (SCL) of adult females ranges from about 66 to 86 cm

worldwide (Witzell 1983). Meylan (1984b) reported that adult females in the Caribbean

measured 62.5-91.4 cm SCL. Weight is typically to 80 kg (Pritchard et al. 1983), with an

historical record of a 280 lb [127 kg] individual caught at Great Sound, Grand Cayman (Carr

1952). Nietschmann (1981) observed little dimorphism between adult males and females in

Caribbean Nicaraguan waters. He reported average carapace lengths of 77.8 cm (range

71.4-85.1, n=17) and 76.5 cm (range 62.5-87.0, n=32), respectively, and average weights

of 53.4 kg and 54.2 kg, respectively. Adult males are distinguished by a long, thick tail that
77

extends well beyond the carapace margin and well developed, recurved claws on the fore

flippers (Witzell 1983). In the U.S. Caribbean, the curved carapace length (CCL) of nesting

females averages about 87 cm (n=61) (Hillis and Mackay 1989, Richardson 1990). At Buck

Island, situated off the north shore of St. Croix, the CCL of nesting females ranged from

78.7-100.0 cm during 1988-1991 (Zandy Hillis, USNPS, in litt., 11 May 1992). Hatchlings

emerging at Buck Island average 41 mm SCL and weigh 14.0-17.5 g (Hillis and Mackay

1989). At Sandy Point National Wildlife Refuge (NWR), St. Croix, egg diameter averages

39.9 mm (n=36 eggs); hatchlings average 42.4 mm SCL and weigh 18.5-19.5 g (n=60

eggs) (Eckert, unpubl. data). Morphometric data are not available for this species in the

U.S. Pacific.

Hawksbills are distributed circumglobally and appear to be the most confined of all the sea

turtles to tropical waters, although individuals are sometimes encountered in northern

latitudes (e.g., Ireland: Penhallurick 1990). The species is often characterized as sedentary,

or as having "given up migration" (Hendrickson 1980). This stereotype has persisted in part

because of early reports (Carr and Main 1973) that carapace color and morphology varied by

island (or island group) in the southwestern Pacific and were unique, allowing indigenous

peoples to discern distinct populations over rather restricted geographic areas. There is

skepticism, however, that such uniformity exists (Limpus et al. 1983). The hypothesis that

the species is sedentary is further eroded by records of long distance movements of several

hundred to several thousand km (e.g., Carr and Stancyk 1975, Nietschmann 1981, Vaughan

1981, de Silva 1982, Meylan 1982, Parmenter 1983, Marcovaldi 1991, Hillis 1994) as well

as data obtained via satellite telemetry revealing long distance post-nesting movements

(Groshens and Vaughan 1994). In the Caribbean region, juvenile hawksbills tagged in the

USVI have been recovered in the BVI, Puerto Rico, St. Lucia, and St. Martin (Boulon

1989). Hawksbills tagged in Antigua later have been recaptured in Dominica (Fuller et al.

1992) and St. Kitts (K. Orchard, SCHS, pers. comm., 1994). On 20 July 1990 a juvenile

hawksbill (74 cm SCL) tagged six months before at the Biological Reserve of Atol das Rocas

in Brazil, was captured and killed in Dakar, Senegal, some 2,300 miles away (Marcovaldi

1991). It is likely that additional tagging in nesting and foraging habitats will corroborate the

hypothesis that migratory behavior occurs in hawksbills.

Hawksbills forage in coral reefs and other hard-bottom habitats in open bays and coastal

zones throughout the tropics and, to a lesser extent, the subtropics, including Florida, the
78

Gulf of Mexico, the USVI, Puerto Rico, Hawaii, and the U.S. Pacific territories. Despite a

wide variety of foods consumed (see Witzell 1983), recent studies indicate that hawksbills

specialize on sponges in the Caribbean, and predominately on two orders of Demospongea

(Meylan 1988, vanDam and Diez 1994, Vicente 1994). Ten species of sponges accounted

for 79.1 % of the dry mass of all sponges identified in the stomachs of hawksbills from seven

Caribbean countries (Meylan 1988). The predominance of specific taxa in the digesta

suggests a degree of selectivity, perhaps related to distinctive properties of the sponges with

respect to spongin and collagen (Meylan 1985). Few data are available from the Pacific, but

Balazs (1978a) reported that the stomach and intestines of a dead hawksbill (75.6 cm SCL)

entangled in a monofilament gillnet in Kanehoe Bay, Hawaii, were filled with three kinds of

unidentified sponges. Hawksbills are solitary, and intraspecific competition for food

resources has not been documented. Illness and even death have been attributed to the

consumption of hawksbill meat in Sri Lanka, China (Taiwan), Philippines, Indonesia, Papua

New Guinea, and Australia (Torres Strait), as well as Central Pacific and Caribbean islands

(Halstead 1970, Hashimoto 1979, Márquez 1990). Most authors speculate that the toxins

originate in food items consumed by the turtle.

Mating occurs on the surface in shallow waters adjacent to the nesting beach and may last

several hours; polyandrous breeding behavior is implied (Witzell 1983). Gravid females

most often nest on isolated (typically insular) nesting beaches with well developed

supralittoral vegetation. Nests are commonly, but not universally, placed amongst woody

vegetation (Mortimer 1982, Ryder et al. 1989). During 1984 on Mona Island, Puerto Rico,

all nesting attempts except one occurred 1-6 m into the vegetation, generally Suriana

maritima (Olson 1985). Nesting is mainly nocturnal, but some populations nest infrequently

(Limpus et al. 1983) or nearly entirely (Diamond 1976) in the daytime. The nesting

sequence consists of landing, selecting a nest site, clearing the site and excavating a shallow

body pit, excavating the nest chamber, egg laying, refilling the nest chamber and concealing

the site, and returning to the sea. The entire procedure requires an average of 130 min at

Buck Island, USVI (Hillis and Mackay 1989) and 126 min (range 72-196, n=20) in Antigua

(Jim Richardson, Univ. Georgia, pers. comm.), the lengthiest stages being nest excavation

and site camouflage. Average clutch size (yolked eggs) varies geographically and appears to

be strongly correlated with mean carapace size among rookeries (Limpus et al. 1983, Witzell

1985). Non-yolked eggs are uncommon, but in some cases have comprised up to 39 % of

the clutch (Hirth 1980).
79

Nesting is seasonal, but the season is often extended and at a few localities nesting may occur

throughout the year with one or two peaks (Groombridge 1982). Long-term nesting studies

are rare, but one such study began on Cousin Island, Seychelles, in 1973, where about 25

females arrive to nest each year. Cousin Island hawksbills deposit an average of four

clutches (mean=182 eggs) per season (September-March) at intervals of 14-18 days and eggs

incubate for 47-77 days (mean=60 days, n=16) (Diamond 1976). Brooke and Garnett

(1983) later reported a mean clutch size of 161 eggs (the variation apparently due to different

approaches to counting egg shells), an annual clutch frequency of 2.76 (as opposed to four),

with remigrants more productive and neophytes slightly less so, and an average remigration

interval of 2.67 yr. In the western hemisphere, 246 nesters were tagged at Tortuguero, Costa

Rica, between 1955 and 1983. Data from this long-term study suggest that clutches

(mean=158 eggs) are deposited at average intervals of 16.4 days (n=28) and incubate for

56-61 days (mean=58.5 days, n=6) before hatching. Because of deficiencies in beach

coverage, clutch frequency and remigration intervals were not calculated (Bjorndal et al.

1985), but an earlier report noted that the most commonly observed remigration interval at

Tortuguero was three years (Carr and Stancyk 1975).

An ongoing study at Jumby Bay, Antigua, unique in having 100 % of the females tagged,

provides the most complete record of hawksbill reproductive ecology available. Some

nesting occurs in every month of the year, but the major season is mid-June to

mid-November. Females arrive at the nesting beach asynchronously. In 1988, 39

individuals deposited an average of 4.4 clutches (mean=147 eggs/clutch, range 70-203) each

at mean intervals of 14.8 days during June-November; the modal number of nests per turtle

was five. Mean incubation time was 68 days and hatch success averaged 85 % (Corliss et al.

1989). Similar results were obtained in 1987 and 1989, when clutch frequency averaged 4.8

and 4.7 clutches per female and mean hatch success was 79 % and 84 %, respectively (J.

Richardson, unpubl. data). Based on this systematic field effort, it is clear that stock

assessment formulae relying on annual clutch frequencies of fewer than five per turtle are in

danger of overestimating total numbers of nesting females (Richardson et al. 1989). 

Remigration to the Jumby Bay site is high, with 95% of the 1987 cohort (20/21) returning

again in 1989 (n=16) or 1990 (n=4) and 66 % of the 1988 cohort (26/39) returning again in

1990 (n=11) or 1991 (n=15) (J. Richardson, unpubl. data).
80

Hawksbills are studied on beaches in the U. S. Caribbean at Virgin Islands National Park (St.

John), Buck Island Reef National Monument (St. Croix), and Mona Island (Puerto Rico). 

The nesting season on St. John extends from June to December (peak: August-September)

and clutch size and incubation average 141.6 eggs (n=39) and 64 days (n=28), respectively

(Small 1982). Annual reports for the Buck Island project indicate that ca. 15-30 hawksbills

nest each year, annual mean clutch size ranges from 137.3-153.4 eggs (n=262 nests),

inter-nesting intervals average 14-15 days (n=77), and peak nesting occurs in

July-September. Annual hatch success for nests that reach full-term without obvious

disturbance from predators, erosion, or poachers ranges from 77 % to 83 %. On Mona

Island there have been 738 nesting records reported (1113 estimated) during seven survey

seasons since 1974, equivalent to some 36 nesting females per annum (Richardson 1990). 

The Mona Island data show that clutch size averages 141.0 (1989) to 157.6 (1984) eggs,

inter-nesting intervals are typically 14-15 days, incubation requires 47 to 63 days, and

roughly 75 % of the nests are laid during August-November (Richardson 1990). Similar

studies have not been undertaken in the U.S. Pacific.

Mrosovsky et al. (1995) evaluated the effect of incubation temperature on sex determination

in hawksbill hatchlings. Incubation temperatures warmer than approximately 29.2EC

produced females, while cooler temperatures produced males (Mrosovsky et al. 1995). 

Dalrymple et al. (1985) reported a heavily male-biased sex ratio in a Florida nest that was

laid late in the nesting season (25 October 1981) and incubated for some 91 days; maximum

ambient temperature rarely reached 30EC. Once hatchlings leave the natal beach, virtually

nothing is known of their distribution, abundance or survival. A brief study conducted in

Puerto Rico suggests that hatchlings orient toward open ocean once they enter the surf (Hall

1987). There is some evidence that hatchlings and post-hatchlings may associate with

Sargassum weed in the pelagic zone (Meylan and Carr 1982, Carr 1987). Juveniles of

various size classes can be observed in hard bottom habitats throughout the tropics (in the

U.S. they are predictably found as far north as southeastern Florida and the northern Gulf of

Mexico), but no systematic study of their distribution, behavior, or survivorship has been

undertaken.

Data from the capture and recapture of hawksbills in St. Thomas, USVI indicate that wild

juveniles (size range 27.4-60.7 cm SCL) grow an average of 3.36 cm per year (Boulon

1983), while juvenile green turtles (size range 26-62 cm SCL, n=41) in the same habitat
81

average 4.8 cm per year (Boulon and Frazer 1990). In contrast, juvenile hawksbills (40-70

cm SCL, n=4) in the southern Bahamas grew at a rate of 2.4-5.9 cm per year, slightly

faster than co-occurring green turtles (n=62) in comparable size classes, and increased in

mass at a rate about 1.5 times greater than did the green turtles (Bjorndal and Bolten 1988). 

Mature females grow at an average rate of 0.3 cm per year at Tortuguero, Costa Rica

(Bjorndal et al. 1985). In the Pacific, nearly two decades (1969-1988) of systematic study in

the Great Barrier Reef (GBR) of Australia have shown that (a) individuals recruit to the GBR

feeding grounds at a minimum of 35 cm CCL, (b) there are significant differences in growth

rate among size classes, with the maximum mean growth rate (2.17 cm/ year) recorded for

turtles in the 50-60 cm CCL range, and (c) a small turtle (35 cm CCL) recruiting to the

feeding grounds can be expected to begin breeding about 31 years later (Colin Limpus,

Queensland Turtle Research, pers. comm., 1992). Age at maturity is unknown.

Population Size

Nesting beaches are distributed circumglobally, roughly from 30EN to 30ES, and can be

identified in the Eastern and Western Atlantic, Eastern (rare), Central and Western Pacific,

and Indian Ocean; the non-nesting range is equally extensive (Witzell 1983). There are no

world population estimates for hawksbill turtles, but a minimum of 15,000 to 25,000 females

are thought to nest annually in more than 60 geopolitical entities (Groombridge and

Luxmoore 1989). Nesting usually occurs at low densities. Moderate population levels

appear to persist around the Torres Strait islands, in the Red Sea ["possibly 500 nests

yearly", Frazier and Salas 1984] and Gulf of Aden, and probably around the Arnavon Islands

(Solomons), northern Australia, Palau, Persian Gulf islands, Oman, and parts of the

Seychelles (Groombridge 1982). In a more recent review, Groombridge and Luxmoore

(1989) list Papua New Guinea [but see Ehrenfeld 1990], Queensland, and Western Australia

as likely to host 500-1000 nesting females per year, while Indonesia and the Seychelles may

support >1000. The largest known nesting colony in the world is located on Milman Island,

Queensland, Australia where Loop (1995) tagged 365 hawksbills nesting within an 11 week

period. With the exception of Mexico (and Cuba?), nearly all Wider Caribbean countries are

estimated to receive <100 nesting females per year (Meylan 1989). 

Recent surveys have documented relatively large nesting colonies on the shores of the

Yucatan Peninsula (see Frazier et al. 1993). As many as 800-1000 nests are laid annually
82

from Isla Holbox (Quintana Roo) to Isla Carmen (Campeche) (Richard Byles, USFWS, pers.

comm.). A recent review by Frazier (1995) indicated 150 to >300 nests per year in

Campeche (Isla Aguada-Chenkán; possibly this is only 50 % of the annual total), >50 nests

at Celestun-El Palmar, 200-400 nests at Las Coloradas-El Cuyo, and about 200 on Isla

Holbox. Regionally important colonies outside of Mexico exist in Nicaragua, Cuba, the San

Blas Islands and Bocas del Toro region of Panama, Mona Island (Puerto Rico), and the

Grenadines (Hopkins and Richardson 1984), as well as on the Manabique Peninsula,

Guatemala (Rosales-Loessener 1987), near Manatee Bar, Belize (Smith 1992), and on Long

Island, Antigua (Corliss et al. 1989). Groombridge and Luxmoore (1989) maintain that

about half of the known nesting populations are known or suspected to be in decline; in

particular, "the entire Western Atlantic-Caribbean region is greatly depleted." Prolonged

over-exploitation for the international tortoiseshell trade and the widespread collection of

eggs are implicated in the species' demise.

In the U. S., nesting occurs in the Atlantic (Florida), Caribbean (Virgin Islands, Puerto Rico)

and Pacific (Hawaii, insular Pacific territories), but has been poorly quantified in most areas. 

Observed nesting in Florida is rare at 0-2 nests per annum (1979-1990; Anne Meylan, FL

DNR, pers. comm.). Florida may support a slightly higher level of nesting, but nesting

censuses are accomplished primarily by morning track counts, and hawksbill tracks may not

always be differentiated from those of loggerhead turtles (Caretta caretta). In the USVI,

Zullo (1986) reported 16-49 nests per annum in Virgin Islands National Park, St. John. In

1991, 64 crawls were reported from Cocolobo Beach and 11 more from elsewhere on St.

John (Z. Hillis, in litt., 11 May 1992). As many as 50 nests per annum are reported for St.

Thomas, but these are reported only opportunistically by fishermen and surely represent a

gross underestimate (Ralf Boulon, USVI Div. Fish Wildl., pers. comm.). A summary of

data available data for St. Croix suggests that nesting occurs on nearly 50 beaches, with 200

or more nests laid per year; Sandy Point NWR alone receives 10-20 nests per year and

Isaac's, Jack's, and Coakley Bays may receive 10-30 nests each (Eckert 1992). Data from

Buck Island (1988-1991) reveal 60-99 confirmed nests per year out of 158-240 crawls (Z.

Hillis, in litt., 11 May 1992).

In Puerto Rico and its offshore cays, there are about 275 miles of sandy beaches (Cintron and

Cintron 1987), making systematic surveys of hawksbill nesting difficult. In 1981, 22 nests

were reported on Vieques Island (Pritchard and Stubbs 1982). A decade later, systematic
83

survey of this satellite island was initiated by the Puerto Rico Department of Natural

Resources and the U. S. Department of the Navy; 25 hawksbill nests were recorded in 1991

(Benito Pinto, PR DNR, pers. comm.). Between June 1975 and June 1977, 73 nests were

reported from "the Culebra group" (Carr 1978). In 1985, 23 nests were reported on Culebra

Island (Tony Tucker, pers. comm. in Meylan 1989); one nest was reported in 1987 and 48

between 1989-1991 (B. Pinto, pers. comm.). Over 15 years an average of 160 nests per year

were laid on Mona Island (1974, 1984-1989; Richardson 1990); 159 nests were laid in 1991

(B. Pinto, unpubl. data). Nesting on the mainland of Puerto Rico remains largely

unquantified, but a few beaches (i.e., Piñones, Luquillo, Humacao) have been surveyed since

1985. Between 1988-1991, 16-27 hawksbill nests per year were laid at Playa Humacao

(Manuel Corbet, PR DNR, unpubl. data). Only one nesting attempt has been reported from

Luquillo (B. Pinto, pers. comm., 1992).

In Hawaii, nesting is known to occur on the main islands between July and November,

including the sites of Halape and Apua Point at a remote location in the Hawaii Volcanoes

National Park. The most consistently used sites seem to be at Kamehame Point (Hawaii) and

on a black sand beach at the river mouth of Halawa Valley (Molokai); not more than a dozen

females nest per year on all beaches combined (NMFS 1992). The Palau nesting population

of hawksbills is the largest in Oceania north of the equator; nesting is concentrated on the

small beaches of the Rock Islands between Koror and Peleliu islands (Maragos 1991). This

population is severely stressed by chronic egg poaching and the hunting of turtles for jewelry

and crafts (Maragos 1991). Palauan fishermen are unanimous in their opinion that both

green and hawksbill turtles are "far less abundant" than they were 10-20 years ago (Johannes

1986). An estimated 100 females (hawksbills and green turtles combined) nest per year in

American Samoa; many are harvested and populations have "seriously declined" throughout

the territory (Tuato'o-Bartley et al. 1992). Rose Atoll was designated a National Wildlife

Refuge in 1974 and an unquantified (probably low) level of hawksbill nesting occurs there

(Balazs 1978b). Hawksbill nesting is reportedly of very little significance in Guam or the

Northern Mariana Islands (Pritchard 1982; Groombridge and Luxmoore 1989). No nesting

has been recorded in the unincorporated U.S. Pacific island territories (Eckert 1993).

Excluding the U.S. Pacific where firm data are virtually non-existent, the United States

(Caribbean/Atlantic) probably supports a minimum of 650 nests per annum or, based on

annual average clutch frequency of five nests per female (Richardson et al. 1989), perhaps
84

130 nesting females. It is useful to remember that estimates of population size ultimately

depend not only on regular surveys of potential habitats, but on our ability to distinguish a

successful nesting from an unsuccessful attempt. Crawl effort (total crawls/nests) varies

temporally and geographically. In the Seychelles, 56.2 % of crawls were judged to have

resulted in egg-laying (40.1 % were confirmed nests) in 1973-1974, whereas in 1974-1975,

62.0 % were so judged (57.2 % confirmed) (Diamond 1976). On Campbell Island,

Australia, 77 % of crawls resulted in egg-laying (Limpus et al. 1983). On Mona Island,

about half the crawls observed, on average, result in egg deposition (39 % - 76 %;

Richardson 1990). At Pasture Bay, Antigua, 60% of 1987 activities were confirmed nests,

68 % in 1988 (Corliss et al. 1989). Between 1985 and 1990 on Buck Island, USVI, 42.2 %

to 79.0 % of crawls were judged to be nests, while the proportion later confirmed to have

eggs varied between 13.4 % and 55.6 % per year (Hillis and Mackay 1989, Z. Hillis, pers.

comm., 1992).

Far fewer data are available regarding foraging population size and distribution. In waters

adjacent to the continental United States, hawksbills have been reported in every state from 

Texas to Massachusetts, with the exceptions of Connecticut and Maine (Woodard 1980). 

While opportunistic foraging, especially during the summer months, is likely to occur along

much of the eastern seaboard and northern Gulf of Mexico, the species is described as

"predictable" only along the southeast coast of Florida from Palm Beach south into the

Florida Keys (A. Meylan, pers. comm.). The Florida records are mainly of juveniles and

subadults, although adults are occasionally reported, and the individuals involved are few in

number. In the U. S. Caribbean, foraging is generally associated with insular shelf reefs. 

Preliminary observations from field surveys conducted in the USVI and Puerto Rico indicate

that juveniles are typically found in water <18 m deep, mostly associated with fringing reefs

around the coastline, while adults forage in deeper waters on the shelf (R. Boulon, pers.

comm.). Juveniles are less common than adults and may tend to forage over wider

territories. Juvenile hawksbills tagged in the USVI have been subsequently recovered

elsewhere in the Western Atlantic (Boulon 1989). Studies in progress on the ecology and

population dynamics of immature hawksbills at their Mona Island, Puerto Rico foraging

ground is focusing on the turtles' growth rates, diet, sex ratios, and genetic relationship in

comparison with nearby nesting populations (vanDam and Diez 1994). 
85

In the U.S. Pacific there are no hawksbill sightings off the west coast of the continent

(Stinson 1984, Scott Eckert, Hubbs-SWRI, pers. comm., 1994). The Hawaiian population is

small and only known to occur in coastal waters of the eight main and inhabited islands at the

southeastern end of the 2,450 km-long archipelago (Balazs 1982). During recent aerial

surveys of Guam (October 1989-April 1991), only 13.2 % of 76 turtles sighted were

hawksbills (G. Davis, Guam Div. Aquatic Wildl. Res., in litt., 22 August 1991). In

contrast, 83 % of 29 sightings (1971-1991) around Tutuila (American Samoa) were of

hawksbills (Tuato'o-Bartley et al. 1992). Elsewhere in the U. S. Pacific specific foraging

grounds are known in some cases, such as in Palau (Pritchard 1982, Maragos 1991), but in

general the literature for the Caroline Islands, the Republic of the Marshall Islands, and the

Commonwealth of the Northern Mariana Islands does not fully differentiate between green

turtles and hawksbills (e.g., McCoy 1974, Pritchard 1982, Johannes 1986). There are no

documented sightings in the unincorporated U.S. Pacific territories (Groombridge and

Luxmoore 1989). However, with the exception of Johnston Atoll (see Balazs et al. 1990),

surveys are lacking.

Listing Factors

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

Sea turtles of all species are threatened by the destruction or modification of important

nesting and foraging habitats. The number of suitable nesting beaches worldwide has been

reduced by residential and commercial coastal development (particularly high-density

development), artificial beach-front lighting, uncontrolled recreational use, sand mining,

beach armoring (including stabilization structures such as seawalls, groins, and erosion

barriers), mechanized beach cleaning, vehicle traffic, litter, the removal of stabilizing dune

vegetation, and the introduction of exotic plants. In Barbados, 55.6 % of nests studied were

affected by beach lighting, resulting in up to 100 % of newly emerged hawksbill hatchlings

in some nests crawling inland (Horrocks et al. 1989). In Antigua, wind erosion following

the removal of stabilizing vegetation has rendered significant areas unusable for hawksbill

nesting; raking and beach cleaning procedures further exacerbate erosion (Ryder et al. 1989). 

In Hawaii, resorts, marinas, and even proposed rocket launching facilities (at Ka'u near

South Point on the island of Hawaii) threaten hawksbill habitat (NMFS 1992). General

overviews of habitat alterations and its observed consequences are provided by Bjorndal
86

(1982), Groombridge (1982), Coston-Clements and Hoss (1983), Bacon et al. (1984),

Hopkins and Richardson (1984), Raymond (1984), National Research Council (1990), and

NMFS\FWS (1993).

At Carambola Resort on Davis Beach, St. Croix, 75 hatchlings were recently found dead in

runoff basins after having been attracted by beach-front lights and by accent lighting along

pedestrian pathways. The basins were later covered with fine-meshed screens, non-essential

lights were turned off, and other lights are being replaced with low pressure sodium bulbs to

help alleviate the problem. The problem of photopollution is not an isolated one. Hotels

along the north coast of St. Croix bathe surrounding cove beaches, virtually all important to

hawksbill turtles, in light. On Buck Island, a protected area northeast of St. Croix,

researchers patrolling the beach on moonless nights can see their own shadows silhouetted by

the bright lights of Christiansted five miles away (Z. Hillis, pers. comm., 1992). Near

Frederiksted, artificial lighting leads hatchlings inland and away from the sea (Philbosian

1976). In Puerto Rico, where the lighting problem is also widespread, some cooperation has

been obtained from public beaches (e.g., Seven Seas near Fajardo) and private hotels (e.g.,

Palmas del Mar) with regard to reducing the amount of light shining on nesting beaches at

night (B. Pinto, pers. comm., 1992).

At sea, indiscriminate anchoring, chemical and organic pollution, and sedimentation from

dredging and upland deforestation degrade coral reefs and other potentially important

habitats. In October 1988, the 440-foot cruise ship Windspirit illegally dropped anchor west

of Francis Bay on the north side of St. John in Virgin Islands National Park and Biosphere

Reserve, obliterating 283 square meters of coral reef. In October 1990, the anchor chain of

the 438-foot cruise ship Seabourne Pride uprooted and overturned at least 42 boulders of

living coral (some 3 m across) in 20 m of water off of Caneel Bay, also in the Biosphere

Reserve. Furthermore, the cumulative negative effect of countless smaller boats anchoring in

coral reef areas is considered significant. Vessel groundings are also a persistent problem

and there are many incidents of small boats running aground on shallow reefs off the north

shore of St. John (Caroline Rogers, USNPS, pers. comm.). On 15 February 1985, the

350-foot MV/A. Regina ran aground off the east coast of Mona Island in federally designated

hawksbill Critical Habitat. The wreck spilled diesel oil, extensively damaged the reef,

produced a considerable suspension of sediment, and littered the beaches with oil and debris

(Cintron and Cintron 1987). On 25 August 1985, a hawksbill emerged to nest on Playa
87

Sardinera on Mona Island with oil on her fore and rear flippers, plastron, buttocks, cloaca,

head and throat; she was unsuccessful in her nesting attempt and did not return (Kontos

1985).

The recent reality of repeated oil spills in the U. S. Caribbean indicates the very serious

nature of this threat. In September 1989, following Hurricane Hugo, a 42,000 gallon spill of

#6 fuel oil (heavy crude oil) at the Water and Power Authority facility in Christiansted, St.

Croix, left south coast beaches heavily oiled. Pelican Cove, a hawksbill nesting beach, was

buried under 0.3 m of crude oil. Between March 1991 and March 1992, two more spills,

both outside of U. S. waters, threatened U. S. nesting beaches. The first occurred on 6

March 1991, 13 nm north of Nevis, when the Trinidad-registered barge Vestabella, loaded

with about 560,000 gallons of #6 fuel oil, sank in 600 m of water after a towing cable

snapped; the initial oil slick was more than 30 miles long (Simmonds 1991). According to

The Daily News (30 March 1991), a USVI newspaper, tar balls and tar sheets began

appearing on St. John on 21 March; tar balls washed ashore soon thereafter on St. Thomas,

St. Croix, Culebra, Vieques, and the main island of Puerto Rico. One year later, on 15

March 1992, a pipe ruptured during ship-to-shore pumping of #6 fuel oil to a transfer station

at St. Eustatius Terminal on the west coast of St. Eustatius, Netherlands Antilles. One

hundred barrels of crude oil were released to the sea in a slick that headed northwest out

across the rich fishing grounds of the Saba Bank. Heavy seas broke up the slick before it

entered U. S. waters, but tar balls eventually fouled the coast of Puerto Rico (Z. Hillis, pers.

comm., 1992).

Behavioral experiments indicate that green and loggerhead sea turtles possess limited ability

to avoid oil slicks, and physiological experiments show that the respiration, skin, some

aspects of blood chemistry and composition, and salt gland function of 15-18 month old

loggerheads are significantly affected by exposure to crude oil preweathered for 48 hours

(Vargo et al. 1986). There is some evidence to suggest that hawksbills are also vulnerable

to oil pollution. Hawksbills (predominantly juveniles), were only 2.2 % (34/1551) of the

total sea turtle strandings in Florida between 1980-1984, yet comprised 28.0 % of

petroleum-related strandings. Oil and tar fouling was both external and internal. Chemical

analysis of internal organs provided clear evidence that crude oil from tanker discharge had

been ingested (Vargo et al. 1986). Carr (1987) reported juvenile hawksbills (to 20 cm)

"stranded [in Florida] with tar smeared sargassum"; some individuals had ingested tar. He
88

noted that the Gulf Stream at times carries oil from both European sources and the Gulf of

Mexico into Florida waters. More recently, following the Vestabella barge spill, a hawksbill

soaked in oil was found dead near Guayama on the south coast of Puerto Rico (B. Pinto,

pers. comm., 1992).

2. Over-utilization for Commercial, Recreational, Scientific, and Educational Purposes

Hawksbills were heavily exploited in the United States prior to their federal protection under

the Endangered Species Act (ESA) of 1973, as amended. Pre-ESA statistics pertaining to the

harvest and marketing of sea turtles in Florida and other southeastern states are provided by

Rebel (1974). According to Joe LaPlace (pers. comm. in Eckert 1992), a lifelong resident of

St. Thomas, the intense commercial harvest of hawksbills began in the USVI in the 1920's

when dealers offered to purchase the shells for export. Within just a few years there were

fewer turtles in the water and carcasses littered the beaches. After WWII, with the advent of

plastics and nylon, the lucrative export market died out, but an increasing human population

(residents and tourists) kept the pressure on remaining stocks until protection was conferred

in 1970. Today, despite legal protection and depleted local stocks, a black market persists. 

In St. Croix, meat and eggs are in "high demand" in some areas, with hawksbill and green

turtle eggs selling (in season) for $7-10 per dozen and some "being traded for crack

[cocaine]" (Greg Hughes, USFWS, in litt., 12 May 1992). In November 1990, a man in

possession of 140 hawksbill eggs was arrested at Manchineel Beach, St. Croix, and convicted

in December 1991 after pleading guilty. Another arrest was made in November 1991 after

45 eggs were taken from Sandy Point NWR; the poacher was sentenced to three months in

prison and two years probation.

There is also a significant level of clandestine harvest in Puerto Rico, especially on the main

island and Mona Island, for both meat and eggs. Seven carcasses were found on Mona in

1986 (Kontos 1987) and four more in 1987 (Kontos 1988). In some western areas of Puerto

Rico "hawksbill sausages" are made. The fresh blood is spiced and put in ice trays and the

coagulated blood is later fried in hot oil. The "sausages" are a delicacy widely consumed by

residents and visitors alike (Pinto 1992). Also on the main island, two men were arrested in

August 1987 and their vessel, scuba gear, and 20 pounds of hawksbill meat seized; they

were subsequently fined $2,000 for the take and possession of the meat. In June 1988, one

of these men was again convicted of possession, this time of 28.5 pounds of hawksbill meat;
89

his vessel, car, and fishing equipment were seized and he was sentenced to one year

supervised probation and one month in jail. In June 1989, three men pled guilty to counts of

taking and possessing endangered species based on their August 1988 arrest for possession of

150-200 pounds of green and hawksbill sea turtle meat; forfeiture of a large vessel employed

by the men and all gear is pending (M. Christian, in litt., 31 March 1992). At Cayo

Berberia (a mangrove islet off the south coast of Puerto Rico at Santa Isabel), the remains of

more than 50 slaughtered hawksbills were found during a site inspection in December 1989

(Vicente et al. 1989).

Illegal harvest for local consumption is also reported from Hawaii and the U.S. Pacific

territories. An unquantified number of Hawaiian hawksbills are taken for domestic black

markets with spears, nets, harpoons, grappling hooks, firearms from shore, underwater 'bang

sticks', nooses, and by hand capture (NMFS 1992). In October 1989, a 212-foot purse

seiner docked in Guam was seized after U.S. Department of Commerce officials discovered

three small hawksbills aboard the vessel. A fine of $35,000 was deposited in the Wildlife

Conservation Fund as a precondition for release of the vessel (Government of Guam 1990). 

In recent interviews in American Samoa, respondents in 24 % of the villages reported that

turtles (hawksbills and/or green turtles) no longer occurred on their village beaches that

historically had nesting activity. During these interviews it was learned that if a sea turtle of

any age is encountered, the likelihood of it being harvested is high (Tuato'o-Bartley et al.

1992). Similarly, the take of hawksbills in Palau is unrelenting and residents concede to

noticeable declines in recent decades (Johannes 1986, Maragos 1991). Egg theft is also

widespread in the U. S. Pacific, except perhaps in Hawaii where hawksbill nesting is

relatively rare. In Palau, egg theft claims >75 % of all nests (Maragos 1991). In American

Samoa, villagers will collect eggs for consumption "whenever a nest is found"

(Tuato'o-Bartley et al. 1992).

Continuing exploitation in the U.S. is only a small part of a much larger reality. Domestic

harvest, often illegal if not covert, is a persistent problem for hawksbills throughout their

range. In the Caribbean, hawksbill shell fetched US$ 110-130/kg in 1980 and a fishermen

could earn $200 or more for a single turtle (Carr and Meylan 1980). The harvest has not

abated with the increasing scarcity of the resource; rather, hawksbills are taken

opportunistically by divers targeting lobster and other high priced items. Thus it is likely that

individuals will continue to be harvested long after the species would otherwise be
90

commercially extinct. For many coastal Caribbean peoples, the capture and sale of a single

hawksbill can be equivalent to one or two weeks' wages. Nietschmann (1981) estimated that

1,000-1,200 hawksbills were removed annually from eastern Nicaraguan waters by nets,

harpoons, and hooprings [nooses]; in the decade between 1969 and 1978, the price for shell

rose 600 %. Heavy pressure is also exerted on eggs. In Mexico, poachers take more than

half of the nests in Campeche and Yucatan, and 60-70 % of those in Isla Holbox; >50 % of

the eggs (and turtles) arriving at the Pacific coast are taken to regional or national markets

(PESCA 1990). [N.B. The extent to which the recent ban on sea turtle harvest in Mexico

(Aridjis 1990) will ultimately eliminate this market remains unknown]. The scenario is

repeated throughout the world (Carr et al. 1982, Meylan 1984b, Groombridge and Luxmoore

1989, Eckert 1993).

Above and beyond domestic harvest is the ongoing international commerce in hawksbill shell

(tortoiseshell, or 'bekko'), a phenomenon widely touted as the single most significant factor

endangering hawksbill populations around the world. To meet the demands of at least four

separate native industries, Japan has conducted the world's largest international trade in sea

turtles and sea turtle products, focusing not only on the hawksbill, but on green turtles and

olive ridleys (Lepidochelys olivacea) as well (Milliken and Tokunaga 1987). Japanese

imports of raw bekko between 1970 and 1989 totalled 713,850 kg, representing >670,000

turtles; more than half the imports originated from the Caribbean and Latin America

(Milliken and Tokunaga 1987, updated by Greenpeace to 1989). Between 1970 and 1987, a

total of 675,247 kg of stuffed hawksbills representing >587,000 turtles was imported

(Greenpeace 1989). Milliken and Tokunaga (1987) note that in order to maintain these levels

of importation, the annual slaughter of at least 28,000 hawksbills is required. Between 1970

and June 1989, Japan imported 368,318 kg of bekko from the Wider Caribbean alone, the

equivalent of more than a quarter million turtles; in 1988, Japan imported from the Wider

Caribbean the tortoiseshell from nearly 12,000 adult hawksbills (Canin 1989).

In spite of full domestic protection for sea turtles and a ban on their importation into the U.

S., sea turtle products, many of them hawksbill in origin, are not only the most commonly

confiscated products by U.S. Customs at our borders (Andrea Gaski, TRAFFIC(USA), pers.

comm.), but these confiscations are on the increase (O'Connell 1990). And lest we believe

that these products are only purchased in foreign countries, about $150 worth of tortoiseshell

jewelry (allegedly imported from Jamaica) was recently confiscated from a store at the St.
91

Croix airport (G. Hughes, in litt., 12 May 1992). In Puerto Rico, tortoiseshell jewelry

valued at approximately $200 was recently seized by NMFS from a gift shop in La Parguera;

a federal investigation determined that the jewelry had been imported by a distributor from

Colombia and purchased at a jewelry show in San Juan. The items have been forfeited to the

U. S. Government (M. Christian, in litt., 31 March 1992). In 1984, investigations in the

USVI resulted in seizure of two commercial shipments of sea turtle jewelry (valued at $500);

in 1986, the seizure of 43 pieces of jewelry; in 1988, forfeiture of $150 in tortoiseshell

jewely boxes (M. Christian, pers. comm.). It is abundantly clear that the Caribbean tourist

industry is fueling the continued exploitation of hawksbills throughout the region for jewelry,

trinkets, stuffed turtle wall hangings, and polished shells.

The negative effects of utilizing hawksbill turtles for scientific purposes are not likely to be

significant. Contemporary hawksbill research consists primarily of studies of the

distribution, abundance, and behavior of nesting and foraging turtles, nest protection,

measuring and tagging, biotelemetry, non-intrusive physiological measurements, and other

pursuits that are not detrimental to the animals involved. In the United States, the use or take

of hawksbills for scientific or educational purposes is controlled by a permit system designed

to protect endangered and threatened species.

3. Disease or Predation

There are no data on the extent to which disease affects hawksbill turtles in the wild. 

External parasites and commensals representing five phyla and 17 families of plants, and two

phyla (Arthropoda and Mollusca) and six families of animals have been recorded; dozens of

species of internal parasites, mostly trematodes, are also documented (see Witzell 1983).

Ghost crabs (Ocypode quadratus, O. ceratophthalmus), land crabs (Cardisoma guanhumi),

monitor lizards (Varanus indicus), Ameiva lizards, night herons (Nyctanassa violacea), and

barn owls (Tyto alba) are known to prey on eggs and/or hatchlings on the beach (Diamond

1976, Small 1982, Limpus et al. 1983, Márquez 1990, Pinto 1992); the black rat (Rattus

rattus) is a suspected predator (Hillis and Mackay 1989). Exotic and feral species, such as

mongooses (Herpestes auropunctatus), wild pigs (Sus scrofa), and domestic dogs are also a

significant threat in some areas. On St. John (USVI) mongooses destroyed 23 % of the total

1980-1981 egg production, and in 1980 dogs destroyed an additional 19 %; wire enclosures
92

subsequently helped to mitigate this problem (Small 1982). On Buck Island, north of St.

Croix, mongooses destroyed virtually all nests laid (at least 45) in 1981 (Small 1982) but

subsequent eradication programs reduced the mongoose population and in 1988 only 8 % of

nests there were lost to this predator (Hillis and Mackay 1989). On Mona Island in 1987, 35

of 71 (49 %) nests were lost to feral pigs (Kontos 1988). Richardson (1990) reports 50-100

% annual nest loss to pigs on some of Mona's most important hawksbill beaches. After

galvanized fences were installed in June 1990, pig predation ceased (B. Pinto, pers. comm.,

1992).

At sea, hatchlings, juveniles, and adults fall prey to oceanic predators, including tiger sharks

(Galeocerdo arcticus [=cuvier]), groupers (Epinephelus itajara), and crocodiles (Crocodylus

porosus) (summarized by Witzell 1983). Hatchlings were attacked by a black-tipped reef

shark (Carcharhinus spallanzani) in Samoa (Witzell and Banner 1980). In 1982 the carapace

scutes from an estimated 28 kg hawksbill were found in the stomach of a 4 m tiger shark

captured off St. Thomas, USVI (Boulon 1984).

4. Inadequacy of Existing Regulatory Mechanisms

In the United States, both national and international laws protect hawksbill turtles from

harassment, harvest, and commerce. According to the Code of Federal Regulations (50 CFR

17.11), the hawksbill is listed as Endangered throughout its entire range under the U. S.

Endangered Species Act (ESA) of 1973, as amended (35 FR 8495; June 2, 1970). Similarly,

the species is classified as Endangered in the International Union for Conservation of Nature

and Natural Resources' (IUCN) Red Data Book, where taxa so classified are considered to be

"in danger of extinction and whose survival is unlikely if the causal factors continue

operating" (Groombridge 1982). Hawksbills are also included on Appendix I of the

Convention on International Trade in Endangered Species of Wild Fauna and Flora (CITES). 

The U. S. ratified CITES in 1974.

All federal agencies must ensure that actions authorized, funded, or carried out by them do

not result in the destruction or adverse modification of habitat designated as Critical Habitat

for marine turtles pursuant to Section 7 of the ESA. Critical Habitat was proposed (Federal

Register, 24 May 1978; 43 FR 22224-22225) for hawksbill turtles in Puerto Rico by the U.

S. Fish and Wildlife Service and subsequently designated (effective 26 July 1982) as follows:
93

(a) Mona Island. All areas of beachfront on the west, south, and east sides of the island from

mean high tide (MHT) inland to a point 0.1 miles from shore, including all 7.2 km of

beaches. (b) Culebra Island. All areas of beachfront on the north shore of the island from

MHT inland to a point 0.1 miles from shore, including the following beaches: Playas Blanca,

Flamenco, Resaca, Brava, and Larga. (c) Cayo Norte. South beach, from MHT inland to a

point 0.1 miles from shore. (d) Culebrita Island. All beachfront areas on the southwest,

east, and northwest shores of the island from MHT inland to a point 0.1 miles from shore

(Dodd 1978).

Critical Habitat for this species was also proposed elsewhere in Puerto Rico (Vieques Island,

south beaches of Culebra Island), as well as in the USVI (portions of St. Croix, St. John, and

St. Thomas; beaches of Buck Island National Monument) (Dodd 1978), but these proposals

have not been acted upon (Ken Dodd, USFWS, pers. comm., 1992). Critical Habitat has not

been designated for hawksbills (or for any other species of sea turtle) in the U. S. Pacific. A

proposed rule was prepared in 1980 to designate Critical Habitat for the green turtle in

Hawaii, Republic of the Marshall Islands, Federated States of Micronesia, and the Republic

of Palau, but this was never approved by the U. S. Fish and Wildlife Service (K. Dodd, 

pers. comm.). In addition, several beaches were considered by Dodd (1978) to be potential

candidates for Critical Habitat for the hawksbill turtle in the U. S. Pacific, including

American Samoa (Tutuila Island, Rose Atoll, Swains Island) and portions of the Republic of

Palau, Truk District, and Lower Mortlocks. The protection of important habitat is crucial to

the long-term survival of endangered and threatened sea turtles. In light of the fact that no

progress whatsoever has been made during the last decade with regard to the designation of

Critical Habitat for hawksbill sea turtles under the ESA, the evaluation and designation of

such habitat should be a high priority.

Broad legal protection has been effective in eliminating turtle and egg poaching in the

continental USA, but not so in Puerto Rico, the USVI, or the U.S. Pacific. In American

Samoa, for example, most people are unaware that it is illegal to take turtles and enforcement

of existing laws protecting turtles is lacking (Tuato'o-Bartley et al. 1992). In Puerto Rico the

theft of eggs is nearly ubiquitous and inadequate local law enforcement is a widely

recognized problem. Meat is available for the asking on the black market and, in spite of

restrictions, some businesses continue to exhibit and sell items fashioned from hawksbill shell

(Pinto 1992). Poaching also continues in the USVI. The Virgin Islands Department of
94

Planning and Natural Resources' Law Enforcement Division lacks sufficient funds and

manpower to fully cover the beaches and phones (at least on St. Croix) are unmanned on

weekends when most people observe violations (Z. Hillis, in litt., 11 May 1992). The

enforcement problem is only exacerbated by the diffuse distribution of this species, by some

nesting and foraging habitats being coincident with relatively remote territories, such as

Mona Island in the Caribbean and the varied archipelagoes of the Pacific, and by the

inexperience or inattention of Customs officers that permits products derived from

endangered species to pass into and out of U.S. Caribbean and Pacific islands. Vigilant

enforcement and more visible public awareness campaigns are clearly needed.

Elsewhere in the world, despite the protection conferred by CITES, harvest continues for

several reasons. First, while CITES prohibits trade in sea turtle products amongst nations, it

does not restrict domestic harvest or commerce. Second, while 126 nations have ratified

CITES (WWF/IUCN 1994), some have exercised their right to take exemption to treaty

provisions as they pertain to sea turtles. When Japan ratified CITES in 1980, it placed

reservations on the hawksbill, green, and olive ridley, effectively exempting itself from the

ban on their trade (Greenpeace 1989). [N.B. Japan has since withdrawn all reservations on

sea turtle species]. Similarly, Cuba took (and still maintains) reservations on hawksbills and

green turtles when it acceded to the Convention in 1990. Third, some countries, such as

Indonesia which ratified CITES in 1979, ignore their obligations as CITES parties and openly

trade in Appendix I species. According to Japanese Customs Statistics, stuffed hawksbills

from Indonesia accounted for nearly half of all "worked bekko" imports between 1979-1986

(208,586 kg out of 440,914 total kg; Milliken and Tokunaga 1987). Other CITES parties

clandestinely participate in the trade by re-routing their shipments through a non-CITES

country. Finally, some important hawksbill suppliers do not belong to CITES at all. 

Between 1970 and June 1989, Haiti, a non-CITES country, exported bekko representing

>21,000 hawksbills to Japan (Greenpeace 1989).

5. Other Natural or Manmade Factors

On the beach, eggs can be lost to flood waters, inundation, and erosion (Small 1982, Hillis

and Mackay 1989, Horrocks et al. 1989, Hillis et al. 1990). In 1989, 18 % of hawksbill

nests at Buck Island Reef National Monument (St. Croix) were lost to flood tides and erosion

effected by Hurricane Hugo. During 1990, there was a three-fold increase in the false crawl:
95

nest ratio on Buck Island due to steep erosion berms, downed trees, and exposed root masses

which remained as Hugo's legacy (Hillis 1991). The problem persisted in 1991, as fallen

vegetation continued to interfere with nesting attempts in the beach forest (Hillis 1992). In

Hawaii, an entire nesting beach at Harry K. Brown Beach Park in Kalapana was recently lost

to a lava flow from the Kilauea volcano (NMFS 1992). 

In nearshore waters, hawksbills are periodically captured in the cooling water intakes of

industrial facilities, such as Florida Power and Light Company's St. Lucie Power Plant on

Hutchinson Island. Between March 1976 (when the St. Lucie Plant opened) and November

1988, six hawksbill captures occurred in the Plant (Ernest et al. 1989). An additional three

were recorded as of 1 June 1992, the most recent being a female weighing 191 lb [87 kg] and

captured on 1 March 1991; all turtles were released unharmed (Eric Martin, Applied

Biology, pers. comm.). In offshore waters, such as in the Gulf of Mexico, oil and gas

structures act as "artificial reef habitats" and the explosive removal of these structures may

result in the death of turtles (Gitschlag and Renaud 1989). 

Incidental capture and entanglement at sea are unquantified and potentially important sources

of mortality. The Sea Turtle Stranding and Salvage Network (Atlantic and Gulf coasts of the

U. S.) reported 27 stranded hawksbills in 1988, 35 in 1989, 61 in 1990 (when an unusually

large number of post-hatchlings washed ashore in Texas), 33 in 1991 and 47 in 1992 (Wendy

Teas, NMFS, pers. comm.). While hawksbills typically represent <1.5 % of total sea turtle

strandings (Schroeder and Warner 1988, Teas and Martinez 1989), the problem is a

persistent and worrisome one. Gill nets, longlines, and shrimp trawls all take turtles in Gulf

of Mexico waters (Hildebrand 1987, NRC 1990) and hawksbills strand on the Texas coast

during virtually all months of the year (Schroeder and Warner 1988, Amos 1989, Teas and

Martinez 1989, 1992).

Of the 25 sea turtles found entangled on the Texas coast during 1986 and 1987, 24% were

juvenile hawksbills (mean 24.3 cm CCL) (Plotkin and Amos 1988). Hawksbills

(predominantly juveniles) have been reported entangled in monofilament gill nets, "fish

nets", fishing line, and synthetic rope; in most cases flippers were lost as a result, and in one

case an animal was recovered with a piece of plastic onion bag entangled around its neck

(Balazs 1985). Balazs (1985) summarized published incidents of the ingestion of marine
96

debris by hawksbills and reported that 88.9 % of the articles recovered were plastic bags,

plastic and styrofoam particles, and tar; 90.9 % of the individuals involved were juveniles.

The specific effects of marine debris and pollution, including oil and pesticides, on

hawksbills, their eggs, and their prey have yet to be determined. The ability of hawksbill

populations to recover will ultimately depend not only on protection from direct harvest and

the conservation of important nesting beaches, but also on clean oceans and a healthy littoral

zone, especially as this pertains to coral reefs and other foraging and resting grounds.

Conclusion

The hawksbill is listed as Endangered by the U. S. Endangered Species Act of 1973, as

amended (USFWS 1989) and the IUCN Red Data Book of endangered and threatened species

(Groombridge 1982). It is also included on Appendix I of CITES. With the exception of the

Kemp's ridley (Lepidochelys kempii), the hawksbill is considered by many to be the most

endangered of all the marine turtles. Of 65 geopolitical units where estimates of relative

hawksbill nesting density exist, 38 of them have hawksbill populations that are suspected or

known to be in decline and an additional 18 have experienced "well-substantiated declines"

(Groombridge and Luxmoore 1989). Hawksbills are a shared resource in the Western

Atlantic, meaning that the status of U.S. populations is tied to factors operating outside our

jurisdiction. Small juveniles in Texas waters are likely to originate from the nesting beaches

on Mexico's Yucatan Peninsula (Hildebrand 1987, Amos 1989) and perhaps other Caribbean

beaches as well. Tagging studies have shown that juveniles tagged in St. Thomas are

frequently recaptured outside of USVI waters, including locales as distant as St. Lucia 650

km to the southeast (Boulon 1989). The same is true in the U.S. Pacific, where hawksbill

stocks are likely to be shared in common amongst the islands of the central and western

Pacific. 

The greatest threat to the hawksbill worldwide is commerce in raw and worked shell

('bekko', or tortoiseshell). The advent of plastics in the 1940s seemed to promise a more

secure future for the species, but this has not been so. Until recently, tens of thousands of

hawksbills were sacrificed each year to meet the demand for jewelry, ornamentation, and

whole stuffed turtles (Milliken and Tokunaga 1987). Hawksbills are mobile, and thus

international trade threatens not only the stocks of exporting countries but all hawksbill
97

populations, including those of the U.S.. Japan's 1988 bekko imports from Jamaica, Haiti,

and Cuba represented some 13,383 hawksbills; it is extremely unlikely that this volume could

have originated solely from local waters (Greenpeace 1989).

In a move destined to benefit hawksbill populations in the U.S. and throughout the world,

Japan announced on 19 June 1991 its intention to end its international trade in bekko. The

decision followed a determination by the U.S. Government that Japan was "undermining the

effectiveness of international programs for the conservation of sea turtles", a determination

that, under the authority of the Pelly Amendment to the Fisherman's Protective Act of 1967,

carried the threat of a partial or complete embargo on fish and wildlife products imported

into the U.S. from Japan (Donnelly 1991). Japan kept its promise to ban all imports of turtle

shell by 31 December 1992, and subsequently removed its Eretmochelys reservation under

CITES in 1994. 

There is little doubt that U.S. Caribbean nesting colonies have been severely depleted during

the twentieth century. The same can be said for the Pacific, where the advent of modern

fishing gear has made harvesting turtles more efficient, and the means to travel with

motorized vessels safely over great distances has enabled hunters to gather mating and nesting

turtles at previously inaccessible breeding sites. Today the illegal domestic harvest of eggs

and turtles continues in the United States, especially in Caribbean and Pacific island

territories. Law enforcement, as well as conservation and management efforts, are hindered

by diffuse nesting distributions and the remoteness of some rookeries. It is not easy to

determine whether remaining populations are stable, increasing, or declining. Systematic

surveys of selected index beaches are essential in order to evaluate the present status of

remnant stocks. Despite predictions that hawksbills would be extinct on Mona Island before

the year 2000 (Kontos 1988), there was cautious optimism after seven years of study (1974,

1984-1989) that annual fluctuations in nesting activity were not projecting a long-term

decline (Richardson 1990).

While some areas under U.S. jurisdiction, notably Mona Island, St. John, and Buck Island in

the Caribbean and Palau in the Pacific, have the potential to serve as important refugia for

the species, the degradation and loss of habitat continues in the United States and elsewhere. 

Coastal development, including beach armoring, traffic, and artificial lighting is proceeding

at a rapid pace. Coral reefs, vital as food and shelter for hawksbills, are shattered by
98

indiscriminate anchoring, broken by divers, fouled by pollutants and sedimentation, stressed

by disease, and smothered by spilled oil. At-sea entanglement in active and abandoned

fishing gear, the ingestion of marine debris, and petroleum fouling are unquantified and

potentially serious concerns to the turtles themselves. In summary, depleted U.S.

populations are not currently considered to be declining, but neither are there indications of

recovery despite more than a decade of legal protection. In addition, habitat destruction and

clear evidence of over-exploitation continue on a global scale. The best available commercial

and scientific data indicate that the hawksbill turtle (Eretmochelys imbricata) should remain

listed as Endangered throughout its range, pursuant to Section 4 of the Endangered Species

Act of 1973, as amended.
99

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109

Kemp's Ridley Sea Turtle, Lepidochelys kempii 

Prepared by Michael Weber

Biological Background

Kemp's ridley sea turtle has also been known as the Atlantic ridley, Mexican ridley, bastard

turtle, or grey loggerhead in English; as tortuga lora or tortuga cotorra in Spanish; and as

tortue batarde in French. The specific name derives from Mr. Richard Kemp of Key West,

Florida, who provided the first type to Garman, who then described the species. The name

"ridley" replaced "bastard turtle" several decades ago, reflecting common usage along the

U.S. Gulf of Mexico coast where the animal was once common. The derivation of the word

ridley for sea turtles in the Genus Lepidochelys is unclear.

Kemp's ridley exhibits several distinguishing features: two pairs of prefrontal scales,

five (very rarely more) pairs of costal scutes, a pore in each inframarginal scute in the

bridge, paddle-shaped forelimbs, each with one claw, and a shell, in adults, nearly as wide as

long (Mager 1985, Pritchard 1989, Ross et al. 1989). Kemp's ridley is one of the smallest

of all marine turtles. Adult females measure 58 - 80 cm in straight carapace length (SCL),

and weigh 40 - 50 kg (Ross et al. 1989). The tail of an adult male extends well beyond the

carapace; the tail of an adult female does not (NRC 1990). Juvenile and hatchling male and

female Kemp's ridleys are morphologically indistinguishable. Hatchlings are roughly 4.1 cm

in carapace length. Compared to the olive ridley (Lepidochelys olivacea), Kemp's ridley is

similar in size and general conformation, but has smaller orbits in a broader head with more

massive, powerful jaws (Pritchard 1989, NRC 1990). The carapace coloration of Kemp's

ridley appears light olive to light gray in adults; charcoal gray in juveniles, and dark gray to

black in hatchlings. In juvenile Kemp's ridleys, the carapace can be wider than it is long

(Pritchard 1989). The plastron or ventral shell appears white or cream-colored in adults and

juveniles, and usually the same color as the dark gray carapace or a slightly lighter shade in

hatchlings.

It is unknown for certain how long Kemp's ridleys live, or how long it takes them to reach

sexual maturity in the wild. Recent estimates using skeletochronological data fitted to a Von

Bertalanffy growth curve suggest that female Kemp's ridleys may attain the size of a sexually
110

mature individual within 11 - 12 years (Zug 1990). In captivity and on a steady, high protein

diet, Kemp's ridleys as young as five to seven years have produced eggs and hatchlings

(Wood and Wood 1988). 

Kemp's ridleys are distributed primarily in the Gulf of Mexico and western Atlantic Ocean. 

Kemp's ridleys may also occur in the eastern Atlantic Ocean, however most documented

occurrences have been juvenile individuals (Bleakney 1965, Brongersma 1972, Pritchard and

Marquez 1973). Unlike most sea turtles, Kemp's ridley has a very restricted nesting range. 

The principal Kemp's ridley nesting beach is located near Rancho Nuevo, Tamaulipas,

Mexico where 95 % of the nests are laid along 60 km of beach. Hurricane Gilbert scoured

the beach in 1988, removing the foredune over a long stretch of beach and leaving a mixture

of gravel, sand, and rock rubble. In 1989, females apparently expanded their nesting up to

30 km northward of the traditional nesting beach (NRC 1990).

Nesting also occurs sporadically at other beaches in Mexico and in Texas (Ross et al. 1989),

and typically involves only one or very few individual turtles. There have been a few

instances of Kemp's ridleys nesting on beaches outside their traditional range in recent years. 

Sightings of female Kemp's ridleys emerging on beaches, their tracks, and\or their eggs and

hatchlings have been documented on Florida's Gulf and Atlantic coasts (Meylan et al. 1990)

and on the Atlantic coast of North Carolina (T.A. Conant unpublished data cited in Bowen et

al. 1994) and South Carolina (Anonymous 1992).

Unlike sea turtles of other genera, Kemp's ridleys emerge synchronously during the day to

nest in aggregations called "arribadas" meaning "arrival" in Spanish. Strong onshore winds

appear to stimulate nesting emergence and to concentrate nesting effort in a narrower area

than on calm days (Pritchard 1989). More than half of the adult females nest every year

between April and mid-August, while the remainder may or may not skip certain years (NRC

1990). Kemp's ridleys lay on average 3.075 clutches per season (Rostal 1991), at 20 - 28 day

intervals, with an average of 103 eggs per clutch. The nesting process takes less than an

hour (Chavez 1969). 

After a 48 - 65 day incubation period, viable eggs hatch and the hatchlings emerge about

dawn (Mager 1985). After leaving their nests, hatchlings swim determinedly at the surface,

eventually encountering overwhelming currents, such as the Loop Current, which carry them
111

elsewhere in the Gulf of Mexico or into Atlantic waters where they presumably inhabit

Sargassum drift lines, convergences, eddies, and rings where they feed at the surface (Carr

1986, 1987) until reaching a carapace length of about 20 cm (NRC 1990). Post-pelagic stage

Kemp's ridleys measuring about 20 - 25 cm carapace length become benthic feeders when

they enter inshore and nearshore waters, particularly areas where there are seagrass beds or

mud bottoms favored by crabs, from Long Island Sound to the northern Gulf of Mexico. 

Concentrations of juvenile ridleys measuring 20 - 58 cm carapace length are documented for

Cape Cod Bay, Long Island Sound, Chesapeake Bay, and bays and sounds of North

Carolina, northwest Florida, western Louisiana, and along the Texas Gulf coast (Keinath et

al. 1987, Ross et al. 1989, Morreale and Standora 1991, Rudloe et al. 1991, Manzella and

Williams 1992, Kenyon et al. 1994, Landry et al. 1994, Epperly et al. 1995). 

Along the U.S. Atlantic coast, juvenile ridleys inhabit bays and sounds from Spring until Fall

when most turtles depart inshore areas and move south as water temperatures decline

(Keinath et al. 1987, Byles 1988, Morreale and Standora 1993). For reasons unknown, some

Kemp's ridleys remain in inshore waters during Winter and are later found cold-stunned after

water temperatures decrease to near freezing temperatures (Morreale and Standora 1993). 

Juvenile Kemp's ridleys in the Gulf of Mexico occupy inshore waters and nearshore waters

from shore to approximately 50 m in depth (Rudloe et al. 1991, Shaver 1991, Renaud 1993). 

Juvenile Kemp's ridleys move from shallower to deeper waters as temperatures cool during

Winter (Henwood and Ogren 1987). 

Within the Gulf of Mexico, adult kemp's ridleys occupy nearshore waters from shore to

approximately 50 m in depth (Byles 1988, Shaver 1991). Post-nesting adults tracked by

satellite from Rancho Nuevo, Mexico migrated along narrow corridors in coastal waters

generally less than 50 m deep and swam to feeding areas where they established

circumscribed ranges (Byles 1988). These female ridleys moved as far south as Cabo

Catoche at the northeast point of the Yucatan Peninsula, where the Gulf Loop Current may

have discouraged further movement, and as far north as northwestern Lousiana (Byles 1988). 

Nothing is know about the migrations of adult male Kemp's ridleys. 

When adult Kemp's ridleys are not migrating to or aggregated near their principal nesting

beach, they inhabit crab-rich waters, such as those close to the Mississippi River Delta and

Laguna del Carmen in the Bay of Campeche (Pritchard 1989, NRC 1990). Adult Kemp's
112

ridleys are carnivorous benthic feeders, preferring crabs, but also occasionally eating

molluscs, fish, shrimp, and vegetation (Mortimer 1982, Lutcavage and Musick 1985, Shaver

1991, Burke et al. 1993, Werner and Landry 1994). The stomach contents of stranded

Kemp's ridleys often include fish remains. Since Kemp's ridleys are not likely able to

capture live fish, these stranded animals were probably feeding on dead fish, in many cases

those discarded by shrimp trawlers (NRC 1990, Shaver 1991).

Both eggs and hatchlings are vulnerable to a wide array of predators, including coyotes,

ghost crabs, vultures, caracaras, and hawks. Marine predators of hatchlings include jackfish,

red drum, sharks and humans (NRC 1990). Juveniles and adults are vulnerable only to

larger predators such as sharks and humans.

Population Size

The best index for population size of Kemp's ridley, like other sea turtles, is the total number

of nests and nesting females found on nesting beaches (NRC 1990). Based on a 1947 film by

amateur naturalist Andres Herrera of Tampico, Mexico, an estimated 42,000 Kemp's ridleys

nested at Rancho Nuevo on June 18, 1947 (Hildebrand 1982). Since that time, the nesting

population has drastically declined. By 1966, when the Mexican government initiated

protection of the nesting beach, the largest arribadas included no more than 2500 females.

Assuming that each nesting female lays 3.075 clutches per season (Rostal 1991), the nesting

population is now estimated to be 509 females (Table 1), or a little more than one percent of

the nesting population in 1947. 

There is no quantitive information on the abundance of adult males and only a little on

juveniles. Ogren (1989) has observed that juvenile Kemp's ridleys are now regularly

encountered in the northern Gulf of Mexico, whereas surveys in the 1950s had encountered

none. 

Listing Factors

1. Present or Threatened Destruction, Modification or Curtailment of Habitat or Range
113

In 1977, the Government of Mexico designated Rancho Nuevo a Natural Reserve. Although

development within the reserve remains minimal, land clearing has been intense. There are

two additional causes for concern that incompatible development may be in the offing. First,

a local group of townspeople, who formed a fishing cooperative, constructed several concrete

buildings in the reserve adjacent to the Barra Coma dune and just north of the turtle camp. 

The cooperative caused minor environmental problems, such as small spills of oil and gas on

the ground and careless discard of garbage. Concern about the cooperative has decreased,

since the townspeople, inexperienced as fishermen, have not succeeded economically (R.

Byles pers. comm.). Second, a road through the reserve from Barra Coma to Barra del

Tordo has been proposed. Although construction of the road has been delayed and direct

impacts on the nesting beach are unlikely, the proposal suggests that pressure to develop the

beach is building.

Point and non-point source discharges of agricultural and industrial chemicals, petroleum

products, and domestic sewage may have indirect effects on Kemp's ridleys by reducing food

sources through degradation of the habitat of prey species (such as sea grass beds and blue

crabs), and direct effects such as reduction in health and fitness of individual animals that

may be manifested by mortality or the disruption of physiological functions. Many of the

347 industrial discharges in the U.S. Gulf of Mexico are found along the eastern

Texas\western Louisiana coast that is favored by juvenile ridleys (OTA 1987). Similarly,

agricultural runoff of nutrients and pesticides, urban runoff of oil, heavy metals, and other

substances, and discharge of nutrients from the Mississippi and other rivers may have led to

the loss of prey nursery areas. Levels of discharges for much of the Gulf, particularly

Mexico, are unknown. Direct effects of these pollutants on Kemp's ridleys have not been

thoroughly investigated.

Besides possible indirect effects upon sea turtles of heavy metals and other toxic substances in

muds and waters discharged in the course of routine operation of exploratory and production

oil and gas platforms, Kemp's ridleys may be exposed to oil released in catastrophic spills

from oil rigs and tankers. Evidence suggests that sea turtles do not avoid oil slicks and can

die by coating and ingestion of oil (Fontaine 1989). 

Chronic release of small amounts of oil from oil rigs and from commercial oil tankers and

freighters emptying their bilges creates tarballs that Kemp's ridleys may confuse with food. 
114

Tarballs may cling to the mouth of a turtle preventing it from eating; if ingested, a tarball

may clog the esophagus. The toxic components of a tarball can also have fatal consequences.

Nearshore habitats important to Kemp's ridleys are regularly exposed to routine and

catastrophic release of oil and other hazardous materials. The U.S. Gulf of Mexico includes

six of the top ten ports in the amount of crude oil and refined products handled and in the

number of barge movements (Townsend 1990). Mexico's largest offshore oil field, located

in the Bay of Campeche, has already been the source of the major blow-out and spill from

the Ixtoc I well, and is certainly the source of routine discharges.

Oceanographic features such as convergences and currents concentrate discarded debris,

making the Gulf of Mexico particularly hazardous to Kemp's ridleys (O'Hara and Debenham

1989). Debris may affect sea turtles in several ways. Pieces of netting and plastic bags, for

instance, may entangle sea turtles, making it difficult for them to swim, feed, digest, or

evade predation. Sea turtles also ingest both large and small pieces of debris, sometimes

causing reduced nutrient absorption, intestinal damage, or possibly a false feeling of satiation

(Ross et al. 1989). Ingested debris is often present, but cannot always be implicated as a

cause of death in necropsied animals (Plotkin and Amos 1988).

There is little information on the impacts of dredging and filling on Kemp's ridleys in the

Gulf of Mexico. Past dredging of channels has contributed to the loss of coastal wetlands,

particularly in Louisiana, that juvenile ridleys favor. The loss of similar developmental

habitat has not been quantified for the Atlantic coast.

2. Overutilization for Commercial, Recreational, Scientific and Educational Purposes 

Taking of eggs from nests and from slaughtered nesting females together with incidental

drowning in shrimp trawls and other fishing gear led to the historical decline of the nesting

population. Both Mexican and U.S. law now prohibit intentional taking of Kemp's ridley sea

turtles and their eggs. Working cooperatively, the two governments have all but eliminated

poaching on the nesting beach (Ross et al. 1989). Incidental capture of Kemp's ridleys in the

U.S. shrimp trawl fishery which once killed more Kemp's ridley than all other human

activities combined (NRC 1990), continues to threaten the existance of this species despite

federal regulations requiring the use of TEDs in all areas, at all times.
115

3. Disease or Predation

Since the cooperative Mexican \ U.S. recovery program began in 1978, predation on nests

has been held to no more than five percent each year. Little is known about mortality caused

by disease or parasites, except for animals reared in captivity (Leong et al. 1989).

4. Inadequacy of Existing Regulatory Mechanisms

 

Deliberate taking of Kemp's ridleys was prohibited by law in the United States and Mexico

in 1973 (Marquez 1989). Shrimp trawling in the vicinity of the nesting beach at Rancho

Nuevo is also prohibited, but enforcement of this prohibition should be strengthened. The

U.S. Endangered Species Act's prohibition on taking of any kind has been vigorously

enforced, with the exception of incidental capture in shrimp trawls. Federal regulations now

require the use of Turtle Excluder Devices (TED) on all shrimp vessels in the U.S. range of

the Kemp's ridley. Enforcement of TED regulations in recent years has not been sufficient

enough to keep mortality levels at a minimum. Several hundred juvenile Kemp's ridleys

recently stranded along the Texas and Louisiana Gulf coast (Shaver 1994), presumably a

result of incorrect installation of TEDs in some shrimp trawl nets. Reducing mortality

among juvenile Kemp's ridleys is the key to reversing the decline in numbers of nesting

females (NRC 1990). 

5. Other Natural of Manmade Factors 

Increases in the number of pleasure boats in nearshore and estuarine areas are no doubt

leading to increased numbers of collisions with sea turtles. The level of mortality from this

human activity is unknown, however it is probably small (NRC 1990).

There is evidence that Kemp's ridleys may be killed in the removal of offshore oil rigs

through explosives (Klima et al. 1988). The level of mortality from this activity is unknown,

but is probably small. Measures to reduce this source of mortality have been implemented

(Richardson 1989).

Kemp's ridleys are also captured in small numbers in other fisheries, such as the Long Island

Sound and Chesapeake Bay pound net fisheries (NRC 1990). Powerplant intake pipes entrap
116

and drown a small number of Kemp's ridleys (NRC 1990). There is potential for incidental

killing of Kemp's ridleys in dredge operations throughout the Gulf of Mexico and southeast

Atlantic coastal waters. Three juvenile Kemp's ridleys were killed in dredges at the Kings

Bay Naval Base in Georgia (Ross et al. 1989). 

Conclusion

The Kemp's ridley is the most endangered of all sea turtle species and shows little sign of

recovery. A minor increase in nesting may be an artifact of greater nesting beach coverage

or may indicate that the nesting population has increased. The lower average clutch size (95

eggs) that has been reported in recent years may indicate an influx of novice nesters (P.

Pritchard pers. comm.).

Efforts to restore the species must concentrate upon protecting subadult and adult animals

(NRC 1990) and must take place over a long period of time because of the species' slow

maturation. The key to this task is the use of TEDs throughout the U.S. and Mexican shrimp

fisheries in adult and developmental habitats, and continued protection of the species' nesting

beach.

The best available commercial and scientific information indicates that Kemp's ridley is

severely depleted and in danger of extinction (Ross et al. 1989, NRC 1990). Therefore, this

species should continue to be listed as an endangered species throughout its range. 
117

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122

Table 1. Annual nest counts at Rancho Nuevo, Mexico and estimated size of the nesting

population of Kemp's ridley turtle (Lepidochelys kempii) (R. Byles, pers. comm.). (Note: the

length of beach surveyed increased from the early years until the present. The area surveyed

has held constant since 198?. Number of nesting females is calculated based on an average

3.075 clutches per turtle per season Rostal (1991)). 

Year # Nests # Nesting Females

1978 924 301

1979 954 310

1980 868 282

1981 897 292

1982 750 244

1983 746 243

1984 798 260

1985 702 228

1986 744 242

1987 737 240

1988 842 274

1989 878 286

1990 992 323

1991 1155 376

1992 1275 415

1993 1184 385

1994 1566 509
123

Olive Ridley Sea Turtle, Lepidochelys olivacea

Prepared by Peter C. H. Pritchard, Ph.D. and Pamela T. Plotkin, Ph.D.

Biological Background

The olive ridley is a small, hard-shelled marine turtle, one of the two species of the genus

Lepidochelys , and a member of the family Cheloniidae. The species may be identified by

the uniquely high and variable numbers of vertebral and costal scutes. Although some

individuals have only five pairs of costals (the number shown by almost all individuals of the

congener Lepidochelys kempii), in nearly all cases some division of costal scutes occurs, so

that as many as six to nine pairs may be present. Division of the "standard" scutes occurs

from the rear of the carapace, so that a specimen with, say, seven pairs of costals shows

division of the homologs of costals IV and V. Asymmetry in the number of costal scutes is

frequent.

In addition to the division of the costal scutes, the vertebral scutes also show frequent

division, as do the scales on the dorsal surface of the head. The prefrontal scales, however,

typically number two pairs. The carapace is wide in subadults and adults, although less so

than that of L. kempii. In anterior profile it is typically elevated and flat-topped, with flat,

sloping sides. The plastron is large, with the usual six pairs of large scutes and sometimes a

small intergular and interanal also. The inframarginals typically number four on each side,

each of which is perforated by a pore located towards its posterior margin. The head is

relatively large, as compared with that of Chelonia or Eretmochelys, but is smaller than that

of adult Caretta, and slightly smaller than that of L. kempii.

The skull of L. olivacea differs from that of L. kempii in many ways. The orbit is

consistently larger, the width across the pterygoids is consistently less, and the alveolar

ridges are sharp and are only evident on the rhamphothecae, not on the underlying

maxillary-palatine sutural area. There are also numerous differences between the lower jaws

of the two species, as follows: i) the entire bony alveolar surface of L. olivacea is flat (rather

strongly concave in L. kempii) ii) the overall depth of the jaw of L. olivacea is somewhat less

than that of L. kempii iii) the tip of the coronoid bone, that provides attachment for the

temporal muscle closing the jaw, is rounded off in L. olivacea (bluntly pointed in L. kempii )
124

iv) the dorsal mandibular vacuity is relatively larger in L. olivacea that in L. kempii v) the

articular surfaces of the mandible of L. olivacea tend to be directed upward (directed more

posteriorly in L. kempii) vi) the bones comprising the articular surfaces (the articular, prearticular, and angular) are loosely sutured in adult L. olivacea.

In size, female olive ridleys in Surinam range in carapace length from 62 to 74 cm (mode 69

cm); in Pacific Honduras from 58 to 74 cm (mode 65 cm); and in Pacific Mexico from 56 to

78 cm (mode 68 cm) (Pritchard 1969). Márquez et al. (1976) reported upon the carapace

lengths of thirteen adult male olive ridleys from Baja California. They ranged from 58.5 to

69.0 cm.

 

The range of the olive ridley is essentially tropical. In the eastern Pacific nesting takes place

from southern Sonora, Mexico, south at least to Colombia. Non-nesting individuals

occasionally are found in waters of the southwestern United States (California), whilst they

occur abundantly in Pacific Colombia and Ecuador, but only in small numbers in Peru and

Chile.

The olive ridley has been recorded occasionally from Galapagos waters, but it is essentially

very rare throughout the islands of the Pacific, and indeed even in the western Pacific it is

scarce everywhere, although widespread low-density nesting occurs. In the Indian Ocean it

only achieves abundance in eastern India and Sri Lanka, although minor nesting occurs

alongside the green turtles at Hawke's Bay, Pakistan, and some nesting also occurs in New

Britain, Mozambique, Madagascar, peninsular Malaysia, and various other localities.

In the Atlantic Ocean, the olive ridley occurs widely, but probably not in great abundance, in

waters of West Africa, from about Mauritania southward at least to the Congo. In the

western Atlantic, nesting formerly occurred abundantly in eastern Surinam (Eilanti and

Bigisanti Beaches), as well as in western French Guiana and northwestern Guyana. 

Non-nesting individuals occur regularly as far west as Isla Margarita and Trinidad, but they

rarely penetrate any further into the Caribbean than this. The species occurs in Brazil, and

nests in the states of Bahia and Sergipe (G. Marcovaldi, pers. comm. to H. Reichart), but it

seems to be rare.
125

Geographic variation in olive ridleys is subtle, and no subspecies are currently recognized.

However, the number of costal scutes apparently varies from one area to another (Pritchard

1969), specimens with only five pairs of costals being somewhat more abundant in the

eastern Pacific than elsewhere. In addition, overall carapace coloration is typically somewhat

lighter in the western Atlantic than in the eastern Pacific. and the shell is typically more

elevated in the eastern Pacific than elsewhere.

The olive ridley was formerly known as the Pacific Ridley, until it became generally

recognized that the species also occurred in tropical parts of the Atlantic Ocean also (as well

as the Indian Ocean). In Mexico it is called "golfina" or "tortuga bestia," whereas in Pacific

Costa Rica it is usually called "lora" or "carpintera." From Guatemala to Nicaragua it is

called "paslama". In Trinidad it is called "batali"; in Guyana, "terecay"; and in Surinam and

French Guiana, "warana". Malay names include "penyu lipas", "penyu rantau," and "penyu

abu abu". In India, the Bengali name is "samudrik katha"; the Hindi name is "gahda

kachua"; the Oriya name is "samudrik kachchima;" the Tamil is "pul amai," "sith amai," or

"kadai amai," and the Telegu name is "samudram thabelu."

The most dramatic aspect of the life history of the olive ridley is the habit of forming great

nesting aggregations, generally known as "arribadas," sometimes as "arribazones" or

"morriñas". Although not every adult olive ridley participates in these arribadas, the vast

majority of them do. Formerly these nesting concentrations occurred at several beaches

along the Pacific coast of Mexico, including Piedra del Tlacoyunque (Guerrero), Bahia

Chacahua, and El Playon de Mismaloya (Jalisco), but in recent years the Mexican arribadas

have been largely restricted to La Escobilla (Oaxaca), although smaller nesting concentrations

have been reported from Morro Ayuta (Oaxaca). In Costa Rica, a major nesting aggregation

is found at Ostional, on the Nicoya Peninsula, and smaller arribadas occur at Nancite, in the

Santa Rosa National Park. Smaller arribadas also occur in Nicaragua at La Flor (Ruiz 1994)

and Chacocente and at several localities in Panama. In the Indian Ocean, four arribada sites

have been reported in the Indian State of Orissa, the most important being Gahirmatha

Beach. In the Atlantic, only small arribadas, numbering at most a few hundred animals per

night, have been reported from a single locality (Eilanti, Surinam). 

Arribadas may be precipitated by such climatic events as a strong offshore wind (in Surinam)

(Pritchard 1969), or by certain phases of the moon and tide (at Nancite, Costa Rica)
126

(Cornelius 1986), but there is a major element of unpredictability at all arribada sites. This

unpredictability, and the apparent ability of gravid females to wait for weeks while holding

fully-shelled eggs, may be an important aspect of the survival advantage of

arribada-formation, a phenomenon usually interpreted as one that evolved as a

predator-saturation device.

Individual olive ridleys may nest one, two or three times per season (Pritchard 1969, Plotkin

1994), typically producing 100-110 eggs on each occasion. The internesting interval is

variable, but for most localities it is approximately 14 days for solitary nesters and 28 days

for arribada nesters (Pritchard 1969, Kalb and Owens 1994, Plotkin 1994). The genus is

also unique in that ridleys of both species commonly, and probably typically, nest each year,

without intervening non-breeding seasons as shown by dermochelyids and other cheloniids. 

The ridleys nesting in an arribada could not be sustained by the productivity of immediately

adjacent marine ecosystems, and the species is indeed migratory. Pritchard and Trebbau

(1984) have documented the post-nesting dispersal of Surinam ridleys to the waters of all of

the Guianas, Trinidad, and eastern Venezuela as far as Isla Margarita. Cornelius and

Robinson (1986) illustrated comparable dispersal of olive ridleys from Nancite, Costa Rica,

to the waters of all countries from Mexico to Peru, with some open sea recoveries, the

farthest being 2,400 km west of Costa Rica. Results of recent investigations (Arenas and

Hall 1990, Pitman 1990, Plotkin et al. 1993, Plotkin 1994) support previous contentions

(Hendrickson 1980) that olive ridleys reside in oceanic habitats of the eastern Pacific Ocean

during the non-reproductive portion of its life cycle. Plotkin (1994) documented a nomadic

migratory behavior of female and male olive ridleys from the reproductive aggregation at

Nancite, Costa Rica and suggested that this pattern evolved in response to spatio-temporal

variations in productivity and available food resources of the eastern Pacific Ocean. 

The overall distribution of the olive ridley shows interesting parallels with that of the utterly

different leatherback turtle (Dermochelys coriacea). Both occupy oceanic habitats (see

Morreale et al. 1993 for leatherback migrations) and both nest primarily on Pacific shores of

the American tropics and in the Guianas, in moderate numbers in tropical West Africa, and

in relatively small numbers elsewhere, both being extremely rare, for example, throughout

Australia, or Pacific oceanic islands.
127

Despite its local abundance, there are surprisingly few data relating to the feeding habits of

the olive ridley. However, those reports that do exist (Caldwell et al. 1969, Carr 1961,

Márquez et al. 1976, see Mortimer 1982 for review) suggest that the diet in the western

Atlantic and eastern Pacific includes crabs, shrimp, rock lobsters, jellyfish, and tunicates. In

some parts of the world, it has been reported that the principal food is algae (Carr 1961,

Biswas 1982).

Population Size

Because of the continued existence of several large arribadas, it is probable that the olive

ridley is, in terms of absolute numbers of adult individuals in existence, the most abundant

sea turtle species in the world. Nevertheless, there is evidence of downward trends at several

arribada beaches. The various populations are under considerable stress, and the

concentration of such a large proportion of the reproductive animals into a few arribadas may

be a liability, not only in that such aggregation facilitates industrial-scale exploitation, as it

has in Mexico as well as on the feeding grounds in Ecuador, but also because arribadas do

not seem to be an efficient method of guaranteeing maximum reproductive efficiency. 

Indeed, at the relatively undisturbed arribada beach of Nancite, within Santa Rosa National

Park, Costa Rica, it has been estimated by Cornelius (1986) that only about 5 % of eggs laid

actually produce hatchlings, and D. Robinson (pers. comm.) has suggested that, under

natural conditions, it may be solitary nesters, many fewer though they may be, that bear

principal responsibility for the reproduction of the species. Contrary to Robinson's

hypothesis, Eckrich and Owens (1995) recently found that significantly more first-night

solitary nests were predated by mammalian predators than arribada nests at Nancite, thereby

suggesting that solitary nests conribute little to the overall reproduction of the species at this

beach. It is unknown if mammalian predation adversely impacts solitary nests laid on nonarribada beaches and if so, to what extent this occurs. 

The number of ridleys nesting during an arribada is difficult to count, although

methodologies to estimate arribada size have been developed that are useful if nesting is well

supervised by competent biologists (Cornelius and Robinson 1985, Gates et al. In press). On

the other hand, estimates by laymen of numbers of turtles in a given arribada are probably so

inaccurate as to be useless.

 
128

Because nesting in successive years is commonplace for olive ridleys (Pritchard 1969,

Cornelius 1986, Plotkin 1994), and may well be the norm for the species, the erratic nesting

population trend lines often shown by loggerhead or green turtle populations, that very rarely

nest in successive years, are not shown by olive ridley populations. It is thus much easier

and more justified to draw conclusions about overall ridley population trends from a few

years of comprehensive nest counts than it is for those species with multi-year nesting cycles.

There is a little evidence to suggest that olive ridley arribadas may not be permanent

phenomena, but rather may be subject to shifts over the years from one site to another. 

Thus, the Mexican arribadas at Playon de Mismaloya and Piedra de Tlacoyunque disappeared

within a few years of their discovery in the 1960s, even though it was the more southerly

nesting population at Escobilla that was subjected to the most intensive industrial slaughter,

and that still survives. Moreover, older inhabitants of the village of Ostional, located behind

the arribada beach of Ostional, Costa Rica, seem to unite in the opinion that arribadas first

occurred at that site in or around 1961 (S. Cornelius pers. comm.).

Published estimates for the sizes of nesting populations in the Western Atlantic, Costa Rica,

Mexico, and India are as follows:

Data to document the declining western Atlantic population are clear and unequivocal. 

Although Pritchard found shells of twenty of more beach-slaughtered olive ridleys on a visit

to Shell Beach, Guyana, in 1964, in the course of the last eight seasons (1987 - 1994)

intensive season-long patrols have recorded only one or two nestings per year. Nesting was

never abundant in French Guiana, and the rumored occurrence of small arribadas at Kourou

does not seem to have been based on fact. Data given by Reichart (1993) indicate that 31

nestings were recorded in French Guiana in 1977, 101 in 1978, 33 in 1979, 6 in 1980, no

data for 1981 to 1985, 52 in 1986, and 479 in 1987. The key country for ridleys in the west

Atlantic region is Surinam, and there the data show an alarming downward trend (Table 1).

At Nancite, Costa Rica, a 1982 estimate (quoted by Mager 1985) suggested that 221,000

ridleys nested there each year. Estimates of the number of female ridleys nesting at Nancite

show a decrease in the population size during the 1980's (Valverde et al. 1990). Most recent

estimates show a significant decrease: an estimated 37,123 ridleys nested at Nancite in 1990

(Zanella and Mo 1990), 34,189 nested in 1991 (Calvo and Mo 1991), and 31,029 nested in
129

1992 (Torre and Mo 1992). While a precipitous decline has occurred at Nancite, the number

of females nesting at Ostional, approximately 90 km away, appears to be stable and possibly

increasing. Mager (1985) provided a 1982 Ostional estimate of between 260,000 and

435,000 nesting ridleys.

Pacific Mexico supported an estimated 1,185,000 adults (including 593,667 females) prior to

1969; and 485,000 adults in 1976 (Cliffton et al. 1982). Large arribadas occurred at El

Playon de Mismaloya (Jalisco), Piedra del Tlacoyunque (Guerrero), Bahia Chacahua, La

Escobilla (Oaxaca), and El Morro Ayuta (Oaxaca). Several populations collapsed during the

1970s (Tlacoyunque, Chacahua, Mismaloya) (Ross 1982) and in the early 1970s only

between 179,000 and 400,000 nesting ridleys arrived at the remaining arribada beaches

(Groombridge 1982). J. Woody (pers. comm. to A. Mager) estimated the Mexican

population at 153,300 adults in 1981, 164,200 in 1982, and 79,900 in 1983. At La

Escobilla, the only remaining site in Mexico where large arribadas still occur, an estimated

202,470 ridleys nested in 1987 (Ruiz and Marin 1988). An estimated 200,000 + female

olive ridleys nested at La Escobilla during the September 1995 arribada and it is expected

that the total number of turtles nesting during the 1995-1996 season will exceed 500,000

turtles (R. Byles, pers. comm., 1995). 

Although nesting in a dispersed fashion is very widespread along the eastern coast of India

south to Sri Lanka, and to a lesser degree along the west coast as well, the great majority of

olive ridleys nest in four arribadas in Orissa. Gahirmatha, located in the Bhitarkanika

Wildlife Sanctuary, Orissa, adjacent to the western Bay of Bengal, is perhaps the most

important arribada beach in India. The estimated number of ridleys nesting during arribadas

at Gahirmatha since 1984 has fluctuated from as little as no turtles to as many as 805,000

individuals (mean = 398,636 turtles) (Mohanty-Hejmadi and Sahoo 1994). 

Listing Factors

In the western Atlantic, nesting of the olive ridley outside Surinam is desultory and, as far as

is known, has never been abundant. Accordingly, numerical trends of nesting animals in

Surinam, where turtle beaches are closely monitored by STINASU, holds the key to

population evaluation in the entire western Atlantic.
130

The nesting beaches themselves (principally Bigisanti and Eilanti) have been well protected

by personnel from STINASU, the Surinam nature protection foundation, for over two

decades. Consequently, it seems that some at-sea stress is depleting the population despite

good annual production of hatchlings and effective protection of nesting adults. The tag

returns of adult turtles that nested in Surinam indicates post-nesting dispersal to feeding

grounds extending from eastern Surinam and western French Guiana through Guyana to

Eastern Venezuela (Isla Margarita, Trinidad, etc.) (Pritchard and Trebbau 1984). This entire

area is heavily trawled for shrimp, and trawlers have been the principal source of returns of

tags from this nesting population. Consequently, it appears inescapable that incidental

capture in trawls has been the cause of the progressive depletion of this population.

Although comprehensive data to back up this conclusion are not available, Reichart (1989)

reported that, in the 1960s, a single ship engaged in periodic experimental trawling exercises

along the coast of the Guianas caught 39 olive ridleys in the course of a one year period. A

rough estimate indicated that 90-100 shrimp boats operate off the coast of French Guiana,

and an ex-crewmember interviewed by Reichart suggested that the typical incidental catch of

ridleys was 2-3 animals per day. Even larger numbers of shrimp trawlers operate out of

Guyana, and some indication of the magnitude of the problem thus becomes evident. We

also know for sure that ridleys have frequently been caught by trawlers operating in waters of

eastern Venezuela, as indicated by the returns of tags (Pritchard and Trebbau 1984). TEDs

are currently required and in use in this area, and it is hoped that with sufficient enforcement

of regulations, high mortality of sea turtles in trawls will no longer occur.

In the eastern Pacific, the situation is different and more complex. Solitary nesting by olive

ridleys occurs along an enormous coastline extending from Sonora, Mexico, to Colombia. 

Few of these beaches are patrolled by conservation crews, and egg collection by local people

is intensive throughout the area. In many parts there are as many egg collectors as there are

turtles, and no nests escape. Even so, it is relatively rare for nesting turtles to be killed along

eastern Pacific shores. On the other hand, at a few points including Escobilla and Morro

Ayuta, Mexico; La Flor and Chacocente, Nicaragua; and Nancite and Ostional, Costa Rica,

the turtles nest in arribadas, and these sites of intensive nesting are subject to at least some

measure of protection, though this varies greatly from one site to another.

A critical unknown relates to the intensity of incidental capture of olive ridleys at sea by the

recently banned high-seas driftnet fishery, and the current shrimp fishery, and long-line
131

fisheries in the eastern Pacific Ocean. Trawling is not universal throughout this area, but it

is intensive in Mexico, Guatemala, El Salvador, Nicaragua, Costa Rica, and Panama. 

Estimated turtle mortality from Guatemala to Costa Rica is 60,042 turtles per year, with most

of these mortalities presumably being olive ridleys (Arauz In press). TEDs are not currently

required, nor in use, but there are tentative agreements pending among many eastern Pacific

countries and TEDs may be required in the foreseeable future. The long-line fishery of the

eastern Pacific likely also poses a significant threat to olive ridleys. A large fleet of vessels

operates from Mexico to Panama in nearshore to offshore waters. Although there are no

quantifiable estimates available, this indiscriminate fishery often encounters, entangles,

hooks, and kills olive ridleys in the eastern Pacific Ocean (Plotkin, pers. comm.). Olive

ridleys are also captured incidentally in the western Pacific long-line fishery which operates

near the Hawaiian Islands (P. Plotkin, pers. comm.). 

Natural phenomena such as El Niños are probably very significant to the survival status of

olive ridleys in the eastern Pacific Ocean. It was recently reported (Hill 1995) that warming

trends in the Pacific, caused by frequent occurrence of El Niños since 1976, may be

responsible for the decline in zooplankton in the California Current and the corresponding

decline in higher trophic level vertebrates of this marine ecosystem. Similarly, warming in

the eastern tropical Pacific might be partially responsible for the decline in the number of

olive ridleys nesting along middle America. 

Recent information from several nations along the eastern Pacific coast of the Americas gives

insight into the current status of the olive ridley, as follows.

Mexico

In Mexico, after several decades of legal, industrial-scale take of olive ridleys, mainly for

export of leather products to Japan, the President of Mexico imposed a complete and

indefinite closed season on the utilization of sea turtles within its national jurisdiction (Aridjis

1990). The decree was signed by the Secretary of SEDUE (Urban Development and

Ecology) and the Secretary of Fisheries at Chapultepec Castle in the presence of the

governors of the important coastal states of Jalisco, Michoacan, Guerrero, and Oaxaca. 

Commercial and industrial fishermen had until 12 June 1990 to file inventories of existing
132

stocks of turtle products with the Fisheries Department. These inventoried products may

continue to be marketed under strict conditions until stocks are depleted.

An Associated Press article (published in the Milwaukee Journal, Nov. 25 1990) commented

on the changes that had occurred in Mexico as a result of the decree. The primary

slaughterhouse at San Agustinillo was closed and its roof had fallen in. On the other hand,

clandestine egg collection continues, and although the fleets of turtle boats are no longer seen

near the nesting beach at Escobilla, there are reports of two clandestine slaughterhouses.

The legal quota for olive ridley capture in 1989 was 23,000, but a total (legal plus illegal) of

as many as 70,000 may have been taken. The reported progressive increase in nesting for

the last three seasons is thus all the more surprising in view of the reported level of slaughter

during that period.

Steiner and Tweeten (1990) provide additional first-hand information on the status of the

turtle industry in Mexico in late 1990. They confirmed that the San Agustinillo

slaughterhouse had closed, but that retail sales of turtle products continues. One dealer in the

Tepito market reported an inventory of 1000 pairs of turtle-skin boots. Although aware of

the closure of the fishery and the requirement for filing an inventory of existing stocks, he

had not filled out any paperwork as required by the May 28 decree, and felt confident that

clandestine take would continue to supply the turtleskin boot industry. Turtle eggs were also

seen for sale in the Tepito Market (5000 pesos for three), and rumors were heard of certain

families continuing to capture large numbers of turtles, and new boats with dual 75 hp

engines were still being acquired for that purpose. Sales of sea turtle products in Mexico

have continued despite the 1990 decree. Steiner et al. (1994) recently reported that the

number and value of sea turtle products confiscated at the U.S.-Mexico border prior to 1990

have not changed significantly since 1990. 

El Salvador

In El Salvador, no arribadas have been identified, but there is widespread low-density ridley

nesting. It is estimated that only about 2 % of the nests laid escape predation by humans (C.

R. Hasbun, pers. comm.). In addition, it is reported that dead adult olive ridleys often wash

ashore on the 321 km of the Salvadorean coast. These are reportedly the victims of shrimp
133

trawlers, which are required by law to operate at least three miles from the shore, but which

in fact trawl within a few yards of shore much of the time.

Nicaragua

A Reuters report, filed September 25 1990, gives some information on the status of ridley

nesting in Nicaragua. The arribada beach at Chacocente, where groups of up to about 4000

ridleys come ashore to nest six or eight times per year, is patrolled by police armed with

machine guns and dressed in battle fatigues and combat boots. For years, virtually every egg

was collected immediately and sold to middlemen on the spot, who would then sell the eggs

to local restaurants or export them. However, a new decree bans turtle egg collecting

throughout the entire season (July 1 to January 31) at Chacocente and five other Pacific

nesting sites. The absolute ban has not been well received by the local people, who in some

cases have registered their displeasure by killing turtles caught at sea and stripping them of

their oviducal eggs. A new proposal, to allow egg collection during the first day (only) of

each arribada, is now under consideration.

Costa Rica

In Costa Rica, some protection of olive ridleys occurs at Nancite by virtue of its remote

location and presence in a National Park. The decline in this nesting population is cause for

concern and may indicate an overall decline of olive ridleys in the eastern Pacific. The

larger nesting colony at Ostional is managed for egg production for human consumption, a

scheme that shows great promise and may actually increase the number of hatchlings

produced, but has not yet been definitively evaluated. Hatchling predation by black vultures

(Coragyps atratus) is sufficiently intense as to suggest the need for some kind of control

measures.

Ecuador

In Ecuador, there was formerly a massive directed take of olive ridleys caught on the feeding

grounds near the Colombian border; nesting of ridleys this far south is desultory. Although

there has not been much recent information on the status of the olive ridley harvest in

Ecuador, a decree (dated July 31 1990) by the Ecuadorian Subsecretariat for Fishery
134

Resources, following a recommendation by the Charles Darwin Foundation for the Galapagos

islands, affords complete protection to all species of sea turtles in Ecuadorian waters, and

prohibits the capture, processing, and internal and overseas trade, in all sea turtle species.

West Africa and Indian Ocean

No current data are available to assess the status of olive ridleys in West Africa.

Large nesting aggregations occur in the northern Indian Ocean and Bay of Bengal. MohantyHejmadi and Sahoo (1994) reported on the size of arribadas at Gahirmatha Beach, the largest

nesting aggregation of ridleys in the world, from 1984 to 1994. The estimated number of

turtles nesting at Gahirmatha per season has ranged from zero (no mass nesting) to as many

as 805,000 ridleys. The most current estimates suggest that this population is relatively

stable, but development and enhancement of shrimp trawl fisheries in this region threaten to

change its status. In the Bay of Bengal, along the Orissa coast, several thousand olive ridleys

may be incidentally captured and killed in shrimp trawl nets annually. Some olive ridleys are

also incidentally captured and killed in the gill net fishery which operates near the coast. 

Over 5,000 dead ridleys stranded along the Orissa coast from December 1993 through April

1994 (Pandav et al. 1994). New jetties and fishing ports constructed adjacent to Gahirmatha

(Mohanty-Hejmadi and Sahoo 1994, Eckert and Eckert 1994) undoubtedly will attract more

boats to the area and threaten havoc on this population if sea turtle conservation measures

are not made mandatory. 

Conclusions

The status of the olive ridley in the western Atlantic is unambiguous. Populations are very

low and continue to decline, almost certainly as a result of long-standing incidental capture in

shrimp trawls. This population is in worse condition than that of the endangered Kemp's

ridley (L. kempii) and should clearly be listed as ENDANGERED.

Available data are too few to assess the precise survival status of the species in the eastern

Atlantic and northern Indian Oceans. However, probable widespread egg collection

throughout West Africa and known problems with incidental capture and mortality in shrimp
135

trawlers and gill nets near the arribada beaches of Orissa, India seem to indicate that the

populations there are stressed and should continue to be listed as THREATENED.

In the eastern Pacific, data indicate that some nesting aggregations are in decline, while

others appear relatively stable. There currently are no controls at all on incidental capture by

shrimp trawl and long-line fisheries throughout this vast region, which appears significant

throughout much of the region. We recommend continuation of the status quo listing of the

olive ridley in the eastern Pacific as THREATENED.
136

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139

Table 1. Annual nest counts and estimated population size for the Suriname olive ridley

turtle (Lepidochelys olivacea) (Reichart 1993). 

Year # Nests # Nesting Females

1967 2675 1440 - 2050 

1968 3290 1650 - 2350

1969 1665 830 - 1190

1970 1750 880 - 1250

1971 1595 800 - 1140

1972 1270 640 - 910

1973 890 450 - 640

1974 1080 540 - 770

1975 1070 540 - 760

1976 1160 580 - 830

1977 1030 520 - 740

1978 870 440 - 620

1979 795 400 - 570

1980 1020 510 - 730

1981 1220 610 - 870

1982 1045 520 - 750

1983 1212 610 - 870

1984 940 470 - 670

1985 670 340 - 480

1986 540 270 - 390

1987 659 330 - 470

1988 569 280 - 400

1989 585 290 - 420