The status of leatherback turtle (Dermochelys coriacea) populations inferred from nesting abundance studies worldwide led to the classification of the species as critically endangered (Hilton-Taylor 2000). Longline fisheries have had a great impact on leatherbacks and are implicated in the decline of their populations (Lewison et al. 2004). Leatherback nesting numbers on Playa Grande, Costa Rica, declined precipitously in the 1990s. The decline of the population was related to both adult mortality and high levels of poaching in past years (Spotila et al. 1996). However, despite the alarming reduction in the numbers of nesting females on Playa Grande, the population is still the most important in the eastern Pacific Ocean (Spotila et al. 2000).

Playa Grande is the primary nesting beach inside Parque Nacional Marino Las Baulas (Steyermark et al. 1996), with 2 other beaches also hosting an important number of turtles. Playa Langosta supported 10% to 15% of the nesting population in the early 1990s (Chaves et al. 1996) and Playa Ventanas also received some nesting. Part of the existing National Park was initially established as the Tamarindo Wildlife Refuge in the late 1980s (including Playa Grande and the estuary of Tamarindo). The protected area was enlarged to include all 3 nesting beaches and associated estuaries, and was converted into a national marine park by presidential decree in 1991 (Spotila and Paladino 2004). The increased protection greatly reduced the poaching of eggs (Steyermark et al. 1996). In 1995, the park was again enlarged and was permanently established by law, guaranteeing the protection of all 3 nesting beaches. Efforts to consolidate the park and to protect the beaches and their nesting turtles continue; for a history of these activities see Spotila and Paladino (2004).

To date, leatherback turtles that nest at different beaches within Las Baulas National Park have been studied independently. Females that nest on Playa Grande and Playa Ventanas have been studied for 15 years. Steyermark et al. (1996) and Reina et al. (2002) described population parameters, such as mortality rates, remigration intervals, clutch frequency, clutch sizes, and average female sizes. Meanwhile, Spotila et al. (2000) identified a severe population decline on Playa Grande and Playa Ventanas that threatened the population with extirpation. In a related but independent study, Chaves et al. (1996) investigated leatherback nesting at Playa Langosta in 1991–1992, describing female sizes, clutch sizes, and numbers of nesting females. In 1997–1998, E. Vélez and R. Piedra started a project on Playa Langosta that is ongoing to the present (Piedra et al. 2007). Both Steyermark et al. (1996) and Chaves et al. (1996) found some exchange of nesting leatherbacks between Playa Grande and Playa Langosta. However, saturation tagging and beach patrols were not started until 1997–1998 on Playa Langosta, and, therefore, observations of turtles that exchanged nesting beaches was accidental. By combining the data from all 3 beaches, it is now possible to reassess the leatherback nesting population at Las Baulas.

We describe here, for the first time, the nesting population of Parque Nacional Marino Las Baulas as a whole and analyze the use of the nesting beaches within the park by nesting females within and between seasons. In addition, we evaluate the effects of conservation efforts over the last 15 years on hatchling production and the nesting population of leatherback turtles at Las Baulas.

METHODS

Parque Nacional Marino Las Baulas is located on the Pacific coast of Costa Rica in Guanacaste Province. It includes 3 nesting beaches arranged north to south: Playa Grande (3.6 km), Playa Ventanas (1.0 km), and Playa Langosta (1.3 km). Playa Grande is separated at its south end from Playa Langosta by the Tamarindo Estuary, the town of Tamarindo, and the San Francisco Estuary (Fig. 1).

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Body pit (a depression left by nesting females) counts started on Playa Grande and Playa Ventanas in 1988–1989 and marking of individual female leatherbacks with the use of passive integrated transponder (PIT) tags (Dutton and McDonald 1994) began in 1993–1994 (Steyermark et al. 1996). In 1997, we began putting PIT tags in both shoulders of the leatherbacks. From 1997–1998 to 2003–2004, we identified 1246 leatherbacks and never observed the loss of a PIT tag. In a few cases, a PIT tag migrated and was not found on the first nesting encounter. In those instances, we retagged the turtle but always detected the missing tag upon the next encounter. Therefore, we determined that there was 100% retention of PIT tags over this time period. We began PIT tagging of females on Playa Langosta in 1997–1998. Tagging methods, as well as beach coverage (percentage of all nesting attempts where the female was observed and identified) are described in Steyermark et al. (1996) and Reina et al. (2002). We recorded PIT tag identifications each season at the 3 nesting beaches. Turtles found on Playa Ventanas were considered together with turtles that nested on Playa Grande because so few turtles nested on Playa Ventanas and they usually laid most of their nests on Playa Grande during a season. In addition, we included data collected at Playa Naranjo in 1998–1999 by Drake et al. (2004) and fortuitous observations of turtles elsewhere to document more distant interchange of leatherbacks among beaches in Costa Rica. Playa Naranjo is located in Santa Rosa National Park, 45 km north of Las Baulas.

We estimated the number of nesting females at Las Baulas per season by adding the number of turtles uniquely marked as individuals with PIT tags identified on Playa Grande and Playa Ventanas to the number of turtles identified on Langosta from 1997–1998 to 2003–2004. To account for turtles that may nest outside the main nesting season (October through February), we added the correction used by Reina et al. (2002) based on body pit counts recorded during 2 full years. On average, 7% of body pits were found between the first of March and the 31st of September. Our estimation of the number of female turtles from 1988–1989 to 1992–1993 was based on body pit counts (Reina et al. 2002). We previously determined that the average percentage of turtles that nested only on Playa Langosta within a nesting season, for those years for which we had data, was 10% of the total number of turtles at the park. We added 10% to the estimation of annual number of turtles from 1988–1989 to 1996–1997 to account for turtles that may have nested only on Playa Langosta.

For each season, we calculated the number of turtles identified, the number and percentage of turtles that remigrated in later seasons, the remigration interval for the next time we saw each turtle, and the number and percentage of turtles that we never saw again. We calculated the average remigration interval for the population by first averaging the remigration intervals per individual and then by calculating the average remigration interval for the population across all seasons.

We were confident that any turtle that was unmarked from 2000–2001 to subsequent seasons was a new recruit to the population, because that was 7 years after we began saturation PIT tagging and only 1% of remigrant turtles had remigration intervals longer than 7 years. It was likely that unmarked turtles within the 1999–2000 season were also new recruits, because only 2%–4.6% of turtles had remigration intervals longer than 6 years. However, caution must be applied to this conclusion, because 25% of the turtles that remigrated from the 1993–1994 season did so after 7 years. Therefore, several turtles from the 1992–1993 season, before we began PIT tagging, could have been returning for the first time in 1999–2000.

We used a Cormack-Jolly-Seber (CJS) model to estimate apparent annual survival rates by using the program MARK v 4.0. The CJS model estimates apparent survival rates and recapture probabilities. Because leatherback turtles do not nest every year, the recapture probability depends 1) on the probability of capturing a turtle when she is in the area and 2) on the nesting probability. Beach coverage averages 90% at Las Baulas, and the estimated clutch frequency is 7 clutches. Thus, the probability of missing a turtle every time she nests in a season is 0.10⁷. Thus, because we see every turtle during the nesting season, the probability of capture is 1, and the recapture probability equals the nesting probability. We considered the nesting probability to be a quadratic function of the years since the last nesting. With the lowest Akaike information criterion, this was the best model that fit the data. Eleven seasons were included in this analysis: 1993–1994 to 2003–2004.

We estimated hatchling production per season in the park from the total number of eggs laid on the beach and egg survival. We considered levels of poaching to be 90% until 1990–1991 (Steyermark et al. 1996; Spotila and Paladino 2004), 50% in 1990–1991, 25% in 1991–1992 and 1992–1993, and 0%–2% from 1993–1994 to the present season. We assumed that each female laid an average of 65 eggs per clutch, 7 clutches per season (Reina et al. 2002) and had a hatching success of 50% (Bell et al. 2003). We estimated that 10% of the clutches were lost because of tidal inundation (Nordmoe et al. 2004) before 1998–1999, mainly at the north and south ends of the beach. From 1998–1999, we moved clutches that were in danger of being inundated to a beach hatchery. Early in the season when high tides covered most of the beach, occasionally reaching the vegetation, we moved most of the clutches to the hatchery. Later in the season, when the tides were lower, we relocated clutches to a safer area on the beach or to the hatchery, depending on the height of tides, the number of turtles nesting on the beach at once, and the space available in the hatchery. We counted all hatchlings that emerged in the hatchery and released them into the ocean at various points along the beach, varying the location from day to day.

RESULTS

Population Analysis

From 1993–1994 to 2003–2004, we PIT tagged 1719 leatherbacks at Las Baulas. Only 448 returned to nest again. The average remigration interval for leatherbacks in the population was 3.7 ± 1.4 years. The most common remigration interval duration for females coincided with the large La Niña event that began in mid-1998 and affected the 1999–2000 and especially the 2000–2001 nesting seasons (when a large number of turtles nested; Table 1). The percentage of turtles identified in a given season that were not seen again was between 60% and 79% for the seasons for which enough years (6) had passed (before 1997–1998) for most of the turtles to come back to nest (Table 1). The CJS model indicated that the apparent annual survival rate was 0.78 (95% CI: 0.75–0.80) The annual nesting population decreased precipitously from 1504 in 1988–1989 to an average of 188 from 2000–2001 to 2003–2004 (Fig. 2).

Table 1. Number of leatherback turtles that nested each year and time in years until they first remigrated to nest again.^a

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Effects of Conservation Efforts

There were 343 new leatherbacks that nested from 2000–2001 to 2003–2004, comprising 49.5% of all the females in those 4 seasons (Fig. 3). There were another 127 unmarked turtles that nested in 1999–2000. Most of those turtles were probably new recruits as well. However, because of the long remigration interval (7 years) of 25% of the remigrant turtles from 1993–1994, we cannot be sure that all of the unmarked turtles in 1999–2000 were nesting for the first time. We placed flipper tags on all leatherbacks in 1992–1993, and, even though those tags were probably lost within a few years, turtles that had been flipper tagged usually are missing a piece of tissue from their rear flipper where the tag sloughed off. That disfigurement, called a tag scar, is easily recognized and recorded when a turtle is encountered. Of the 127 unmarked turtles in 1999–2000, only 7 had tag scars. Therefore, up to 120 of those turtles were probably new nesters. Thus, between 343 and 463 new leatherbacks nested at Las Baulas between 1999–2000 and 2003–2004.

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Hatchling production increased substantially since the park was created. There were 30,788 and 30,180 hatchlings produced on the beach in 1988–1989 and 1989–1990, respectively, when the number of nesting females was the highest. The maximum of 153,547 hatchlings were produced on the beach in 1992–1993, despite a decline in number of turtles from 1504 to 1000 (Fig. 4). The lowest hatchling production was 15,374 in 2002–2003 because of the lowest number of nesting turtles that season. However, that hatchling production was approximately half the number of hatchlings produced in 1988–1989, whereas the number of nesting females was 95% lower. In addition, there were a similar number of hatchlings coming off the beaches of Las Baulas in 1998–2003, with 68 to 188 nesting females under the current conservation regime, as there were in 1988–1989 with 1500 nesting females and no park. In 1999 and 2000, the 246 and 417 nesting females, respectively, produced 2 to 3 times as many hatchlings as were produced in 1988 and 1989. Beach protection and the establishment of a hatchery helped to offset the effect of the decline in numbers of nesting females. Because the park was established in 1991, there were 994,451 hatchlings produced, of which 20,951 were produced in the hatchery (Table 2).

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Table 2. Number of hatchlings produced in the hatchery on Playa Grande at Las Baulas Park from 1998–1999 to 2003–2004.

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Beach Exchange

We identified 1252 individual leatherback turtles at Las Baulas from 1997–1998 to 2003–2004. Most of the turtles, 71% ± 3.0 (SE) nested only on Playa Grande within a given season, 10% ± 1.9% nested only on Playa Langosta, and 18% ± 2.4% exchanged beaches at least once in a season (Table 3). In 1998–1999, 6 turtles nested only at Playa Naranjo, 1 turtle nested at Playa Naranjo and Playa Langosta, and 1 turtle from Playa Grande later nested at Ostional.

Table 3. Number and percentage of leatherbacks, Dermochelys coriacea, that nested only at Playa Grande; only at Playa Langosta; at both Playa Grande and Playa Langosta; or at Playa Grande, Playa Langosta, and other locations during each nesting season from 1997–98 to 2003–2004.

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Exchange rates between beaches in different seasons were high. Of 209 turtles that nested in 2 seasons, 136 nested both seasons on Playa Grande, 13 turtles nested only on Playa Langosta, and 60 turtles exchanged beaches between seasons or within a season (Table 4). Of 22 turtles that nested in 3 seasons, 13 nested on Playa Grande all 3 times, none nested exclusively on Playa Langosta each season, and 9 turtles exchanged nesting beaches between seasons. Only 1 turtle nested 4 seasons, and it nested exclusively on Playa Grande all 4 times. In the past, we analyzed the population by counting only the turtles that nested on Playa Grande and Ventanas. Therefore, we did not account for the females that nested only at Playa Langosta during a given year. In doing so, we missed 9% of turtles that nested only 1 season, 15% of turtles that nested 2 seasons, and 9% of turtles that nested 3 seasons, because they nested exclusively at Playa Langosta during at least one of the seasons (Table 5). This underestimated both the population size and the survival rate of the turtles.

Table 4. Nesting beach at which leatherback turtles nested each season for turtles that nested 1, 2, 3, or 4 seasons.

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Table 5. Number of leatherback turtles that nested always at the same beach (Playa Grande or Playa Langosta) or at different beaches between seasons, and number of turtles that would have been missed at least 1 season without information from Playa Langosta.

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DISCUSSION

Parque Nacional Marino Las Baulas comprises 2 major nesting beaches: Playa Grande (including Playa Ventanas) and Playa Langosta, which are both used by a large number of individual nesting leatherbacks. Leatherbacks that nest at Las Baulas show a high fidelity to a single beach, but movements between Playa Langosta and Playa Grande are not uncommon. On average, 71% of leatherbacks nest only on Playa Grande, 10% nest only on Playa Langosta, and 18% nest on both beaches in a given year. In other years, leatherbacks may shift to nest primarily on a different beach. Furthermore, 82% of the turtles that nested at Langosta within 2 seasons also nested at Playa Grande, and 100% of the turtles that nested at Langosta within 3 seasons occasionally nested also at Playa Grande.

There are 3 hypotheses that evaluate this behavior. Tucker and Frazer (1991) hypothesize that a leatherback population contains 2 components. Most of the population is faithful to 1 beach or a small area. Another segment of the population shows less site fidelity and nests on beaches over a larger area. For example, leatherbacks at St. Croix also nest at Culebra and on other beaches in the region (Eckert et al. 1989). Based on aerial surveys and personal communications from other leatherback research projects on the Pacific side of Costa Rica, we know that 85%–90% of the leatherbacks that nest on the Pacific side of Costa Rica do so at Las Baulas. Furthermore, most of the turtles that have been seen at other locations have no tags from Las Baulas. Thus, it appears that Las Baulas leatherbacks are, in general, faithful to the beaches of the park and seldom wander to more distant beaches. Therefore, this hypothesis does not explain nesting of leatherbacks in this population.

Nordmoe et al. (2004) suggest that a proximity hypothesis would better explain nesting distribution than a fidelity hypothesis. The investigators conclude that leatherbacks simply nest close to the site of their previous nest as opposed to being faithful to a particular location on the beach. However, this study was conducted on nest locations only on Playa Grande, whereas the scope of our analysis revealed that there is considerable shifting of nesting by leatherbacks between the beaches of the park, more than expected by chance (>5%). Therefore, the proximity hypothesis cannot address nesting distribution among beaches within the park.

Nearly 90% of leatherbacks at Las Baulas nest at Playa Grande. Of these individuals, 71% nest only on Playa Grande, and 18% nest on both Playa Grande and Playa Langosta. Therefore, an appropriate third hypothesis is that environmental conditions may influence selection of nesting beaches within Las Baulas Park. Thus, Playa Grande may be the principal nesting beach for leatherbacks within the park because of unique aspects of its location and physical characteristics. It is the largest beach and may provide the best physical conditions for approaching the beach from the sea. Playa Grande may also provide nesting conditions, such as sand characteristics, beach slope, or tidal dynamics that are more suitable for leatherback nests and embryonic development. To effectively evaluate this hypothesis, properties of the beach and near-shore physical structure and dynamics require further investigation.

Because of the mixing of leatherbacks between Playa Grande and Playa Langosta, analyzing the nesting population at the 2 beaches independently leads to errors and an underestimate of the population size and nesting dynamics of the population. This also leads to an incorrect estimation of remigration intervals if we consider exclusively turtles that nest at either beach. The data presented here emphasize the importance of all beaches in the national park and the necessity of coordinated protection within the park to ensure the survival of the population.

Nesting cohorts of leatherbacks at Las Baulas showed variable remigration intervals across seasons. The most common remigration intervals for most cohorts coincided with the higher than average nesting season in which 417 turtles nested (2000–2001). The second most common remigration interval coincided with the 1999–2000 season when 246 turtles nested; these were both “La Niña” years of increased productivity in the Pacific Ocean. Remigration intervals in some species of sea turtles have been related to El Niño Southern Oscillation (Limpus and Nicholls 1988; Hays 2000; Solow et al. 2002), and some studies suggest that Las Baulas leatherbacks were also affected by this phenomenon (Steyermark et al. 1996; Wallace et al. 2006). However, additional analyses are required to elucidate this relation.

Adult annual mortality rates calculated with the CJS model (22%) were slightly lower than the 25% previously estimated by Reina et al. (2002). This revised estimate was because of the inclusion of 4 more years of data that documented additional remigrant females. For example, Reina et al. (2002) included the years from 1993–1994 to 1999–2000, whereas the present analysis included data from 2000–2001, when a large number of turtles nested after the occurrence of a La Niña, as well as integrated data from Playa Langosta. Although this revised estimate is more accurate, a 22% annual mortality rate is still unsustainable for the population and double the 11% recently estimated for the leatherback turtles that nest at Sandy Point, St. Croix (Dutton et al. 2005). Demographic modeling (Spotila et al. 1996) indicates that this rate will lead to extinction of the population.

The annual mortality rates and a shrinking temporal buffer suggest this population faces an imminent threat of extinction (Fig. 2). Conservation on the beach has significantly increased the production of hatchlings, and poaching has been reduced from 90% to less than 1% since 1990. Since the park was declared, park rangers, local guides, and researchers have been protecting natural nests on the beach to prevent poaching. Equally, nests threatened by tidal inundation are moved to a hatchery or relocated to a safer area on the beach, thus increasing by 10% the production of hatchlings on the beach. Increasing production of hatchlings will eventually result in a higher proportion of neophyte nesters and enhanced recruitment. Currently, 49.5% of the turtles are new nesters, which probably reflects the increasing protection on the beach in past seasons. At St. Croix, 51.5% of leatherbacks were new nesters after several years of conservation efforts (McDonald and Dutton 1996).

However, fewer clutches were laid since 1988–1989, because the number of nesting females decreased, apparently because of the high levels of adult mortality. Survival of late life-history stages in a population is critical to the growth of turtle populations (Crowder et al. 1994). Congdon et al. (1993, 1994) reported that slow growing and late maturing species such as turtles cannot sustain increased levels of adult mortality because of human harvesting. Crowder et al. (1994) and Crouse and Frazer (1995) concluded that sea turtles, in general, are severely limited in their ability to respond to current levels of mortality. Spotila et al. (1996) reported that simulations indicated that increases in adult mortality as low as 1% caused a slow decline in an otherwise stable leatherback population and that an increase in adult mortality of 5% caused a rapid decline in such a population. Thus, incidental capture of adults and subadults in the oceans by fishing activities could still lead to a population decline, regardless of the conservation efforts on the beach (Crouse et al. 1987; Heppell et al. 1996; Heppell and Crowder 1998). However, the simulations by Spotila et al. (1996) indicated that increased production of hatchlings could, under some circumstances, compensate for an incidental adult mortality of up to 5%.

To understand to what extent beach protection could offset incidental fisheries mortality, we made a first approximation of recruitment for the Las Baulas population. We made the following assumptions based on our previous studies, data in this paper, and other studies. A leatherback at Las Baulas lays 65 eggs per clutch, 7 clutches a season, and nests every 3.7 years (Spotila et al. 1996; Steyermark et al. 1996; Reina et al. 2002; this study). If she has a reproductive life of 20 years (Hughes 1996; Spotila et al. 1996) and 10% of the clutches are lost because of tidal inundation, then she would produce 2215 eggs during her lifetime. Because of protection on the beach, poaching dropped from about 90% in 1988 to near zero by 1993–1994. When considering a hatching success of 50% (Bell et al. 2003), 1108 hatchlings would be produced to replace the female and a male if the population were stable and the sex ratios were 1:1. Thus, 554 hatchlings would be needed to produce an adult turtle without human-induced mortality at any level. Thus, we can assume that production of 550 hatchlings should result in 1 female returning to nest 9 to 11 years later (Zug and Parham 1996; Dutton et al. 2005). We recorded 343 new turtles in 4 seasons, from 2000–2001 to 2003–2004. If we total the hatchlings produced on the beach during 4 consecutive seasons and assume an average age at maturity of 9 years, then the 439,885 hatchlings produced from 1991–1992 to 1994–1995 should have resulted in 794 new turtles recruited during 2000–2001 to 2003–2004. If age at maturity was 10 years, then we would expect 719 new turtles, and, if age at maturity was 11 years, then we would expect 702 new nesters. The actual number of new recruits was less than half these quantities, which suggests that mortality rates in the ocean before maturation are double the levels of those of a stable population. Although age to maturity might be longer and more variable among females than the assumptions used here, the timing of the observed increase in the St. Croix leatherback nesting population suggested that recruitment of neophytes occurred ca. 10–12 years after significant conservation efforts to enhance hatchling production began (Dutton et al. 2005). Given these several assumptions, this first approximation of the expected recruitment to the nesting population is lower than expected, despite conservation efforts on the beach. These conclusions place further emphasis on increased consideration of fishing impacts on the Pacific leatherback population.

Beach protection has proven effective elsewhere to maintain leatherback populations in the Atlantic and Indian Oceans. At St. Croix, 30% of nests were in danger of tidal inundation, erosion, or being poached, and were relocated to safer areas from 1982 to 1996. This population showed an exponential growth since 1991–1992, 9 years after the beach protection was intensified, as a consequence of the increased hatchling production on the beach (Boulon et al. 1996; Dutton et al. 2003). Equally, the nesting population of Tongaland, KwaZulu-Natal in South Africa was protected after the area was declared a marine reserve. Consequently, the annual nesting population increased from 18 to over 100 females per nesting season within 10 years (Hughes 1996). The leatherback populations that nest at St. Croix and South Africa are examples of successful conservation efforts on nesting beaches. The St. Croix population, however, has lower mortality rates in the ocean as adults than Pacific leatherbacks that nest at Las Baulas (Lewison et al. 2004). Beach protection is insufficient as a sole measure to ensure the survival of the Las Baulas nesting population. Our study suggests that only a combination of beach protection and a reduction of mortality rates to the critical life history stages in the ocean can prevent the extirpation of this population.