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Natural and anthropogenic mortalities were expected to occur in all Kemp’s ridley life stages following the nesting season in 2009 (Crowder & Heppell 2011; NMFS et al. 2011; Gallaway et al. 2013; Heppell In press), but the only life stages that could have affected the 2010 nest count were adult females, and subadult females that matured between the ends of nesting seasons in 2009 and 2010. However, natural and anthropogenic mortalities in adult and subadult females were not expected to reach levels high enough to interrupt exponential growth in annual nests (Crowder & Heppell 2011; NMFS et al. 2011; Gallaway et al. 2013; Heppell In press). After the 2010 drop in nests, the substantial increase in nests in 2011 and the slight increase in nests in 2012 (Fig. 1; see also Burchfield & Peña 2013) were encouraging, but probably resulted from population momentum. They suggested that population growth had quickly resumed (Gallaway et al. 2013), and provided hope that exponential growth would soon resume. However, the numbers of nests in 2011-2013 were well below those predicted (Fig. 1).

Figure 1. Observed and predicted annual numbers of Kemp’s ridley nests on Rancho Nuevo, Playa Dos-Barra del Tordo, and Tepehuajes beach segments combined, Tamaulipas, Mexico, 2009-2013.

Figure 2. Kemp’s ridley annual nests and annual hatchlings released at Rancho Nuevo, Playa Dos-Barra del Tordo, and Tepehuajes beach segments combined, Tamaulipas, Mexico, 1966-2013. Numbers are in logarithmic scale to show the approximately straight trends during the late 1980s through 2009, which are evidence that the observed increases in numbers within this interval were exponential.

The drop in nests in 2013, and a preliminary datum for nests in 2014 suggesting a further drop, prompted me to begin drafting this commentary in mid-July 2014. The 2014 drop in nests was confirmed by the 12 August 2014 message posted on CTURTLE entitled “Abrupt Setback for the Recovery of the Critically Endangered Kemp’s Ridley Sea Turtle” (<http://www.lists.ufl.edu/cgi-bin/wa?A0=CTURTLE>). I adapted the 1966-2013 time series of nests and hatchlings data in Figs. 1 & 2 from Gallaway et al. (2013), Benny Gallaway (personal comm., August 2014), and Gladys Porter Zoo’s 2009-2013 annual reports on the Mexico-U.S. Kemp’s ridley population restoration project. Some data points in Figs. 1 & 2 may differ from those published previously, but this does not materially affect my hypothesis.

From mid-July through 06 August, I shared my hypothesis and drafts of my commentary with others, and invited some of them to join me as coauthors, but none accepted. I proceeded with the commentary, believing it necessary to focus additional attention on the unexpected crisis facing Kemp’s ridley recovery, the urgent need to determine the cause or causes of extraordinary post-2009 changes in population trajectory, and the need for funding to support conservation, monitoring, research, and demographic modeling going forward. The Second International Kemp’s Ridley Sea Turtle Symposium (<http://www.kempsridley.info>), to be held on 18-19 November 2014, in Brownsville, Texas, “...will provide an opportunity to highlight the status of the species and immediate needs to get it back on track towards recovery.”

I hypothesize that the 2009 age distribution and momentum of the Kemp’s ridley population in the Gulf of Mexico were fundamentally altered by substantial reductions in numbers of turtles of both sexes at all ages, following the end of the 2009 nesting season (Caillouet 2011). Age distribution is self-explanatory. n addition, population momentum is related to population age distribution. Population momentum was described by Keyfitz (1971) (see also Koons et al. 2006) and discussed by Heppell et al. (2007) and Caillouet (2010, 2011). Koons et al. (2006) is particularly apropos, because their study of vertebrate populations concluded that “population momentum will play a critically important role in the population dynamics of long-lived, late maturing organisms that are exposed to large changes in environmental conditions that are caused naturally or via anthropogenic actions.” In this regard, the data- rich history and current situation of the Kemp’s ridley population provide opportunities for study and modeling of anthropogenic and natural environmental effects on a previously rapidly growing vertebrate population. Although my hypothesis covers both sexes, reduced numbers of males do not limit the number of clutches laid, but reduction in numbers of adult males could affect fertilization of eggs (Coyne & Landry 2007). Therefore, my commentary focuses on females, and the baseline for female Kemp’s ridley age distribution and momentum were those existing at the end of the 2009 nesting season.

A critical underlying assumption of my hypothesis is that there were no post-2009 reductions in annual efforts expended toward searching for and documenting nests in Tamaulipas as compared to efforts expended in 2009 (Caillouet 2011; Gallaway et al. 2013). If searching and documentation efforts decreased after 2009, my hypothesis has no basis. Likewise, failure of this assumption would also affect validity of post-2009 predictions by demographic models (e.g., Crowder & Heppell 2011; NMFS et al. 2011; Gallaway et al. 2013; Heppell In press).

Fundamental alteration of the 2009 age distribution and momentum of the female portion of the Kemp’s ridley population represents (hypothetically) a much greater population setback than previously reported (e.g., Crowder & Heppell 2011; Gallaway et al. 2013). If my hypothesis is true, conservation efforts going forward may not be sufficient to prevent further decline for a decade or more, because of the time-lag between release of new cohorts of hatchlings and maturation of surviving females from these cohorts. However, without such efforts the decline could be protracted further. For these reasons, currently unused methods of enhancing hatchling production may be worthy of consideration, discussion and debate (see Mrosovsky & Godfrey 2010). USFWS cuts in funding for conservation efforts on nesting beaches in Tamaulipas, Mexico pose additional challenges going forward (Plotkin & Bernardo 2014; <http://www.tamucc.edu/news/2014/06/060614%20Hartes%20Heroes.html#.VBmXajF0yos>). These cuts have been replaced by funding from other sources through 2014, but funding for years 2015 and beyond is uncertain. NMFS et al. (2011) listed lack of funding as a threat that could reverse Kemp’s ridley population growth.

Crowder & Heppell (2011) and Gallaway et al. (2013) applied models to predict post-2009 changes in annual nest numbers, under hypothetical scenarios involving mortality estimated for various Kemp’s ridley age groups in 2010. Modeling by Crowder & Heppell (2011) was based on a time series of annual nests and hatchlings ending with year 2003, and it incorporated a hypothetical upper limit of 12,000 nests protected annually in corrals, with the rest left in situ. The stock assessment model applied by Gallaway et al. (2013) was based on an updated time series of annual nests and hatchlings, it did not place an upper limit on nests protected annually in corrals, and it incorporated a metric of annual shrimp trawling mortality in the Gulf of Mexico. No previous demographic model of sea turtle populations had incorporated a metric of shrimp trawling mortality, despite the finding by Magnuson et al. (1990) that incidental capture in shrimp trawls was the most important source of post-pelagic sea turtle deaths at sea. For year 2010, Gallaway et al. (2013) estimated that Kemp’s ridley mortality attributable to shrimp trawling was 12.6% of annual total mortality (anthropogenic and natural combined) for ages ≥ 9 years, and 4.4% of annual total mortality for ages ≥ 2 years. In other words, the larger percentage of estimated annual total mortality in 2010 was left unexplained, but was not attributable to shrimp trawling.

Under my hypothesis, reductions in numbers at age for Kemp’s ridley females following the 2009 nesting season should not be expected to be proportionate. It is more likely that numbers at age were disproportionally impacted. Impacts of anthropogenic and natural factors on Kemp’s ridleys depend on spatio-temporal distributions of the turtles of various ages in relation to spatio- temporal distributions of these factors. A simple example is provided by the 2-yr pelagic stage (ages 0-1 years; Witherington et al. 2012; Gallaway et al. 2013). The pelagic life stage is not vulnerable to incidental capture in shrimp trawls, but all post-pelagic Kemp’s ridleys (ages 2 years and older) are vulnerable to shrimp trawling (Gallaway et al. 2013). Spatio-temporal distribution of the pelagic stage is determined by locations and timing of entry of hatchlings into the Gulf of Mexico, and dispersal thereafter by oceanic surface circulation (Ogren 1989; Collard & Ogren 1990; Márquez M. 2001; Putman et al. 2010, 2013; Witherington et al. 2012). Spatio-temporal distributions of post-pelagic life stages and factors that threaten them have been elucidated by mark-recapture, tracking, in-water sampling, observer programs, bycatch in fisheries, and strandings (e.g., Caillouet et al. 1996; McDaniel et al. 2000; Morreale et al. 2007; NMFS & USFWS 2007; NMFS et al. 2011; Seney & Landry 2011; Garrison & Sasso 2012; Lewison et al. 2013; Shaver et al. 2013).

The post-2009 departure of observed annual numbers of nests from those predicted (Fig. 1) obviously reflected a reduction in nesting by females that were already adults and those that became adults after the 2009 nesting season. Clearly, if numbers of adult and subadult females were substantially reduced by higher than expected mortality, this could explain reduced numbers of nests in 2010 and subsequent years as compared to model-predicted numbers of nests. However, nesting also could have been reduced by non-lethal factors that prevented migration to nesting beaches, egg production, or both (Benny Gallaway, personal comm., August 2014). Migration from foraging areas to nesting beaches and production of eggs require reserve energy, and in any given year some adult females in the population may have insufficient energy reserves to support migration or production of eggs (Márquez-M. 2001; Witzell et al. 2005, 2007; Morreale et al. 2007; Rostal 2007; Shaver et al. 2013; Shaver & Caillouet In press). Energy demands for migration and egg production must be substantial, because Pritchard (1980) observed “...that nesting ridleys invariably look extremely lean, without bulging fat around the limb bases typical of captive animals, and the neck always looks scrawny with the space under the front of the carapace cavernous and shrunken.” It is therefore conceivable that post-2009 annual numbers of nests (Fig. 1) reflected failure of large numbers of potential nesters to nest in 2010 and beyond because they became debilitated, undernourished, or experienced reproductive failure for other reasons following the 2009 nesting season (e.g., see <http://www.fws.gov/contaminants/fws_oscp_05/fwscontingencyappendices/L-WildlifePlans/turtle.pdf>; Caillouet 2010, 2011; Gallaway et al. 2013; Shaver et al. 2013; VanderKooy 2013). Examination of available strandings data for subadult and adult female Kemp’s ridleys in years 2009-2014 and beyond, including live-strandings and fresh carcasses, could shed light on theory that survivors were unable to nest for whatever reason.

In 2009, prospects for Kemp’s ridley recovery were promising. Intensive conservation efforts in Tamaulipas had reversed the previous population decline by 1986, and the nesting range had expanded (Caillouet 2010; Caillouet et al. In press; Shaver & Caillouet In press). The use of turtle excluder devices (TEDs) in shrimp trawls and declining shrimping effort had accelerated the population’s growth (Heppell et al. 2005, 2007; Caillouet 2006, 2010; Gallaway et al. 2013). Annual nests and hatchlings had been increasing exponentially for more than 2 decades (Fig. 2). Hatchling inputs had restored the population’s age structure, and increased the number of cohorts among females nesting in a given year. By 2009, subadult and adult female life stages represented more cohorts but substantially fewer individuals than all younger life stages combined, which represented fewer cohorts but substantially more individuals (see Seminoff & Shanker 2008). NMFS et al. (2011) characterized post-2009 expectations for recovery of the Kemp’s ridley population as follows: “We anticipate that the Kemp’s ridley will attain its downlisting criterion of 10,000 nesting females in a season by 2011. Based on population growth rates of 19% per year, we anticipate that the Kemp’s ridley could attain its delisting criterion of an average of 40,000 nesting females per season over a 6-year period by 2024.” USFWS also was optimistic (<http://www.fws.gov/endangered/news/bulletin-summer2009/brighter-future-for-kemp.html>): “The Kemp’s ridley nesting population is increasing, and we are cautiously optimistic that the species is on its way to recovery.”

The cause or causes of departure of observed annual numbers of nests from those predicted (Figs. 1 & 2) remain to be determined. The Deepwater Horizon oil spill and remedial actions taken to mitigate it (Antonio et al. 2011; Belter 2014; Fikes et al. 2014), incidental capture in shrimp trawls (<http://sero.nmfs.noaa.gov/protected_resources/sea_turtles/documents/shrimp_biological_opinion_2014.pdf>), other anthropogenic threats, and environmental phenomena are being considered as possible contributors (Caillouet 2010, 2011; Crowder & Heppell 2011; Fikes et al. 2014; Gallaway et al. 2013; NMFS et al. 2011; Safina 2011).

Also to be considered is a possible effect of Gulf of Mexico carrying capacity for Kemp’s ridleys (Gallaway et al. 2013). Typically, a population’s growth curve is sigmoid in shape (symmetrical or asymmetrical). Once an exponentially growing population passes the inflection point on its growth curve, growth continues but at a diminishing rate as the population approaches an upper asymptote; the magnitude of the asymptote is related to environmental carrying capacity. At a stakeholders meeting (either April 2004 or February 2006) held by the Kemp’s Ridley Recovery Team (<http://www.fws.gov/kempsridley/meetingschedule.html>) in Houston, Texas, Peter C.H. Pritchard suggested that the inflection point on the Kemp’s ridley population growth curve might occur at a high population level, as a consequence of many years of intensive conservation efforts applied toward its recovery (Caillouet 2006), implying that the population might overshoot environmental carrying capacity. A population crash might result from such a situation. Heppell et al. (2007) recognized that reductions in habitat and prey availability may have decreased the maximum population size that can be attained. They also recognized that reductions in carrying capacity can result from limitations of available nesting habitat and prey. Furthermore, they noted that population growth could not continue indefinitely because of changes in egg survival rates as a result of decreased per capita protection on nesting beaches, possible density-dependent changes in survival and growth, and the potential for increased natural or anthropogenic threats. If carrying capacity for the Kemp’s ridley population was abruptly reduced in 2010 by impacts on Kemp’s ridley prey species (Caillouet 2010, 2011; Shaver et al. 2013; VanderKooy 2013), this could have been a factor influencing post-2009 changes in annual numbers of nests.

Figure 3. Annual average hatchlings per nest for Kemp’s ridley nests on Rancho Nuevo, Playa Dos-Barra del Tordo, and Tepehuajes beach segments combined, Tamaulipas, Mexico, 1966-2013.

The product of annual number of nests and annual average number of hatchlings released per nest estimates reproductive output of nesters in a given year (Witzell et al. 2005, 2007; Heppell et al. 2007). Annual average number of hatchlings released per nest peaked in 1989, thereafter declining to a little more than half the 1989 level by 2013 (Fig. 3). The decline could represent reduction in fecundity of nesters as well as many factors affecting clutch survival rates (Márquez-M. 1994; Witzell et al. 2005, 2007; Heppell et al. 2007). A long-term trend toward younger, less fecund nesters would be expected as a result of exponentially increasing annual numbers of hatchlings released (Fig. 2). The decline in annual average number of hatchlings released per nest emphasizes the need for time series of annual numbers of first-time (or neophyte) nesters, and their annual proportion of all nesters. Regardless of its cause, the observed decline in annual average number of hatchlings per nest suggests it will now take about twice as many nests to produce a given number of hatchlings as it did in 1989. This will be a further challenge to restoration of exponential population growth going forward.

An obvious question arises as to how my hypothesis can be tested. Sample annual age distributions of Kemp’s ridleys would be useful in testing the hypothesis, and they could also be used to validate demographic models. However, annual age distributions are difficult if not impossible to determine by sampling. Direct methods (e.g., skeletochronology, based on dead specimens) and indirect methods (e.g., based on mark-recapture and strandings data) have been used to determine ages of Kemp’s ridleys (Snover et al. 2007; Avens & Snover 2013; Gallaway et al. 2013). Estimated relationships between post-pelagic female Kemp’s ridley age and straight carapace length (SCL) are numerous, but they do not apply as well to estimating ages of adults as they do to estimating ages of juveniles and subadults, since growth slows or stops after Kemp’s ridleys reach maturity (Snover et al. 2007; Caillouet et al. 2011; Avens & Snover 2013). Nevertheless, available annual SCL measurements from all sources can be converted to age using available SCL-age relationships (Snover et al. 2007; Avens & Snover 2013; Gallaway et al. 2013). Statistical comparisons can then be made among annual age distributions derived from each SCL-age relationship applied to SCL measurement data from various sources, or all sources combined. Annual age distributions derived by converting SCL to age can also be compared to model-generated annual age distributions. These approaches can also be applied to nesters only. Model-generated annual age distributions of nesters can also be used to predict the proportion that first-time nesters represent of total nesters in a given year. Comparisons of model- generated age distributions of nesters among years 2009-2014 and beyond are possible.

Comparing annual SCL distributions of nesters, derived from annual samples taken on Tamaulipas beaches over the years, could also be informative in testing my hypothesis, validating demographic models, and validating SCL-age relationships. For example, demographic models can be used to generate annual SCL distributions, which can be compared to annual SCL distributions obtained from samples from various sources, or all sources combined. The need for recent annual SCL distributions of nesters was recognized by Gallaway et al. (2013), and a large sample of SCL measurements of nesters in Tamaulipas was obtained in 2014 (Benny Gallaway, personal comm., August 2014). If nester SCL measurement data are available for other years, they could be valuable for making annual SCL distribution comparisons among years, especially 2009-2014.

Figure 4. Cumulative distributions (%) of SCL (cm) of Kemp’s ridley nesters at Rancho Nuevo, Tamaulipas, Mexico, in 1996 (solid dots) and 1980 (solid triangles). Data adapted from Chavez et al. (1968) and Pritchard (1980), respectively.

I constructed cumulative distributions of SCL for Rancho Nuevo nesters in 1966 and 1980, with data adapted from Chavez et al. (1968) and Pritchard (1980); they did not differ very much (Fig. 4), except for the absence of females > 71 cm SCL in the 1980 sample. The 1966 sample was made up of residual nesters remaining when conservation efforts began at Rancho Nuevo, and the 1980 sample contained residual nesters plus younger nesters added as the result of these conservation efforts (see Márquez-M. 1994). Márquez-M. (1994) tabulated maximum, median, and minimum SCL and sample size for years 1966-1992 (Figs. 5 & 6). Median SCL showed little variation, but maximum and minimum SCL varied quite a bit, in part due to wide variation in sample size. The wide range in SCL of nesters representing multiple cohorts in a given year, coupled with limited variation in their median or mean SCL over years (Márquez-M. 1994; Márquez-M. 2001), suggests to me that the SCL distribution of neophyte nesters may be indistinguishable from that of nesters representing all accumulated cohorts, with the possible exception that neophytes have not had time to grow to the larger SCLs that are possible with much older nesters. Thus it is important to identify neophyte nesters from other nesters in a given year. In fact, Márquez-M. (1994) wrote that “...in the wild, size is not that important in determining sexual maturity.” He noted that females nesting at Rancho Nuevo and in captivity can be as small as 55 cm SCL, and concluded that “...age and size at initial maturity are so variable that they should be checked annually in order to conduct meaningful population analysis.” Genetic differences among females that survive to maturity, and differences in their histories of exposure to factors that affected their growth, can lead to a wide variation in SCL and age at maturity and first nesting (Márquez-M. 1994). Demographic modelers have not incorporated such variation into their model runs; instead, they have assumed a fixed age at maturity. It is not likely that all surviving females of a given cohort mature simultaneously. Gallaway et al. (2013) included an optional maturity schedule or ogive in their stock assessment model, so that variation in age and SCL at maturity could be accommodated in the future. However, this option was never used. In this regard, the approach recommended by Heino et al. (2002) to relate age and size at maturity may be worthy of application to Kemp’s ridley population modeling.

Figure 5. Annual maximum, median, and minimum SCL (cm) of Kemp’s ridley nesters at Rancho Nuevo, Tamaulipas, Mexico, during 1966-1992. Adapted from Márquez-M. (1994).

Figure 6. Sample size (n) for SCL (cm) measurements (see Figure 5) of Kemp’s ridley nesters at Rancho Nuevo, Tamaulipas, Mexico, during 1966-1992. Adapted from Márquez-M. (1994).

Examination of SCL distributions of Kemp’s ridleys strandings in the Gulf of Mexico will also be useful in testing my hypothesis. For example, the annual SCL distributions of Kemp’s ridleys strandings in the Gulf of Mexico in 2010 and 2011 were strikingly different from that in 2009 (Gallaway et al. 2013). Going forward, it will be important to examine annual SCL distributions of strandings and to make statistical comparisons among years, especially 2009-2014 and beyond.

Whether or not my hypothesis is true, there can be no doubt that the Kemp’s ridley population changed after the 2009 nesting season, and these changes likely will not be quickly rectified. Kemp’s ridley recovery has been postponed. Questions remain as to the cause or causes of this setback, what should be done to restore exponential growth in the population, how it will be funded, and how long it will take. I sincerely hope that my hypothesis is wrong, and that sustainable growth in the Kemp’s ridley population will soon resume!

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