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To test the hypothesis of excessive abundance of neritic juvenile Kemp’s ridleys, Caillouet (2019) recommended that age-structured modeling be used to estimate post-1984 annual numbers of neritic juveniles and adults, so that a post-1984 time series of the quotient derived from annual number of adults divided by annual number of neritic juveniles could be examined. If this quotient declined, the decline would support the hypothesis. However, even if such analyses supported the hypothesis and annual hatchling production on Tamaulipas beaches were reduced substantially, it could take 10 yrs or more before effects could be detected, because of the time lag related to age at sexual maturity (Avens et al. 2017, 2020; Caillouet 2019). This lends urgency to implementing the as-yet unfulfilled age- structured modeling and examination of the post-1984 time series of the quotient. Recommendations by Caillouet (2019) are consistent with previous extensive uses of age-structured modeling to assess effects of conservation interventions and other factors affecting status and trends of the Kemp’s ridley population (Márquez-M. et al. 1982; Heppell et al. 1996, 2005, 2007; Heppell & Crowder 1998; TEWG 1998, 2000; Crowder & Heppell 2011; NMFS et al. 2011; Gallaway et al. 2013, 2016a, b; NMFS & USFWS 2015; Kocmoud et al. 2019; Ramirez 2019). Theoretical papers by Schröder et al. (2014) and DeRoos (2018) discuss juvenile versus adult abundances and their effects on population dynamics.

Translocation of nests to on-beach hatcheries is considered highly manipulative (Meylan & Ehrenfeld 2000), but it was necessary, in combination with conservation interventions that reduced at-sea mortality of neritic life stages, to prevent Kemp’s ridley’s extinction and to put this species on a course toward recovery (Marquez-M. 1994; Heppell et al. 2005, 2007; Márquez-M. et al. 2005; Gallaway et al. 2013, 2016a, b; Márquez-Millán & Garduño-Dionate 2014; Burchfield & Peña 2015; Caillouet et al. 2015, 2016a; Kocmoud et al. 2019; Wibbels & Bevan 2019). Egg-to-hatchling survival is lower for nests left in situ, even when in situ nests are protected in various ways (Marquez M. 1987; Pritchard 1990, 2007; TKRRT 1992; Marquez-M. 1994; Márquez et al. 1999; Márquez-M. et al. 2005; Bevan et al. 2014, 2016; Márquez-Millán & Garduño-Dionate 2014; Burchfield & Peña 2015; Burchfield et al. 2020).

Pritchard (2007) questioned whether “the more turtles the better” conservation philosophy applied to Kemp’s ridleys on Tamaulipas beaches should be abandoned. By 2004, the annual number of nests had increased to levels exceeding capabilities to translocate most of them to on-beach hatcheries (Bevan et al. 2014; Caillouet et al. 2016a; Gallaway et al. 2016a, b; Kocmoud et al. 2019). Therefore, a decision was made to reduce numbers of nests translocated to on-beach hatcheries and thus increase annual numbers of nests left in situ. However, in situ nests have continued to be protected in various ways on Tamaulipas beaches (Burchfield et al. 2020) and annual hatchling production has not been substantially reduced.

Arribada nesting on Tamaulipas beaches is the biogeographical norm for Kemp’s ridley (Hildebrand 1963, 1982; Pritchard 2007; Wibbels & Bevan 2019). In the distant past, Kemp’s ridley arribadas overwhelmed predators with ephemeral oversupplies of food, thereby perpetuating the species (Pritchard 2007). However, arribada nesting was disrupted primarily by exploitation of eggs on Tamaulipas beaches and mortality in neritic juveniles and adults caught unintentionally in shrimp trawls (Carr 1963, 1967, 1977; Hildebrand 1963; Marquez-M. 1994; Heppell et al. 2005, 2007; Gallaway et al. 2013, 2016a, b; Márquez-Millán & Garduño- Dionate 2014; Burchfield & Peña 2015; Caillouet et al. 2015, 2016a; Kocmoud et al. 2019; Wibbels & Bevan 2019).

Kemp’s ridley population status and trends have been measured by annual numbers of nests (N_t, where t is calendar year) and hatchlings released (h_t) into the GoM from the Tamaulipas index beach (Rancho Nuevo, Tepehuajes and Playa Dos beach segments combined) (Fig. 1; NMFS et al. 2011; NMFS & USFWS 2015). I emphasize that N_t and h_t comprise most but not all of the nests and hatchlings documented annually on Tamaulipas beaches (Heppell et al. 2007; Burchfield et al. 2020). The US-Mexico recovery plan (NMFS et al. 2011; NMFS & USFWS 2015) provided N_t and h_t thresholds for downlisting Kemp’s ridley from endangered to threatened status; viz., N_t = 25,000 nests (equivalent to 10,000 adult females nesting in a season) and h_t = 300,000 hatchlings released in a season (Fig.1). The downlisting threshold for h_t was exceeded during 2000-2020, except for 2001 when it was 291,268, while N_t remained below its downlisting threshold (Fig. 1).

Figure 1. Trends in Log₁₀-transformed H_t, h_t, N_t and h_t/N_t (where t = calendar year) on the index beach, Tamaulipas, Mexico, 1966-2020, compared to Log₁₀-transformed downlisting thresholds for h_t (horizontal dotted line) and N_t (horizontal dashed line).

Figure 2. Trend in finite multiplication rate, N_t/N_t-1, for the Kemp’s ridley index beach, Tamaulipas, Mexico, 1967-2020. The horizontal dashed represents N_t = N_t-1 (no change between consecutive years t-1 and t). Values of N_t / N_t-1) above the horizontal dashed line indicate increases (N_t > N_t-1), and those below the line indicate decreases (N_t) < N_t-1).

Also shown in Fig. 1 are trends in two derived variables; viz., cumulative numbers of hatchlings released (H_t; Caillouet et al. 2016a) and numbers of hatchlings released per nest (h_t/N_t; Caillouet 2014). The variable H_t reflects total numbers of hatchling ever released from Tamaulipas beaches, beginning in 1966. The variable h_t/N_t reflects annual fecundity of nesters and hatch rates, which are influenced by many factors (Caillouet 2014; Caillouet et al. 2016a). In any year, ht is determined for the most part by N_t, but it has also been affected by the post-1989 decline in h_t/N_t (Caillouet 2014; Caillouet et al. 2016a). Fecundity of nesters declined as the annual proportion of neophyte (first time) nesters increased (Marquez-M. 1994; Heppell et al. 2005, 2007; Witzell et al. 2005; Caillouet 2014; Caillouet et al. 2016a, 2018; Shaver et al. 2016b), and this may have contributed to the decline in h_t/N_t. Intentional increases in numbers of nests left in situ (Bevan et al. 2014) also could have contributed to the post-2003 decline in h_t/N_t. A mark-recapture study of Tamaulipas nesters during 2014-2015 found that 86% were putative neophytes (Burchfield & Peña 2015).

For years 1986–2014, Caillouet et al. (2016a) detected pre-2010 slowing of rates of increase in (1) the relationship between N_t and H_t-10, and (2) the times series of N_t/H_t-10. Caillouet et al. (2018) detected pre-2010 slowing of the rate of increase in N_t (Fig. 1). The finite multiplication rate (N_t/N_t-1) reached a temporary peak in 2000 (Fig. 2; see also Fig. 1B in Caillouet et al. 2018) and its maximum level in 2020 (Fig. 2). Assuming 10 yrs to maturity, its most recent surge may be a response to the 2009 hatchling release (indexed by h_t = 1,025,027), which was the highest on record (Fig. 1). This recent surge may also provide optimism that population growth has resumed; however, the highest N_t within the 1966-2020 time series was 22,415 in 2017, which is 4,239 (23%) higher than its 18,176 level in 2020. Only time will tell whether the nesting setback has ended.

Five years before the Deepwater Horizon (DWH) oil spill occurred in the northern GoM, Heppell et al. (2005) raised concerns that carrying capacity had changed and could prevent Kemp’s ridley from reaching original levels. In 2006, Peter C.H. Pritchard suggested that carrying capacity might be exceeded because of intensive conservation efforts applied over the years (Caillouet 2014). GoM ecosystem alteration and degradation were underway long before the DWH oil spill (Heppell et al. 2007; Jackson 2008; Peterson et al. 2011; Walker et al. 2012; Yasuhara et al. 2012; Karnauskas et al. 2013; Shepard et al. 2013; Benitez et al. 2014; DWH NRDA Trustees 2016; Davis 2017; Hu et al. 2017; Scavia et al. 2017; Ward 2017; Wallace et al. 2020). Gallaway et al. (2013) mentioned the possibility that the assumption of density-independent mortality in age-structured modeling of the Kemp’s ridley population may no longer be valid due to limits imposed by carrying capacity, but Gallaway et al. (2016b) considered density-dependent mortality unlikely for benthic-stage (neritic) Kemp’s ridleys. Kocmoud et al. (2019) suggested that environmental factors caused the remigration interval for nesting females to increase. Avens et al. (2017, 2020) and Ramirez et al. (2020, 2021) compared Kemp’s ridleys in the GoM and western North Atlantic Ocean with regard to age, growth, and maturity as related to environmental factors.

If age-structured modeling shows abundance of neritic immatures to be excessive, then consideration should be given to translocating excess clutches from Tamaulipas to other beaches throughout the northern GoM, to bolster the existing nesting colony on the coast of Padre Island National Seashore, and to establish new ones. Nesting on GoM beaches north and east of Tamaulipas, and along the eastern coast of North America may eventually become more important to Kemp’s ridley population growth, recovery, resiliency, diversity and sustainability as climate warms and sea level rises (Heppell et al. 2007; Poloczanska et al. 2009; Putman et al. 2010a, b; Caillouet 2012, 2019; Pike 2013a, b; Shaver et al. 2013; 2016a, b; Caillouet et al. 2016b, 2018; Bevan et al. 2019; Butler 2019; Fuentes et al. 2019; Griffin et al. 2019; Innis et al. 2019; Reid et al. 2019; Dubois et al. 2020). However, Kemp’s ridley may not be capable of adjusting rapidly enough to climate warming and sea level rise because of its fidelity to reproducing predominantly along the Tamaulipas coast (ibid.). Currently, it is unlikely that many if any Kemp’s ridley hatchlings that enter the western NAO from rare nestings on the US east coast survive (Ramirez, M.D., pers. comm.; Caillouet & Gallaway 2020). Coastal currents and configurations and widths of continental shelves of the GoM and western NAO also influence locations of Kemp’s ridley reproductive and foraging areas (Carr 1980; Rudloe & Rudloe 2005; Putman et al. 2010a, b; Shaver et al. 2013, 2016b; Caillouet & Gallaway 2020). In addition, river inflows (especially that of the Mississippi River) are greater along the GoM coast than along the east coast of North America, and they are essential to sustaining coastal estuaries that support life cycles of key Kemp’s ridley prey species such as blue crab (Callinectes sapidus) (Hildebrand 1982; Vanderkooy 2013; Perry & Vanderkooy 2015; Gallaway et al. 2016b; O’Connell et al. 2019). In addition, restoration of the GoM ecosystem should increase carrying capacity for the Kemp’s ridley population (Caillouet et al. 2018; Caillouet 2019).

Kemp’s ridley’s largest documented single-day arribada occurred on 18 June 1947, and it has been adopted as a benchmark for this species’ recovery (Bevan et al. 2016; Wibbels & Bevan 2019). Therefore, consideration should be given by CONANP, USFWS and NMFS to examining existing daily nest counts during 1966-2020 to find the largest single-day nest count in each of those seasons. The trend in largest single-day nest counts would be informative as an index of single-day arribada size and progress toward recovery. My guess is that it would show the Kemp’s ridley population to be far from recovery, even though its downlisting criterion for females nesting in a season has been approached, while that for hatchlings has been exceeded in 20 of the last 21 years (Fig. 1).

Acknowledgements. This review would not have been possible without availability of N_t and h_t data pairs for 1966-2020, which are maintained by CONANP and Gladys Porter Zoo, Brownsville, Texas. N_t and h_t data pairs for 1966-2014 were obtained from NMFS & USFWS (2015), and those for 2015-2020 were obtained from Jaime Peña and Patrick Burchfield, Gladys Porter Zoo. I thank Alberto Abreu-Grobois, Larisa Avens, Benny Gallaway, Nathan Putman, and Matthew Ramirez, as well as anonymous peer reviewers and Matthew Godfrey for reviewing various drafts of the manuscript. I commend CONANP and its predecessor agencies, National Park Service, USFWS, NMFS and Texas Parks and Wildlife Department, as well as other federal, state and local government agencies, corporations, businesses, universities, conservation organizations, local communities, and individuals in Mexico and the US, for their roles in Kemp’s ridley conservation. This review is dedicated to the memories of Peter C.H. Pritchard (who died on 25 February 2020), Henry H. Hildebrand, Archie F. Carr, and Andrés Herrera. It is also dedicated to René Márquez-Millán.

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