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Herrera’s movie also showed men exploiting eggs (Hildebrand 1963; Carr 1967). However, seven decades before the 1947 arribada, Prieto (1873) reported that marine turtles and their eggs contributed to the commerce of Tamaulipas. In the early 1920s, Mexico’s federal government began promulgating laws, regulations, and acts aimed at reducing harvest of sea turtles and their eggs on land and at sea (Marquez-M. 1994; Márquez et al. 1998; Márquez-M. et al. 2018). Despite such measures, the Kemp’s ridley population declined substantially and was still declining when Hildebrand (1963) urged promulgation of conservation measures to prevent extinction of this species.

In 1966, Mexico’s federal government initiated on-beach patrols and annual protection of as many nesting females, nests, and hatchlings as possible near Rancho Nuevo (Chavez et al. 1968; Heppell et al. 2005, 2007; Márquez-M. et al. 2005, 2018; Pritchard 2007; Márquez-Millán et al. 2014). At the same time, Mexico’s federal government initiated (1) an annual count of nests (N_t, where t is calendar year), which provided an index of annual abundance of nesting females, and (2) a corresponding annual count of hatchlings (h_t) released into the GoM, which provided an index of annual reproductive output of these nesting females (Caillouet et al. 2015b, 2016, 2018; Wibbels & Bevan 2016). Annual production of hatchlings (h_t) was being restored, but annual nests (N_t) continued to decline (Fig. 1), because not enough time had elapsed for the new recruits to reach maturity (Marquez-M. 1994). Carr (1977) called for action to save Kemp’s ridley from extinction, noting that the preceding decline in the population was caused by overexploitation of eggs combined with heavy natural predation pressures, but the decline in progress was brought about by incidental capture in shrimp trawls.

Figure 1. Trends in Kemp’s ridley H_t, h_t, and N_t, where t is calendar year, H_t is cumulative annual number of hatchlings released, h_t is annual number of hatchlings released, and N_t is annual number of nests (i.e., clutches of eggs laid) on the Tamaulipas, Mexico nester-abundance-index beach during 1966-2018. Horizontal dotted lines represent annual minima, h_t (300,000) and N_t (25000 ≈10,000 adult females), for downlisting Kemp’s ridley to threatened status (see NMFS et al. 2011).

In 1978, the US-Mexico Kemp’s ridley restoration and enhancement program was initiated (Márquez Millan et al. 1989; Márquez-M. et al. 2005, 2018; Pritchard 2007; Márquez-Millán et al. 2014; Caillouet et al. 2015b). The population continued declining to near extinction by 1985 (Fig. 1; Byles 1993). During the 1947-1985 population decline, magnitudes of ecological roles in aquatic and terrestrial habitats (Bjorndal & Bolten 2003; see review by Lovich et al. 2018) fulfilled by Kemp’s ridley no doubt had diminished substantially, and the GoM ecosystem likely adjusted to declining abundance of all life stages.

In any analysis or modeling of trends in N_t and h_t (Fig. 1), consideration should be given to the intermittent increases in length of the Tamaulipas nester-abundance-index beach over the years, from that of Rancho Nuevo exclusively, to the maximum comprising Rancho Nuevo, Tepehuajes and Playa Dos combined (see map Figure 1 in Márquez et al. 1999; Turtle Expert Working Group [TEWG] 1998, 2000; Heppell et al. 2005, 2007; Márquez-M. et al. 2005; NMFS et al. 2011; Márquez-Millán et al. 2014). Also, during 1966-1977, most nests found and counted (N_t) were left in situ, thus clutches of eggs that were translocated to protective, on-beach hatcheries represented small proportions of N_t; the counts of hatchlings released (h_t) in those years originated only from clutches that were translocated and protected (TEWG 1998, 2000; Márquez et al. 1999; Márquez M. 2001; Márquez-M. et al. 2005). During 1978-2018, eggs from all nests found were translocated to protective, on-beach hatcheries, except for those deliberately left in situ (TEWG 1998, 2000; NMFS et al. 2011; Gallaway et al. 2013, 2016a; Caillouet et al. 2016), because either they exceeded the capacity of on-beach hatcheries, or it became logistically impossible to translocate all eggs (Bevan et al. 2014). A robust examination of archived records could be helpful in evaluating levels and efficacy of monitoring N_t and h_t over the years.

Cumulative beneficial effects of conservation interventions that reduced mortality on Tamaulipas beaches and at sea, combined with other factors, reversed the decline in N_t by 1986 (Fig. 1; Byles 1993; Marquez-M. 1994; Caillouet 2010; Caillouet et al. 2016), and led to rapid increase in N_t to 19,361 by 2009 (Crowder & Heppell 2011; NMFS et al. 2011; Gallaway et al. 2013, 2016a, 2016b; Caillouet 2014; Caillouet et al. 2015b, 2016, 2018; Dixon & Heppell 2015; NMFS & USFWS 2015; Mazaris et al. 2017; Kocmoud et al. 2019). The other factors were those that contributed to reduction in mortality associated with shrimp trawling in GoM waters, including the 1976 US-Mexico treaty that phased out US shrimp trawling in Mexico’s GoM waters by 1980, the seasonal Texas Closure to shrimping that began in 1981, the use of turtle excluder devices (TEDs) first required by US federal regulations initiated in 1987 and expanded thereafter, hurricane damage to GoM shrimp trawlers and processing facilities, and deteriorating economic conditions within the GoM shrimp industry (Condrey & Fuller 1992; Iversen et al. 1993; Lewison et al. 2003, 2013; Caillouet et al. 2008, 2016; Nance et al. 2008, 2010; Gallaway et al. 2013, 2016a, 2016b). Kemp’s ridley mortality in all life stages, except the oceanic stage, had been greatly reduced (Márquez Millan et al. 1989; Kemp’s Ridley Recovery Team 1992; Lewison et al. 2003, 2013; Heppell et al. 2005, 2007; Márquez-M. et al. 2005, 2018; Crowder & Heppell 2011; Finkbeiner et al. 2011; NMFS et al. 2011; Márquez-Millán et al. 2014; NMFS & USFWS 2015; Caillouet et al. 2016; Valdivia & Suckling 2019). Post-1985 increases in N_t and h_t through 2009 suggested that all Kemp’s ridley life stages had increased in abundance, and that their ecological roles and contributions to biodiversity and population resilience within GoM ecosystem were being restored.

The US-Mexico recovery plan (NMFS et al. 2011) established the following demographic criteria for downlisting Kemp’s ridley status according to the US Endangered Species Act from endangered to threatened status: at least 10,000 females (≈ 25,000 nests) nesting in a season on the nester-abundance-index beach, and at least 300,000 hatchlings released annually from that beach. Horizontal dotted lines in Fig. 1 depict these downlisting thresholds for h_t and N_t, and vertical dashed lines connect the points for h_t and N_t in 2000 and 2010. The recovery plan’s population model predicted that these thresholds would be reached by 2011, and that N_t would continue increasing at a rate 19% per year through 2020, assuming that survival rates within each life stage remained constant (NMFS et al. 2011). However, this rapid increase in N_t was interrupted in 2010, the year in which the Deepwater Horizon (DWH) oil spill occurred in the northern GoM (Bjorndal et al. 2011; Caillouet 2011; Crowder & Heppell 2011; Gallaway et al. 2013). Kemp’s ridley strandings increased in the northern GoM during 2010 and 2011, and the DWH oil spill and shrimp trawling received the most attention as possible causes (Caillouet 2011; Gallaway et al. 2013). During 2010-2018, N_t ranged 10,987-22,415 (Fig. 1), with its lowest in 2014 and highest in 2017, all of which were below predicted levels (Caillouet 2014; Caillouet et al. 2015b, 2016, 2018; Crowder & Heppell 2011; Dixon & Heppell 2015; Gallaway et al. 2013, 2016a, 2016b; Kocmoud et al. 2019; NMFS et al. 2011; NMFS & USFWS 2015). This represented a major setback in Kemp’s ridley nesting (Caillouet et al. 2016). Deepwater Horizon Natural Resource Damage Assessment Trustees (2016) concluded that the oil spill was unlikely to have had a direct impact on Kemp’s ridley nesting in 2010, but could have contributed to reduced numbers of nests in subsequent years through direct and indirect pathways. Gallaway et al. (2016b) estimated there were 61,330 Kemp’s ridley deaths in 2010, of which they attributed 5% to incidental mortality in shrimp trawls, 19% to natural causes of mortality, and 76% to undetermined anthropogenic causes of mortality other than shrimp trawling. Various hypotheses have been put forward to explain the 2010-2018 nesting setback, but none have been confirmed with certainty (Gallaway et al. 2016a, 2016b; Caillouet et al. 2018; Kocmoud et al. 2019). Despite the nesting setback, h_t ranged 291,268-1,025,027 during 2000-2018, exceeding the 300,000-hatchling threshold established by NMFS et al. (2011) in all but one (2001) of the last 19 years (Fig. 1). Maintenance of such high levels of h_t for almost 2 decades demonstrates the dedication of Mexico to its on-beach conservation interventions in Tamaulipas and their efficacy.

Neritic immature and adult sea turtles are subject to compensatory density-dependent functions (National Research Council 2010). Lowered per capita availability of food for neritic Kemp’s ridleys can reduce somatic growth rates, increase age at sexual maturity (ASM), and reduce body condition of adults and their ability to migrate to nesting beaches; it can reduce the ability of adult females to develop eggs and nest, as well as increase inter-nesting and remigration intervals (Bjorndal et al. 2014; Caillouet 2014; Caillouet et al. 2016, 2018; Gallaway et al. 2016b; Avens et al. 2017; Omeyer et al. 2017; Craven et al. 2019; Kocmoud et al. 2019). It is unlikely that oceanic stage Kemp’s ridleys compete with conspecific neritic immatures or adults for food and other resources, but likely that neritic immatures and adults do compete for food and other resources. It is unlikely that availability of nesting beaches has limited Kemp’s ridley population growth; as the population increased, nesting spread within and beyond Tamaulipas, arribada size increased on beaches of Tamaulipas and elsewhere, and arribada nesting is the norm (Márquez et al. 1999; Jiménez-Quiroz et al. 2003; Heppell et al. 2005, 2007; Márquez-M. et al. 2005; NMFS et al. 2011; Márquez-Millán et al. 2014). However, average h_t/N_t has been declining since it peaked in 1989, and its cause has not been determined (Caillouet 2014; Caillouet et al. 2016).

Density independence has been assumed in most modeling of the Kemp’s ridley population (TEWG 1998, 2000; NMFS et al. 2011; Heppell et al. 2005, 2007; Coyne & Landry 2007; Crowder & Heppell 2011; Gallaway et al. 2013, 2016a; Dixon & Heppell 2015; NMFS & USFWS 2015). However, density-dependent effects on various vital (demographic) rates were evident before 2010 (Heppell et al. 2005, 2007; Caillouet 2014; Caillouet et al. 2016, 2018; Gallaway et al. 2016b; Shaver et al. 2016; Avens et al. 2017). The regression model applied by Caillouet et al. (2018) showed that density dependence affected N_t before 1962 and after 2004, with an interval of density-independent changes in N_t in between. Caillouet et al. (2018) hypothesized that slowing of the rate of increase in N_t after 2004 was caused by a combination of declining carrying capacity for Kemp’s ridleys due to degradation of the GoM ecosystem, exponential growth of the population, and declining per capita availability of food for neritic immatures and adults, including natural prey and scavenged discarded bycatch from shrimp trawling. Factors that could have contributed to declining per capita availability of food included intraspecific competition among neritic immatures and adults, their interspecific competition with other marine predators and scavengers, effects of fisheries for crabs, and reductions in discarded bycatch from the shrimp fishery (Gallaway et al. 2016b; Avens et al. 2017; Caillouet et al. 2018; Craven et al. 2019; Kocmoud et al. 2019). The more abundant loggerhead sea turtle (Caretta caretta) may also compete for food with Kemp’s ridley (Hart et al. 2018; Lamont & Iverson 2018).

Caillouet et al. (2018) prompted my examination herein of four novel and simple rates of change calculated from N_t, h_t, and H_t, where H_t is the cumulative annual count of hatchlings released from the nester-abundance-index beach (see Figure 7 in Caillouet et al. 2016), to determine whether these rates exhibited pre-2010 evidence of density dependence. The range in t for these calculations was 1966-2018. Because assumed ASM affects results of Kemp’s ridley population models, I incorporated three different values (8, 10, and 12 years) for M (minimum ASM) in calculating some of these rates. The range in published estimates of ASM for wild Kemp’s ridleys is 6.8-21.8 years (Snover et al. 2007; NMFS et al. 2011; Avens et al. 2017). Models applied by TEWG (1998) and Heppell et al. (2005) incorporated assumed ASMs of 8, 10, and 12 years. Each value of M was assumed constant over t, as is usually the case with ASM in various models (but see the review by Bernardo 1993). In modeling, increasing the value of ASM increases the number of cohorts of neritic immatures in the estimated population, because the number of cohorts in the oceanic stage is typically held constant (TEWG 1998, 2000; Heppell et al. 2005, 2007; Coyne & Landry 2007; Crowder & Heppell 2011; NMFS et al. 2011; Gallaway et al. 2013, 2016a, 2016b; Kocmoud et al. 2019).

The four rates were:

Figure 2. Trends in Kemp’s ridley N_t /N_t-M, where t is calendar year, N_t is the annual number of nests (clutches of eggs laid) on the Tamaulipas, Mexico nester-abundance-index beach during 1974-2018, 1976-2018, and 1978-2018, for assumed minimum age at sexual maturity, M, of 8 years, 10 years, and 12 years, respectively. Horizontal dotted lines represent N_t /N_t-M=1.

Regardless of assumed M, 2000 and 2009 were pivotal years for the trend in N_t / N_t-M, as shown by substantial slowing of its trends after 2000 and again after 2009. Early values of N_t / N_t-M were below 1 for the three values of M (Fig. 2). Starting values of N_t /N_t-M were 0.244 for an M of 8 years, 0.185 for M=10 years, and 0.154 for M=12 years. Through 2000, N_t /N_t-M increased least rapidly for M=8 years, more rapidly for M=10 years, and most rapidly for M=12 years, because increases in M postponed the starting points for N_t /N_t-M, thereby shortening the interval between starting years and 2000. Caillouet et al. (2018) plotted the time series of N_t /N_t-1, referring to it as the finite multiplication rate. Its highest level occurred in 2000 and its lowest in 2010 (Figure 1 B in Caillouet et al. 2018). The plot of residuals for the demographic model applied in the recovery plan (Figure 5 in NMFS et al. 2011) provided evidence of an inflection point in growth of N_t in 2000 when the largest positive residual occurred.

All values of N_t /H_t-M were below 1 (Fig. 3), because H_t was so much larger than N_t throughout the time series (Fig. 1), as expected. Starting values of the trends in N_t /H_t-M were highest (0.0480) for M=8 years, intermediate (0.0363) for M=10 years, and lowest (0.0302) for M=12 years. All trends in N_t /H_t-M were steeply downward during the pre-1986 population decline. The pre-2010 minimum N_t /H_t-M was 0.00168 in 1989 for M=8 years, 0.00206 in 1989 for M=10 years, and 0.00252 in 1993 for M=12 years. Regardless of assumed M, 2000 and 2009 were pivotal years for the trends in N_t /H_t-M. Caillouet et al. (2016) were the first to plot N_t /H_t-M, limiting it to M=10 (see their Figure 7).

Figure 3. Trends in Kemp’s ridley N_t /H_t-M, where t is calendar year, N_t is annual number of nests (clutches of eggs laid) on the Tamaulipas, Mexico nester-abundance-index beach during 1974-2018, 1976-2018, and 1978-2018, for assumed minimum age at sexual maturity, M, of 8 years, 10 years, and 12 years, respectively, and H_t-M is cumulative annual number of hatchlings released on the nester-abundance-index beach during 1966-2010, 1966-2008, and 1966-2006, respectively.

Figure 4. Trend in Kemp’s ridley h_t /h_t-1, where t is calendar year, h_t is annual number of hatchlings released on the Tamaulipas, Mexico nester-abundance-index beach during 1967-2018. The horizontal dotted line represents h_t /h_t-1=1.

Variation in values of h_t / h_t-1 was relatively wide during the first ≈ 2 decades, then narrowed through 2000 (Fig. 4). Interestingly, the highest value of h_t /h_t-1 occurred in 1976, prior to the beginning of the US-Mexico Kemp’s ridley restoration and enhancement program. Years 2000 and 2009 were pivotal for h_t / h_t-1, with h_t / h_t-1 exhibiting a general decline with increased variability after 2000.

Starting values of the trends in h_t / H_t-M were lowest (0.81) for M=8 years, intermediate (1.18) for an M=10 years, and highest (1.57) for M=12 years (Fig. 5). For all values of M, the trends in h_t /H_t-M were downward as the population declined; the downward trend of h_t /H_t-M was least steep for an M=8 years, intermediate for M=10 years, and most steep for M=12 years. Interestingly, the pre-2010 minimum h_t /H_t-M was 0.00168 in 1989 for M=8 years, 0.00206 in 1989 for M=10 years, and 0.00252 in 1993 for M=12 years, although minimum N_t occurred in 1985. Regardless of assumed M, 2000 and 2009 were pivotal years for the trend in h_t /H_t-M , with drops in 2010 marking the beginning of the nesting setback.

Figure 5. Trends in Kemp’s ridley h_t/H_t-M, where t is calendar year, h_t is annual number of hatchlings released on the Tamaulipas, Mexico nester-abundance-index beach during 1974-2018, 1976-2018, and 1978-2018, for assumed minimum age at sexual maturity, M, of 8 years, 10 years, and 12 years, respectively, and H_t-M is cumulative annual number of hatchlings released on the nester-abundance-index beach during 1966-2010, 1966-2008, and 1966-2006, respectively. Horizontal dotted lines represent h_t/H_t-M=1.

Given that density dependence appears to have begun reducing the rate of growth of the Kemp’s ridley population around year 2000, I hypothesize that the annual number of neritic immatures became excessive around that year. I recommend that age-structured modeling that incorporates estimates of annual mortality attributable to shrimp trawling in the GoM (e.g., Gallaway et al. 2013, 2016a, 2016b) be conducted to estimate annual numbers of adults and neritic immatures in the population during 1985-2018. For such modeling, data covering Mexico’s shrimp trawling in the GoM should be acquired and combined with data covering US shrimp trawling in the GoM, for purposes of estimating total annual Kemp’s ridley mortality attributable to shrimp trawling within the entire GoM. The 1985-2018 trend in the annual quotient calculated by dividing estimated annual number of adults by estimated annual number of neritic immatures can then be examined. If this annual quotient initially increased then later declined, the decline would suggest density-dependent limitation of population growth and show when it began developing. This approach is consistent with earlier modeling that estimated the potential of experimental reintroduction of Kemp’s ridley nesting to Padre Island National Seashore and use of TEDs to contribute to Kemp’s ridley population growth (Caillouet et al. 2015b; Heppell et al. 1996, 2005, 2007; Heppell & Crowder 1998; NMFS et al. 2011; NMFS & USFWS 2015; Shaver & Caillouet 2015; TEWG 1998, 2000). I welcome the application of other types of age-structured modeling to estimate annual numbers of adults and neritic immatures for use in calculating the suggested quotient and its trend beginning with 1985.

Pritchard (2007) pondered the possibility that the unstated goal of producing “as many turtles as possible” should be abandoned, “not only because natural population constraints will eventually be felt on the feeding grounds but also because there is almost certainly some level of density of an arribada at which the sheer number of turtles is counterproductive, leading to degradation of the beach and massive, although accidental, destruction of eggs laid by previous nesters”. Natural population constraints on the feeding grounds seem more likely to have begun limiting the Kemp’s ridley population’s growth rate than arribada density, because most nests have been protected in on-beach hatcheries, beginning in 1978 (Caillouet 2006).

In all modeling of Kemp’s ridley population dynamics to date, additions from immigration and losses from emigration have been ignored. However, the proportion of the population retained within the GoM is much greater than that in the Atlantic (NMFS et al. 2011; Putman et al. 2013; NMFS & USFWS 2015). I assume that Kemp’s ridley immigration represents the return of neritic stage turtles from the North Atlantic Ocean (NOA) to the GoM, and emigration represents transport of oceanic stage turtles into the NAO combined with movement of neritic stage turtles from the GoM to the NAO. Migration distances from the NAO to western GoM nesting beaches are longer than those from within the GoM. Annually, the oceanic stage is much more abundant than the neritic stage, so losses to the NAO likely exceed gains by the GoM. The total number of GoM tag returns for Kemp’s ridleys tagged along the US east coast is low, although most of them were documented for Tamaulipas nesters (Caillouet et al. 2015b). It is time for future Kemp’s ridley population models to incorporate metrics of emigration and immigration, based on dispersal in the oceanic stage and examinations of available catch-mark-recapture and tracking data for neritic stage turtles, to determine whether there is a net loss to the NAO, and if so to estimate its magnitude. Notwithstanding possible net loss from the population through emigration into the NAO, nestings on the US east coast appear to be increasing, and may someday reach levels important to sustaining the population, just as nestings in the GoM in locations other than Tamaulipas provide “safety nets” for the species.

Heppell et al. (2007) and Wibbels & Bevan (2016) suggested that demographic criteria for delisting Kemp’s ridley may be unachievable. Consideration should be given to adding a recovery criterion related to achieving an annual population age-structure similar to that in 1947; i.e., one with a much higher proportion of adult females than currently exists (Caillouet et al. 2018). A challenge is that the effects of any changes in conservation interventions will not be detectable via the N_t metric for approximately a decade. If reduced GoM carrying capacity for Kemp’s ridleys is currently the dominant factor limiting population growth, then ongoing efforts to restore the GoM ecosystem may mitigate its effects (see Caillouet et al. 2018; National Academies of Sciences, Engineering, and Medicine 2017; Peterson et al. 2011). In the interim, continued monitoring of N_t and h_t will make it possible to observe the effects of status quo maintaining or increasing h_t (NMFS et al. 2011). Consistent with this status quo approach was the suggestion by Caillouet et al. (2016) that the most expedient way to restore Kemp’s ridley population growth toward recovery would be to translocate more clutches to protective corrals, leaving fewer in situ, but I would recommend against it at this time.

If annual numbers of neritic immatures in the population are already excessive and preventing population recovery as defined by NMFS et al. (2011), it would seem prudent to begin reducing numbers of neritic immatures by reducing annual numbers of hatchlings released from Tamaulipas beaches. This could be achieved by leaving more clutches in situ without protection (NMFS et al. 2011; TEWG 1998, 2000). This could free some of the personnel and resources now devoted to collecting, translocating, and protecting clutches, to focus on researching (1) past, present, and future annual proportions of putative neophyte nesting females (Caillouet 2014), (2) cause(s) of the post-1989 decline in h_t /N_t (Caillouet 2014), (3) past, present, and future annual carapace length-frequency distributions of nesting females (Caillouet 2014), (4) past, present, and future health and body condition of nesting females (5) past, present, and future remigration intervals (via catch-mark-recapture) of nesting females (Gallaway et al. 2016b; Kocmoud et al. 2019), (6) sampling methods that ensure accurate counts of nesting females (NMFS et al. 2011; Rees et al. 2018) and hatchlings released, (7) life-long tags or marks for mass-tagging cohorts of hatchlings (Caillouet & Higgins 2015), (8) past, present, and future annual sex ratios of hatchlings, and (9) detection of tags or marks on nesting females (external and internal). Consideration should also be given to updating and modifying the bi-national recovery plan (Caillouet 2006; Caillouet et al. 2015a), including the demographic criteria for downlisting and delisting. Whether or not the analyses and modeling recommended herein are conducted, continued on-beach conservation interventions (at a level to be determined) and monitoring on the coast of Tamaulipas are essential to Kemp’s ridley population recovery within the GoM, and they are required to maintain and enhance the secondary and tertiary nesting colonies in Veracruz and Texas that contribute to the population’s diversity and resilience (NMFS et al. 2011; NMFS & USFWS 2015; Tecolutla Turtle Project 2018).

Acknowledgements. I am grateful that Mexico’s Comisión Nacional de Áreas Naturales Protegidas (CONANP) made N_t and h_t data pairs for 1966-2018 available. The data pairs for 1966-2014 were obtained from NMFS & USFWS (2015), and those for 2015-2018 were obtained from Jaime Peña, Gladys Porter Zoo, Texas, via annual reports for the “Mexico/United States of America population restoration project for the Kemp’s ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas Mexico.” Special thanks are due to Nathan F. Putman, Benny J. Gallaway, William E. Grant, and my daughter Theresa E. Caillouet, who reviewed various versions of the manuscript and offered helpful comments. MTN editor, Matthew H. Godfrey, and two anonymous peer reviewers also made helpful comments. I commend all who have participated in Kemp’s ridley population recovery efforts in Mexico and the US, including those in federal, state, and local government agencies, corporations, businesses, universities, conservation organizations, and communities (volunteers). I dedicate this review to Peter C.H. Pritchard and René Márquez-Millán, and to the memories of Henry H. Hildebrand, Archie F. Carr, and Andrés Herrera.

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