|Scientific Name:||Limulus polyphemus|
|Species Authority:||(Linnaeus, 1758)|
Monoculus polyphemus Linnaeus, 1758
|Red List Category & Criteria:||Vulnerable A3bd ver 3.1|
|Assessor(s):||Smith, D.R., Beekey, M.A., Brockmann, H.J., King, T.L., Millard, M.J. & Zaldívar-Rae, J.A.|
|Reviewer(s):||Shin, P., Botton, M.L., Carmichael, R., Dharmarajah, V., Ho, B., Ling, D.J., Novitsky, T. & Tanacredi, J.|
Conceptually, the Horseshoe Crab assessment integrates information on management and conservation actions, major threats, habitat and population responses, and population genetic structure (Figure 2 in the Supplementary Material). The population responses, in terms of abundance, geographic range, and viability, along with genetic structure inform risk at the regional level, which in turn, informs the assessment of species' extinction risk (Figure 2 in the Supplementary Material).
|Previously published Red List assessments:|
|Range Description:||Horseshoe Crabs have persisted for more than 200 million years (Botton and Ropes 1987, Shuster 2001, Tanacredi 2001, Anderson and Shuster 2003) and unmistakable fossil forms of Horseshoe Crabs have been found as far back as 450 million years ago (Sekiguchi 1988, Rudkin and Young 2009). Today Horseshoe Crabs are found in two regions of the world. Three species, Tachypleus tridentatus (Leach, 1819), T. gigas (O. F. Müller, 1785), and Carcinoscorpius rotundicauda (Latreille, 1802), inhabit the coastal waters of Asia from India to Japan, including the East Indies and Philippines. Limulus polyphemus, the “American” Horseshoe Crab, is found along the Atlantic coastline of North America ranging from the Yucatán Peninsula, México (19°N) to the Gulf of Maine, USA (42°N) (Figure 1 in the supplementary material). Six subpopulations are recognized, determined by genetic population structure: Gulf of Maine (USA), Mid-Atlantic (USA), Southeast (USA), Florida Atlantic (USA), Northeast Gulf of México (USA), and Yucatán Peninsula (México).|
Atlantic Coast of North America
Horseshoe Crabs occur along the Atlantic coast of the US, from Maine to Florida. The centre of distribution and the highest densities occur in the Delaware Bay region.
Northern extent. The distribution of Horseshoe Crabs extends to the north along the mid-Atlantic and northeastern states, and Horseshoe Crabs have been observed as far north as Cobscook Bay, Maine. However, recent efforts to locate breeding populations (Schaller 2002, Schaller et al. 2005) reveal that the northern extent of the species is Frenchman Bay east of Mt. Desert Island, Maine (Frings and Frings 1953, Kingsley 1901, Moore and Perrin 2007). Records exist of Horseshoe Crabs in Nova Scotia including one living specimen from Lahave Island, southwest of Halifax, Nova Scotia (Wolff 1977), but breeding populations are not known to occur currently in Canadian waters.
Southern extent. Horseshoe Crabs have been observed nesting in all coastal counties along the east coast of Florida and on the Florida sea islands including Amelia Island but they appear to be less common along the northeast coast of Florida (Florida Fish and Wildlife Conservation Commission online survey data, Gerhart 2007) than elsewhere in the state. Horseshoe Crabs occur in the inlets (e.g., Ponce, Sebastian), lagoons (e.g., Mosquito) and rivers (e.g., Halifax, Banana, Indian) and associated islands along the Southeast coast of Florida from Ponce Inlet to Merritt Island, collectively referred to as the Indian River lagoon system (Ehlinger and Tankersley 2009). Farther north, Horseshoe Crabs are commonly found nesting along the shores of all the sea islands of Georgia, from Cumberland to Tybee Island (Sandifer et al. 1980; D. Saunders, University of Georgia Marine Laboratory at Skidaway Island and H.J. Brockmann, University of Florida, pers. comms), and they are common on all the coastal islands of South Carolina. Horseshoe Crabs are relatively common in the fishery independent monitoring (FIM) trawls in South Carolina throughout the year (South Carolina Department of Natural Resources).
Gulf of México
U.S. Coast. Horseshoe Crabs breed in all coastal counties along the west coast of Florida (Florida Fish and Wildlife Conservation Commission online survey data) including the Panhandle of Florida, the Florida Keys, and the Marquesas (Mikkelson 1988) but there are no records as far west as the Dry Tortugas (T. Ziegler, Fisheries Biologist, Everglades and Dry Tortugas National Parks, pers. comm). Florida FIM trawl surveys in the Gulf have recorded Horseshoe Crabs in every month of the year (Bob McMichael, Fish and Wildlife Research Institute/FIM, pers. comm). Farther west, Horseshoe Crabs are rarer along the coasts of Alabama and Mississippi compared to Florida; only three Horseshoe Crabs have been captured in Mississippi state trawl surveys since 1995 (Darcie Graham, University of Southern Mississippi Assistant Director, Center for Fisheries Research and Development, pers. comm). However, Horseshoe Crabs breed regularly on the northern side of the barrier islands of Alabama in Mississippi Sound, west of Mobile Bay at Dauphin Island (Hedgpeth 1954; Richmond 1962; R. Carmichael, Dauphin Island Sea Lab, pers. comm.), with breeding pairs rarely observed on the Fort Morgan Peninsula and Gulf Shores area of Alabama (R Carmichael, Dauphin Island Sea Lab, pers. comm.; Estes 2015). Horseshoe Crabs breed on Mississippi’s Petit Bois Island (R Carmichael, Dauphin Island Sea Lab, pers. comm.; Estes et al. 2015) Horn Island (S. J. VanderKooy, Gulf States Marine Fisheries Commission, pers. comm.) and West Ship Island (Fulford and Haehn 2012). Solitary Horseshoe Crabs are occasionally found on the Gulf of Mexico (southern) side of the barrier islands in Alabama and Mississippi (R. Carmichael, Dauphin Island Sea Lab, pers. comm). The western extent of historically recorded Horseshoe Crab breeding in the Gulf of México is the Chandeleur Islands, the eastern most barrier islands of Louisiana (Cary 1906). Louisiana has no records of Horseshoe Crabs in their trawl surveys (Martin Bourgeois, Louisiana Department of Wildlife and Fisheries pers. comm). There are no records of Horseshoe Crabs from the Texas trawl surveys (Glen Sutton, Texas Parks and Wildlife Department, pers. comm.), and only one historic record of a Horseshoe Crab collected at Padre Island in 1940-41 (Hedgpeth 1954). A small population was introduced by humans in Galveston Bay, Texas (Britton and Morton 1989), but apparently this population has not persisted because Horseshoe Crabs are not known to currently exist in Texas (P. Montagna, University of Texas, pers. comm).
Mexican coast. Along the Mexican coast of the Gulf of México and Caribbean Sea, Horseshoe Crabs occur on the Yucatán Peninsula, i.e., in the states of Campeche, Yucatán and Quintana Roo (Britton and Morton 1989), with only rare reports of Horseshoe Crabs from Veracruz (Chávez and Muñoz-Padilla 1975) and no reports from the coasts of Tabasco or Tamaulipas. Horseshoe Crabs are common on the west coast of the Yucatán Peninsula from Laguna de Términos and Isla del Carmen (southern Campeche) up to Celestún (Yucatán), as well as along the north coast of the peninsula (Yucatán and the north coast of Quintana Roo). There are reports of Horseshoe Crabs on the east (Caribbean) coast of the peninsula at least south to Tulum (Ives 1891, Gómez-Aguirre 1979) and Punta Allen (R. Sapién, Universidad Nacional Autónoma de México, pers. comm.), on the northern limit of Bahia de la Ascension in Quintana Roo (Zaldívar-Rae et al. 2009). Horseshoe Crabs in México are primarily associated with the mangrove communities in coastal lagoons and estuaries.
Although there are no published accounts of Horseshoe Crabs anywhere in the Caribbean, Mikkelson (1988) reported that “Old books on the fauna of the West Indies describe Horseshoe Crabs on the coast of Jamaica”. A few Horseshoe Crabs have been observed across a number of years in the Bahamas by Dr. Kathleen Sullivan-Seeley (Department of Biology, University of Miami, pers. comm.) who has been conducting an invertebrate survey in this area. In her logs (1986-2002) she notes the presence of Horseshoe Crabs at Chub Cay, Normans Cay on Shroud, Bogue Sound on South Caicos, Elizabeth Harbor on Exuma, and also on the islands of New Providence, and Eleuthera (Sullivan-Seeley, University of Miami, Bahamas log, unpub. data). Dr. Sullivan-Seeley has not observed Horseshoe Crabs in Jamaica or on the southeast coast of the Dominican Republic, where she has conducted surveys over the course of six years with dives and trawls (pers. comm). Scientists in Cuba report that they have not seen breeding Horseshoe Crabs in Cuba (E. Perera, Center for Marine Research in Havana, pers. comm.) although individual animals may be found there occasionally.
Transport and Introduction
The distribution of Horseshoe Crabs does not appear to have been influenced by transport and introduction into new areas. There are a number of old accounts of Horseshoe Crabs being sighted in Europe (e.g., Southwell 1873, Lloyd 1874) and from 1968 to 1976 at least eighteen were collected by fishermen or found northern European beaches (Wolff 1977). The presumption is that these animals were transported by humans across the Atlantic (Wolff 1977). Reports of Limulus in waters along Israel and western Africa were probably also due to transplanted animals (Mikkelsen 1988, Anderson and Shuster 2003). Two previous large-scale introductions of Limulus, one into San Francisco Bay on the Pacific coast of the U.S. (MacGinitie and MacGinitie 1949) and the other along the southern coast of the North Sea (Lloyd 1874), did not result in the permanent establishment of the species in either location (Wolff 1977).
Native:Mexico (Campeche, Quintana Roo, Yucatán); United States (Alabama, Connecticut, Delaware, Florida, Georgia, Louisiana, Maine, Maryland, Massachusetts, Mississippi, New Hampshire, New Jersey, New York, North Carolina, Rhode Island, South Carolina, Virginia)
|FAO Marine Fishing Areas:|
Atlantic – northwest; Atlantic – western central
|Range Map:||Click here to open the map viewer and explore range.|
|Population:||Population Genetic Structure|
Due to their morphological similarity to mid-Mesozoic taxa, Horseshoe Crabs are considered to be evolutionarily static and have been referred to as phylogenetic relics (Selander et al. 1970). However, close inspection has revealed the presence of considerable variability and geographic differentiation in morphology (Shuster 1979, Riska 1981) and genetic diversity (Selander et al. 1970, King et al. 2005). A range of molecular genetic techniques applied across multiple studies has been utilized in attempts to assess population structure (stock identification) in Horseshoe Crabs. These studies, which now include the first range-wide surveys of nuclear DNA variation in any Horseshoe Crab (Xiphosura: Limulidae), did not support the null hypothesis of a homogeneous gene pool for Horseshoe Crabs inhabiting the Atlantic coast of North America. The pattern of genetic variation observed was consistent with that identified previously in surveys of morphological variation (Shuster 1979, Riska 1981).
A survey of allozyme variation among four broadly distributed collections suggested that Atlantic and Gulf of México populations of Horseshoe Crabs were genetically differentiated (Selander et al. 1970). A subsequent study of mitochondrial DNA (mtDNA) variation identified a major genetic discontinuity distinguishing northern from southern populations with a phylogeographic break occurring around Cape Canaveral, along Florida’s Atlantic coast (Saunders et al. 1986). At a finer scale, Pierce et al. (2000) reported little evidence of gene flow between Delaware and Chesapeake Bay Horseshoe Crab populations as reflected by sequence variation in the mtDNA cytochrome oxidase subunit I region, though variation at randomly amplified polymorphic DNA (RAPD) markers was uniform, implying that gene flow may be sex-biased. Most recently King et al. (2005 and 2015) have surveyed neutral (assumed) genetic variation at 13 microsatellite DNA markers of 1,841 Horseshoe Crabs sampled at 35 spawning locations (Table 1 in the supplementary material) from northern Maine to the Yucatán Peninsula, México. This extensive intraspecific examination of the nuclear genome (nDNA) has revealed the presence of considerable allelic diversity and differentiation (population structuring) that appears to reflect the actions of various evolutionary processes.
These recent findings (King et al. 2005 and 2015) suggested the presence of similar levels of genetic diversity and variation among the collections punctuated with a series of genetic discontinuities of varying “depth” across the species’ range that could indicate demographic independence, regional adaptation, and reflect vicariant geographic events. Populations sampled within these regional groupings exhibited shallow but statistically significant differentiation. Moreover, patterns of population relatedness were consistent with the observation that populations at both ends of the species’ range are more differentiated from proximal populations than those in the middle (or zone of abundance).
Patterns of genotypic variation in the nDNA at the individual and population scales suggest three major zones of genetic discontinuity: 1) the Southeast (and northward) from the Florida Atlantic (Florida Indian River [FIR] and Biscayne Bay (FBB]) collections; 2) the Florida Atlantic (to the southern tip of Florida) from the Florida Gulf of México collections; and 3) the Florida Gulf of México from the Yucatán, México collections (Figures 1 and 3, Table 1 in the supplementary material). The latter discontinuity was identified using collections from a single locality on the northeast coast of the Yucatán Peninsula (San Felipe-Río Lagartos); hence, there may be other zones of genetic discontinuity within the Mexican part of the distribution. Narrower zones of genetic discontinuity were evident between a) the Gulf of Maine and Mid-Atlantic collections, b) the Mid-Atlantic and Southeast collections (SC and GA), and c) the Tampa Bay and Cedar Key collections. An additional zone of discontinuity may exist between Alligator Point and St. Joseph Bay along the Florida panhandle. The relatively small sample size from St. Joseph Bay precludes a determination at this time. This phylogeographic pattern implies there are at least seven demographically distinct lineages across the species’ range that are relevant in conservation considerations. In addition to support for recognition of these zones of discontinuity, these data suggest low levels of gene exchange between collections on either side of these genetic discontinuities. Additional data across the Gulf of Mexico may further divide or unify the population structure of this region.
In addition to the demographically discrete lineages identified for Horseshoe Crabs, a series of metapopulations and other individual collections delineated within each discrete lineage may be considered distinct management/recovery units for future management planning purposes. Metapopulations may exist in the Gulf of Maine (Maine and New Hampshire collections), the entire Mid-Atlantic region (with some substructure within), the upper Chesapeake Bay collections (MDT, MDF), the Southeast assemblage (SBB, SBE, GSA, and GSI), southwest Florida Gulf of México (FMI, FCH, FTB), and the northwest Florida Gulf of México (FCK, FAP). Within areas bounded by zones of genetic discontinuity, there appears to be substantial gene flow between each population and its nearest neighbours; the presence of these metapopulations appears to bode well for the demographic fitness of some regions.
Genetic diversity was sufficiently high at the nDNA markers that each individual surveyed (N = 1,841) possessed a unique multilocus genotype. This allowed assessment of sex-specific gene flow patterns which indicated a trend toward reduced philopatry (breeding site fidelity) in male Horseshoe Crabs throughout their range. This trend toward decreased female vagility and increased male vagility peaked in the region between the Chesapeake and Delaware Bays. There is significantly more male migration between these two bays than female gene flow. These results agreed with those of Pierce et al. (2000) in regard to both structure of mtDNA variation and lack of structure detected by the nDNA RAPD markers. This sex-biased dispersal persists even with direct access between the bays via the Chesapeake and Delaware Canal constructed in the early half of the 19th century. These findings imply that should a population become extirpated, gene flow alone may not be sufficient to repopulate an area due to limited larval dispersal potential (Botton and Loveland 2003) and female migration (Swan 2005) between embayments.
Any further quantification of the degree of migration between Delaware and Chesapeake Bays is difficult due the absence of structure between sample collections from the two bays. Additional mitochondrial DNA data validated by tagging studies targeting females are required to allow quantification of the effective migration between Chesapeake and Delaware Bays. Swan (2005) found that among the 14 Horseshoe Crabs that were observed to have travelled >100 km from their tagging location, one was recovered in Chesapeake Bay, but not during spawning season.
In their entirety, these research findings suggest a series of discontinuities across the species’ range that could indicate regional adaptive significance or reflect vicariant geographic events. Regional groupings (Figure 3 in the supplementary material) may warrant management unit status based on the presence of statistically significant allele frequency heterogeneity, allocation of genetic diversity, and a high percentage of correct classification to region of origin. Moreover, a pattern of male-biased gene flow was observed among all collections from Maine to Florida’s eastern Gulf of Mexico coast that suggests management efforts might best be targeted at a finer scale. These findings also provide justification for release of Horseshoe Crabs sampled by scientific or commercial interests near the original collection site.
The presence of demographically distinct and evolutionarily significant lineages delineated by zones of genetic discontinuity is consistent with the findings of researchers assessing behavioural patterns. The integration of the information from the nuclear genome with previously identified allozyme (Selander et al. 1970) and mitochondrial DNA variation (Saunders et al. 1986, Pierce et al. 2000) and ecological data should prove essential to developing an ecologically and evolutionarily sound management strategy.
Based on the major zones of discontinuity in the genotypic patterns of nDNA, we structured the risk assessment into the following regions and then integrated the regional assessments to the species level. The transnational genetically-defined regions (also referred to a subpopulations) were (Figures 1 and 3 in the supplementary material):
Data availability determined whether quantitative analysis was valid for trend estimation. In the absence of a quantitative analysis, descriptive summaries of observations were included to infer qualitative trends in relative abundance and distribution.
Qualitative Trends in the Gulf of Maine
The northernmost population of Horseshoe Crabs was studied for 10 years from 2001 to 2010 (Schaller 2011, Schaller and Dorsey 2011). The study included daily surveys in Taunton Bay, Maine during late May and June each year where 6,964 spawning Horseshoe Crabs were tagged and released (sex ratio of 1.8 males to females). Based on this study, the authors were “cautiously optimistic” that the population in Taunton Bay was stable (Schaller 2011). Pete Thayer (Maine Department of Marine Resources, pers. comm.) stated that “Over the late 90’s to late 2000’s, Horseshoe Crabs were fished down a bit for eel bait until a seasonal closure regulation was enacted, from which point they appeared to be bouncing back in the survey’s final years.” Moore and Perrin (2007) who tracked Horseshoe Crabs in Taunton Bay, a tributary of Frenchman Bay, during 2003-2005 observed no emigration; and thus considered the populations to be resident to the embayment.
In 2001 to 2004, spawning surveys were conducted at five sites in the Gulf of Maine to establish baseline data (Schaller et al. 2005). Schaller et al. (2005) remarked that Horseshoe Crab spawning density is sparse throughout Maine and that three historical breeding sites are no longer used by Horseshoe Crabs for spawning. Of the five sites where surveys were conducted in all years (2001 to 2004), counts of spawning Horseshoe Crabs increased in three surveys and decreased in two.
Quantitative Trends in Gulf of Maine (New Hampshire), Mid-Atlantic, Southeast, Florida-Atlantic, and Northeast Gulf of México
Data were available from 40 fishery-independent data sets covering Mid-Atlantic and Florida regions (New Hampshire to Florida; regions as defined above and in Figure 1) over a range of years. The fishery-independent data sets were selected by the Atlantic States Marine Fisheries Commission (ASMFC) for stock assessment (ASMFC, Sweka et al. 2013). ASMFC selects datasets that are overseen or conducted by state or federal agencies or academic institutions using standardized methodology and survey design. State agencies rely on these datasets to comply with ASMFC monitoring requirements. The basic data were individual counts of Horseshoe Crabs within sampling units; the demographic (age-class, sex) and temporal and spatial resolution of each dataset is described in Sweka et al. (2013: Appendix B) and summarized in Table 2 (in the supplementary material).
We analysed trends from each dataset and then used meta-analysis techniques to summarize inference at the regional or sub-regional level because the data came from many independent monitoring programs. We grouped the datasets from the Mid-Atlantic region into sub-regions because of geographic differences in harvest pressure and environmental conditions. The sub-regions were New England states (NH, RI, MA), New York area (CT, NY), and Delaware Bay area (NJ, DE, MD, VA). In addition, datasets represented the Southeastern (NC, SC, GA), Florida Atlantic (FL), and Gulf of México (FL) regions. There were no state-specific datasets from NC; however, data from an offshore monitoring program (SEAMAP) included waters off the NC coast. The time series varied among the datasets. The New England area included the longest time series, with one data set from 1959 and several that started in the 1970s. Data sets from the New York and Delaware Bay areas started in the late 1980s. Data sets from the Southeast included several that started in the mid-1990s.
The objective of the meta-analysis of regional trends was to determine change in Horseshoe Crab populations during the periods defined by the available data. The trend analyses involved fitting a linear regression to the data, which had been standardized by subtracting the mean and dividing by the standard deviation. Standardization was required for the trend analysis results based on individual datasets to be combined using meta-analysis techniques.
We used three meta-analysis techniques described by Manly (2001:123-125):
For those regions or sub-regions with negative weighted slope (i.e., Gulf of Maine (NH), New England area, New York area, Northeast Gulf region), population reduction over 40 years, which approximates three generations based on age-structured population models (Sweka et al. 2007), can be projected assuming the current linear trends continue and the index represents population abundance (see the Supplementary Material for the formula used for this projection.
Continuation of these negative trends would result in projected population reductions of 100% in Gulf of Maine (NH), 92% in New England, 11% in New York, 55% in Florida Atlantic, and 32% in Northeast Gulf of México. Although not accounting for carrying capacity limits to population growth, projections indicate population increases in the Delaware Bay of 116% and in the Southeast region of 218% over 40 years.
Qualitative Trends in the Yucatán Peninsula
It has been claimed that population sizes in México dwindled dramatically between the 1960s and the early 1990s, especially in the Laguna de Términos area (Gómez-Aguirre 1979, 1980, 1983, 1985, 1993). However, these claims cannot be fully substantiated as they were not based on formal quantitative surveys. There have been no periodical quantitative assessments of population sizes in México, which would allow comparisons between at least two points in time for specific localities.
Population Viability Analyses
There have been several efforts to develop population models useful for assessing the viability of populations subject to harvest and to explore population dynamics (Gibson and Olszewski 2001, Grady and Valiela 2006, Davis et al. 2006, Sweka et al. 2007, McGowan et al. 2011b, Smith et al. 2013). Modelling studies have focused on Horseshoe Crab populations in Rhode Island (Gibson and Olszewski 2001), Cape Cod (Grady and Valiela 2006), and Delaware Bay (Davis et al. 2006, Sweka et al. 2007, McGowan et al. 2011b, Smith et al. 2013). All analyses concluded that over harvest results in population depletion, but that some levels of reduced harvest can be compatible with a viable population. Gibson and Olszweski (2001) and Davis et al. (2006) used production models to examine biomass recovery of populations in Rhode Island and Delaware Bay that had been depleted by overfishing. Gibson and Olszweski (2001) estimated intrinsic growth rate of 0.5 (finite rate of 1.6) for population in Rhode Island and concluded that recovery would take 10 years under no harvest and 20 years under harvest well below recent levels. Davis et al. (2006) concluded that the Delaware Bay population had been overfished and projected that recovery could occur within four years, but likely would take longer than 15 years. Grady and Valiela (2006) and Sweka et al. (2007) used life-history structured models to examine population dynamics of populations in Cape Cod embayments and Delaware Bay, respectively. Sensitivity analyses indicated that population growth was most sensitive to variation in early life stage and juvenile survival. The generation time according to the age-structured population model (Sweka et al. 2007) model is 13.7 years. The modelling by Grady and Valiela (2006) suggested seasonal closures along with low levels of harvest are required for sustainability. Sweka et al. (2007) examined the role of density-dependent egg mortality on population abundance under different harvest levels, and consistent with previous analyses, identified sustainable harvest levels that allowed for population growth. The Sweka model became the basis for predictive modelling to support adaptive management of Horseshoe Crab and Red Knots (Calidris canutus rufa) in Delaware Bay (see the Conservation Actions section, Harvest Management; McGowan et al. 2011b, Smith et al. 2013).
|Current Population Trend:||Decreasing|
|Habitat and Ecology:||Reproduction|
Timing of spawning varies with latitude. Increased water temperatures in the spring stimulate adult Horseshoe Crabs to migrate from deeper waters where they overwinter toward shallow waters where they spawn (Shuster 1982, Watson et al. 2009), although in the Yucatán spawning activity is associated with decreased water temperatures (J. Gutiérrez and J. Zaldívar-Rae, Anáhuac Mayab University, unpub. data). On the Yucatán Peninsula, L. polyphemus seems to spawn throughout the year (Alvarez-Añorve et al. 1989, Barba-Macías et al. 1988 (field report), Bonilla-González et al. 1986, Rosales-Raya 1999), although a markedly seasonal winter-spring spawning pattern, with a peak in December, was recently detected in Chuburná, on the northern coast of the peninsula (J. Gutiérrez and J. Zaldívar-Rae, Anáhuac Mayab University, unpub. data). Similarly, in South Florida (FWC online survey) and Indian River Lagoon area of Florida (Ehlinger and Tankersley 2007) Horseshoe Crabs have been observed spawning during every month of the year with peaks in April, May and August. In Mississippi spawning is observed from early April to mid-November with a peak in April and May (Fulford and Haehn 2012). Along the coast of the Florida Panhandle, breeding occurs from February through October with peaks in March and April (Rudloe 1980, FWC online survey). In Georgia and South Carolina spawning has been reported from March to July, peaking in May (Thompson 1998). In Delaware Bay, spawning occurs from April through at least July, with peak spawning in May and June (Shuster and Botton 1985, Michels et al. 2008, Smith and Michels 2006). In Long Island Sound, spawning generally begins in early May and peaks by the end of May (Beekey and Mattei 2009). In Cape Cod, Massachusetts, spawning occurs from May through July with apparent peaks in May and June (Barlow et al. 1986, Widener and Barlow 1999, James-Pirri et al. 2005).
Spawning is strongly associated with the new and full moon high tides during the spawning months, which are the highest tides in the month (day or night) (Rudloe 1980, Brockmann 2003b, Barlow et al. 1986, Smith et al. 2002b). However, Smith et al. (2010) reported that the association between spawning activity and lunar period was only slightly higher than what would be expected from chance alone. Within these new and full moon tidal ranges, Horseshoe Crabs seem to prefer the higher of the two unequal tides in a day (Barlow et al. 1986, Cohen and Brockmann 1983, Rudloe 1980). In Delaware, spawning activity predominantly occurs during the evening high tides (Shuster and Botton 1985, Smith et al. 2010). In Pleasant Bay, Cape Cod, spawning occurred at a similar level during all daytime high tides regardless of lunar phase (James-Pirri et al. 2005, Leschen et al. 2006). In microtidal areas, such as Indian River Lagoon, Florida, breeding activity is episodic (Ehlinger et al. 2003) or breeding activity is affected by increased water level from wind-blown surge, such that deeper water results in a larger number of crabs (Rudloe 1985). Even where there is a 1m tidal inundation, water level from wind-blown surge strongly influences the numbers of spawning horseshoe crabs (Brockmann and Johnson 2011).
Spawning adults prefer sandy beach areas within bays and coves that are protected from wave energy (Shuster and Botton 1985, Smith et al. 2002a, Jackson et al. 2002, Landi et al. 2015). Nests are primarily located between the low tide terrace (tidal flat) and the extreme high tide water line (Penn and Brockmann 1994, Weber and Carter 2009). Weber and Carter (2009) found that 85% of nests were deposited between the tidal flat and the nocturnal high tide wrack line on the western shore beaches of Delaware Bay. Penn and Brockmann (1994) noted that nests occurred in a narrower band along the high tide line of beaches on the west coast of Florida. On the Yucatán Peninsula, Horseshoe Crabs spawn on small beaches limited by mangroves or near the edges of small mangrove islands within coastal lagoons where organic matter abounds and microbial decomposition is high (Zaldívar-Rae et al. 2009). Spawning is sometimes observed on offshore sandbars and oyster bars (Wenner and Thompson 2000). On the Mississippi coastal islands, breeding is limited to the north side in intertidal sand beach habitat (Fulford and Haehn 2012). Some sub-tidal nesting also occurs in sands with high oxygen, such as the sand flats just off the beach. Most nesting beaches have nearby nursery habitats for juveniles (Botton and Loveland 2003). Geographic differences in nest site selection can be explained by differences in wave energy, beach morphology, and geochemistry (Botton et al. 1988, Penn and Brockmann 1994, Smith et al. 2002a, Landi et al. 2015). Sediment grain size in particular can influence spawning site selection as environmental conditions in the sand affect development (moisture, temperature, and oxygen gradients) (Penn and Brockmann 1994, Jackson et al. 2008). Previous studies suggest that females avoid laying eggs in eroded beaches that are high in hydrogen sulphide and where sediment pore water is low in oxygen - presumably because these conditions are detrimental to egg development (Botton et al. 1988, Penn and Brockmann 1994). In Massachusetts, New Jersey, and Delaware, spawning beaches are typically coarse-grained and well drained, as opposed to Florida beaches, which are typically fine-grained and poorly drained (Penn and Brockmann 1994). In Long Island Sound, nests can be found on beaches ranging from coarse grained and well drained to cobble dominated substrates to fine grained and poorly drained muddy substrates (Beekey and Mattei 2009). In Yucatán, spawning pairs seem to prefer the high tide line of beaches where coarser sand and rubble are mixed with the more common fine sand/clay substrates, usually at the base of man-made structures and roadsides that reach the water (J. Gutiérrez and J. Zaldívar-Rae, Anáhuac Mayab University, pers. obs).
Female crabs typically arrive at the spawning beach with a male attached to their posterior opisthosomal spines. Males use their modified, claw-like pedipalps to grasp onto females’ terminal spines (Loveland and Botton 1992, Brockmann 2003a) as they head from the water to the spawning beaches (Shuster 1982). Active mate choice by females has been observed only in an experimental setting (M Hart and HJ Brockmann, University of Florida, unpub. data). Males locate female mates offshore using visual (Barlow et al. 1982, Barlow and Powers 2003) and chemoreceptive cues (Saunders and Brockmann 2010). In addition, unattached males come onto shore and gather around nesting pairs as satellites (Brockmann and Penn 1992). Satellite males are attracted to nesting pairs by visual and chemical cues and appear to be attracted to larger females (Hassler and Brockmann 2001, Schwab and Brockmann 2007). Attached males do not differ in size from unattached males but they are in better condition, more active, have a higher sperm concentration, remain attached longer and probably are younger (more recently moulted into the adult) than satellite males (Brockmann and Penn 1992, Loveland and Botton 1992, Brockmann 2002, Duffy et al. 2006, Sasson et al. 2012) and they feed little while attached whereas females and satellites males do feed (Smith et al. 2013). Some pairs do not have satellite males (monandry) and these females tend to be smaller and lay fewer eggs than those nesting in association with many males (polyandry) (Johnson and Brockmann 2012). Monandry seems to be the rule on at least two sites on the northern coast of the Yucatán Peninsula (Chuburná and Chabihau), with very few pairs having a single satellite male (J. Gutiérrez and J. Zaldívar-Rae, Anáhuac Mayab University, unpub. data). Although a single attached male can fertilize all of the eggs, when satellite males are present (two to four) they may fertilize a majority of the female’s eggs (Brockmann et al. 1994, 2000). However, with increasing numbers of satellites their reproductive success declines (Brockmann et al. 2000, Johnson and Brockmann 2010). Differences in mating strategies have been attributed to variation in spawning density, operational sex ratios, and male condition (Duffy et al. 2006, Brockmann and Smith 2009, Mattei et al. 2010). Single females have been observed excavating nests on spawning beaches in Long Island Sound where spawning indices are extremely low (0.002 females m-2) however it is unknown whether eggs were deposited or not (Mattei et al. 2010). Elsewhere, when females arrive on the beach without males, they do not lay eggs (Brockmann 1990).
Operational sex ratios (OSR) on beaches are typically male biased (Loveland and Botton 1992). Unattached males typically return to the beach more frequently than females creating male-biased sex ratios and male-male competition for mates (Brockmann 1990; Brockmann and Johnson 2011; Smith et al. 2002a, 2010; Brockmann and Smith 2009; Rudloe 1980). The mean OSR in unharvested populations is generally 1.5-2.4 males to females (Brockmann and Johnson 2011, Schaller 2002, Mattei et al. 2010, Wenner and Thompson 2000, Rudloe 1980) but in harvested populations the sex ratio is normally 3-8 males/female (Kreamer and Michels 2009, Carmichael et al. 2003, James-Pirri et al. 2005, Smith et al. 2002a, Smith et al. 2009b). Mean OSR during the 2012-2013 spawning season in Chuburná, Yucatán, was 1.1 males to females, consistent with a generally monandric mating system (J. Gutiérrez and J. Zaldívar-Rae, Anáhuac Mayab University, unpub. data).
Horseshoe Crabs are the only extant marine arthropod with external fertilization of their eggs that does not brood the eggs (Brusca and Brusca 1990). On spawning beaches, females excavate a pit below their body and deposit two to five clusters of eggs at depths from 5 to 20 cm (Rudloe 1979, Brockmann 1990, Leschen et al. 2006, Brockmann 2003b). Males externally fertilize the eggs as they are being deposited. Horseshoe Crab fecundity varies with latitude and with female size (Botton et al. 2010). Shuster (1982) reported 88,000 eggs per female for the Delaware Bay. Average fecundity was correlated with female size in Pleasant, Bay Massachusetts with fecundity ranging from 14,500 eggs for a 201 mm prosomal width (PW) female to 63,500 eggs for females >261 mm PW (Leschen et al. 2006).
Cluster size also varies latitudinally. In Florida, cluster size was reported to be 2,236 (Brockmann 1990) rising to 2,365-5,836 eggs/cluster in Delaware Bay (Shuster and Botton 1985, Weber and Carter 2009). In Long Island Sound, cluster size averages 3,741 eggs (Beekey et al. 2013) compared to 640-1,280 in Cape Cod (Leschen et al. 2006). Cluster size is not correlated with female size (Brockmann 1996, Leschen et al. 2006), but larger females lay more clusters per spawning season than smaller females. Females typically lay multiple nests during one tidal cycle and in many cases return over multiple tidal cycles (Brockmann 1990, Brockmann and Penn 1992, Brousseau et al. 2004, Smith et al. 2010). Brockmann (1990) reported that in Florida females returned to nest on average 3.4 times and most spawned during only one tidal cycle (five days of high tide around new or full moon), whereas males returned over two or more tidal cycles (Brockmann and Penn 1992). In Delaware Bay females spawned over two to five consecutive nights, remaining within 50 to 715 m of their established spawning beach before moving away from the beaches several days after the tidal cycle (Brousseau et al. 2004, Smith et al. 2010).
Horseshoe Crab development has been the subject of several classic monographs (Packard 1872, 1885; Kingsley 1892, 1893; Munson 1898) and a number of more recent studies (Brown and Clapper 1981, Sekiguchi et al. 1982, Shuster and Sekiguchi 2003). A review of the developmental biology and ecology is provided by Botton et al. (2010). Unfertilized eggs are typically greenish blue or grey and between 1.6-1.8 mm diameter. Unfertilized eggs have a large volume of yolk surrounded by a tough outer chorion. Egg development is dependent on temperature, salinity, moisture, and oxygen. Trilobite larvae hatch from the eggs within 2-4 weeks, although some larvae may overwinter within nests and hatch out the following spring (Botton et al. 1992). Hatching is triggered by environmental cues associated with high water conditions (hydration, physical disturbance, hypoosmotic shock that helps to maximize survival by preventing larvae from being stranded on the beach (Botton et al. 2010). Trilobite larvae are weak swimmers and generally rely on vertical movement to take advantage of selective tidal stream transport. Larvae settle within a week of hatching and begin molting (Shuster 1982). Larvae and juvenile crabs remain in the intertidal flats, usually near breeding beaches (Botton and Loveland 2003). These findings suggest limited larval dispersal (Botton and Loveland 2003). Approximately two weeks after hatching, larvae molt to the juvenile stage (second instar stage) where the telson is formed. Many juveniles reach the fourth instar by the end of their first summer (Botton et al. 1992). Over time, the older juveniles move out of intertidal areas to deeper waters (Botton and Ropes 1987) where they remain until they reach maturity.
Horseshoe Crabs undergo stepwise growth shedding of their exoskeleton at least 16 or 17 times before reaching sexual maturity (Shuster 1950), a process that takes 9-10 years (Shuster and Sekiguchi 2003). Females are typically larger at maturity than males. Smith et al. (2010), reviewing several studies, reported the average prosomal width growth increment (ratio of PW from instar i to i+1) for all instars was 1.28 (range: 1.15–1.52). Growth is relatively rapid during the first several years progressing through stages I–V in the first year, stages VI–VII the second year, stages VII–IX the third year, with a single molt per year until reaching maturity (Shuster 1982). Shuster (1950) approximated that it took 9 to 12 years for Horseshoe Crabs to reach sexual maturity. Sekiguchi et al. (1982) concluded that Limulus polyphemus molts 16 times and matures in their ninth year; females molt 17 times and mature in their tenth year. Smith et al. (2010) found that males in Delaware Bay tended to mature at age 10 and 11, while females tended to mature at ages 10, 11 and 12. Marked adults have been observed during 6-10 years, which means that some individuals may reach at least 20 years of age (Ropes 1961, Shuster 1958, Botton and Ropes 1988, Grady et al. 2001, Swan 2005, Brockmann and Johnson 2011). Limulus polyphemus attains their average largest size at the central portion of their range (Delaware Bay) but are significantly smaller north of Long Island Sound and in the southernmost part of their range in the Gulf of México (Shuster 1979, Graham et al. 2009) and México.
Migration and Dispersal
The general movement patterns of Horseshoe Crabs include (1) juveniles move from spawning beaches to deeper waters as they age, (2) juveniles reach sexual maturity in their natal estuary or migrate offshore to mature in the ocean, and (3) adults migrate annually from the ocean or deep bay waters to spawn on estuarine beaches (Baptist et al. 1957, Shuster 1979, Shuster and Botton 1985, Botton and Ropes 1987, Botton and Loveland 2003, Smith et al. 2009a). There is considerable evidence, however, that migratory patterns may be more complex. Smith et al. (2009a) suggested Horseshoe Crabs in Delaware Bay exhibit sex-specific migratory patterns. Until about age eight years, juveniles of both sexes remain within the bay. After age eight years, females begin to migrate to the continental shelf as older juveniles and mature in the ocean. In contrast, males tend to remain within the bay to mature. After reaching maturity, both sexes migrate from the ocean or deep bay waters to spawn on the estuarine beaches. While the greatest proportion of the Delaware Bay Horseshoe Crabs appear to migrate to the continental shelf (Botton and Ropes 1987, Hata 2008), tagging data indicate that some Delaware Baycrabs and most crabs across the New England States remain within local regions and overwinter in local embayments (Botton and Ropes 1987, James-Pirri et al. 2005, Swan 2005, Smith et al. 2006, Moore and Perrin 2007, Beekey and Mattei 2009). Landi et al. (2015) found that spawning beach locations within Long Island Sound tended to be those closer to offshore locations where adults were caught in trawl surveys. These data are further supported by stable isotope analyses, which indicate adult crabs are loyal to local feeding grounds (Carmichael et al. 2004, O’Connell et al. 2003). Finally, acoustic telemetry data and tracking studies have shown that many animals remain year-round within one bay or estuary (Rudloe 1980, Ehlinger et al. 2003, Beekey and Mattei 2009, Watson and Chabot 2010, Brockmann and Johnson 2011).
Factors contributing to natural mortality include age and excessive energy expenditure during spawning, which can result in stranding, desiccation, and predation. Loveland et al. (1996) reported that the natural mortality rate in adults is low with the single greatest source due to beach stranding. Botton and Loveland (1989) concluded that stranding mortality, which they estimated to be about 10% of the total population in Delaware Bay in the mid-1980s, is likely to vary among estuaries because it is affected by population density, weather and tidal conditions, and beach geomorphology. The condition of the individual, which is probably age related, is also a factor in stranding-related mortality (Penn and Brockmann 1995, Smith et al. 2010). Carmichael et al. (2003) found in Pleasant Bay, Massachusetts, adults had a lower estimated mortality rate than juveniles, and there was no significant difference in estimated mortality rate between adult males and females. In contrast, Butler (2012) found through analysis of mark-recapture data from Delaware Bay that adult male annual survival (77%, SE = 0.04) was greater than adult female survival (65%, SE = 0.09). Adult and juvenile Horseshoe Crabs make up a portion of the Loggerhead Sea Turtle's (Caretta caretta) diet in the Chesapeake Bay (Keinath 2003, Seney and Musick 2007), but the severity of Horseshoe Crab mortality due to predation from sea turtles, alligators in the southeast (Reid and Bonde 1990), and other marine animals remains unknown.
Shorebirds feed on Horseshoe Crab eggs in areas of high spawning densities such as the Delaware Bay (Botton et al. 1994, Botton et al. 2003). Horseshoe Crab eggs are considered essential food for several shorebird species in the Delaware Bay, which is the second largest migratory staging area for shorebirds in North America (Clark and Niles 1994). Despite significant shorebird predation on Horseshoe Crab eggs, such activity probably has little impact on the Horseshoe Crab population (Botton et al. 1994). Horseshoe Crabs place egg clusters at 5-25 cm deep (Weber and Carter 2009), which is deeper than most short-billed shorebirds can penetrate. Many eggs are brought to the surface by wave action and burrowing activity by spawning Horseshoe Crabs (Nordstrom et al. 2006). These surface eggs that are consumed by birds would not survive, due to desiccation (Botton et al. 1994). Horseshoe Crab eggs and larvae are also a seasonally preferred food item of a variety of invertebrates and finfish (Shuster 1982). In Florida, many shorebirds winter, particularly along the west coast, and many birds that migrate to South America stop in Florida on their return north. The highest migrating shorebird abundances are in Tampa Bay, Florida Bay, and Apalachicola Bay, all in the Gulf of México (Sprandel et al. 1997). Because populations of Horseshoe Crabs are relatively small in Florida, their eggs provide a less dependable food source than those in Delaware Bay so the presence of Horseshoe Crab eggs in the diet of Florida shorebirds is probably opportunistic (Gerhart 2007). Unreliability of Horseshoe Crab eggs as a food source also seems to be the case in the Yucatán Peninsula, where Horseshoe Crab abundances are relatively low in comparison to those of US populations (J. Gutiérrez and J. Zaldívar-Rae, Anáhuac Mayab University, unpub. data).
Limulus polyphemus has been described as an ecological generalist (Shuster and Sekiguchi 2009) able to tolerate a wide range of environmental parameters throughout its distribution although Sekiguchi and Shuster (2009) suggest that individual subpopulations may have narrower tolerances than the species as a whole. Habitat requirements change throughout the Horseshoe Crab life cycle, extending from intertidal beach fronts and tidal flats for eggs and larvae, to the edge of the continental shelf for adults.
Larval Habitat Requirements
Spawning habitat must include a sufficient depth of porous, well-oxygenated sediments to provide a suitable environment for egg survival and development (Botton et al. 1988). Nest depth on the western shore of Delaware Bay generally ranged between 3.5 and 25.5 cm (mean 15.5, SD 3.5), although nest depth may be affected by wave energy, bioturbation, or other factors after deposition (Weber and Carter 2009). These results are similar to those found by previous investigators on Delaware Bay beaches (e.g. Hummon et al 1976, Penn and Brockmann 1994, Botton et al. 1994). In the Laguna de Términos and Champotón areas of Campeche, México, nest depths range from 2 to 30 cm, and substrate composition in nesting sites varies widely, from an estuarine locality (Icahao, near Champotón) where up to 60% of the substrate was medium-grain to cobble, to a coastal lagoon site (Isla Pájaros, in Laguna de Términos) where 70% of the substrate was loam-clay to fine sand (Rosales-Raya et al. 1997).
Rate of egg development is dependent on interstitial environmental parameters including temperature, moisture, oxygen, and salinity (French 1979, Jegla and Costlow 1982, Laughlin 1983, Penn and Brockmann 1994) and disturbance (bioturbation) from external forces (Jackson et al. 2008). Optimal development occurs at salinities between 20 and 30 ppt (Jegla and Costlow 1982, Laughlin 1983), although populations from microtidal lagoon systems that often experiences high salinities (>50 ppt) had an optimal range of 30 to 40 ppt, with hatching occurring at salinities as high as 60 ppt (Ehlinger and Tankersley 2004, 2009). Egg development occurs most readily at temperatures ranging from 25 to 30°C (Jegla and Costlow 1982, Laughlin 1983, Penn and Brockmann 1994, Ehlinger and Tankersley 2004). Penn and Brockmann (1994) found optimal development of Horseshoe Crab eggs from Delaware and Florida to occur at oxygen concentrations between 3 and 4 ppm and moisture content between 5 and 10%. In Campeche, México, the salinity of interstitial water surrounding nests ranged from 25 to 59 ppt (Rosales-Raya et al. 1997).
Juvenile Habitat Requirements
Nearshore, shallow water, intertidal flats are considered essential habitats for development of juvenile Horseshoe Crabs (Botton 1995), since juveniles usually spend their first two years on the sand and mud flats just off the breeding beaches (Rudloe 1981). The diet of juveniles is varied including particulate organic matter from algal and animal sources (Gaines et al. 2002, Carmichael et al. 2004). Older juveniles are exclusively subtidal (Shuster and Sekiguchi 2009), with each succeeding stage moving toward deeper water. In the Delaware Bay, females begin to leave the Bay and move to continental shelf waters around age 7 to 8 to mature in the ocean (Hata and Hallerman 2009, Smith et al. 2009a). Smith et al. (2009a) provide evidence that males remain in the Bay until maturity (age nine years), but Hata and Hallerman (2009) found evidence of significant numbers of immature males on the shelf one to two years prior to reaching maturity. Delaware Division of Fish and Wildlife's 16-foot (4.9 m) bottom trawl survey data indicated that more than 99 percent of juvenile horseshoe crabs (<16 cm prosomal width) were captured at salinities >5 ppt (Michels 1996). In the southeast, juveniles have been reported to be active throughout the year, foraging in the intertidal zone within a few meters of the nesting beach (Rudloe 1981). They alternately crawl at the surface of the substrate and bury in the sand or mud, feeding on benthic organisms. As Horseshoe Crabs mature, the diet composition shifts to larger prey, and Horseshoe Crabs are known to be important predators of benthic meiofauna (Carmichael et al. 2004, Carmichael et al. 2009, Botton 2009).
Adult Habitat Requirements
Adult Horseshoe Crabs have been found as far as 35 miles (56 km) offshore at depths greater than 200 m; however, Botton and Ropes (1987) found that 74% of the Horseshoe Crabs caught in bottom trawl surveys conducted by the National Marine Fisheries Service (NMFS), Northeast Fisheries Science Center were taken in water shallower than 20 m. They are observed in a wide range of salinity regimes, from low salinity (<10 ppt) areas such as the upper Chesapeake Bay, to the hypersaline (>50 ppt) environments of the Indian River Lagoon in Florida. During the spawning season, adults typically inhabit bay areas adjacent to spawning beaches. In Delaware Bay, Horseshoe Crabs are active in the Bay area at temperatures above 15°C (Shuster and Sekiguchi 2009, Smith et al. 2010), while crabs in Great Bay, NH increase activity at temperatures above 10.5°C (Watson et al. 2009). In the fall, adults may remain in local embayments or migrate into the Atlantic Ocean to overwinter on the continental shelf. The northern extent of all Horseshoe Crab species may be limited by duration and severity of winter temperatures. The lack of Horseshoe Crab populations in the western Gulf of México, which has suitable beach spawning habitat, may result from the local hydrodynamic and tidal regime along with an absence of barrier islands to attenuate wave energy (R Carmichael, Dauphin Island Sea Lab, pers. comm). Nearly all Horseshoe Crab populations occur in areas with semi-diurnal tides of moderate amplitude, but tides of this type are not observed in the western Gulf of México.
Botton and Haskin (1984) and Botton and Ropes (1989) found the primary prey for adult Horseshoe Crabs are Blue Mussels (Mytilus edulis) and Surf Clams (Spisula solidissima). Recent declines in Surf Clam in the mid-Atlantic are being attributed to climate-change induced increases in water temperatures during late-summer and fall (E. Powell, Rutgers University, pers. comm). Adult Horseshoe Crabs are known to be important predators of a variety of benthic macrofauna (Carmichael et al. 2004, Carmichael et al. 2009, Botton 2009). The effects of a declining prey base on Horseshoe Crab population carrying capacity is unknown.
Horseshoe Crabs are an important part of the ecology of the coastal systems in which they are found (Botton 2009). They provide food for endangered sea turtles (Keinath 2003) and migrating shorebirds (Haramis et al. 2007). Their burrowing activities affect the habitat available for other species through bioturbation (Gilbert and Clark 1981, Kraeuter and Fegley 1994), and adult predatory activities affect the intertidal and subtidal meio- and macrofauna (Wenner and Thompson 2000, Ehlinger and Tankersley 2009).
|Continuing decline in area, extent and/or quality of habitat:||Yes|
|Generation Length (years):||13-14|
|Movement patterns:||Full Migrant|
|Use and Trade:||American Horseshoe Crabs are commercially harvested. Currently, most harvest is for use as bait in other fisheries (eel and whelk in the U.S.). Harvest by the biomedical industry for production of Limulus amebocyte lysate (LAL) is significant and increasing, but currently less than for bait and does not result in certain mortality as does bait harvest. Harvest for the marine life aquaria trade or scientific and educational collection is small in comparison to other uses, but is significant in Florida where juveniles are collected in large numbers (Gerhart 2007). Substantial evidence suggests that over-harvest can result in depleted populations and localized extirpations (Widener and Barlow 1999, Carmichael et al. 2003, Rutecki et al. 2004, Schaller et al. 2005, Gerhart 2007, Smith et al. 2009b, McGowan et al. 2011b).|
Historically, Horseshoe Crabs along the Delaware Bay were harvested heavily (1 to 5 million per year) for fertilizer dating back to the mid-1800s (Shuster and Botton 1985). Harvest of Horseshoe Crabs for fertilizer declined to a negligible level by the 1960s (Shuster 2003, Kreamer and Michels 2009).
Presently the largest coast-wide harvest is for bait. Horseshoe Crabs are commercially harvested primarily for use as bait in the conch (Busycon spp.) pot and American Eel (Anguilla rostrata) pot fisheries (ASMFC 2009). The increase in harvest of Horseshoe Crabs during the 1990s was largely due to increased use as whelk bait (Smith et al. 2009b). Coast-wide landings of all four whelk species have increased 62% since 2005 (ASMFC 2013), although harvest of Horseshoe Crabs for bait has declined since 1998 through quota regulations and has been stable since mid-2000s (Table 4 in the Supplementary Material).
Between 1970 and 1990, commercial harvest increased from less than 20,000 pounds (9 mt) to above 2 million pounds (907 mt) annually (ASMFC 2013). Reported harvest increased rapidly during the late 1990s to nearly 6 million pounds (2,722 mt) or nearly 3 million animals in 1998 (Table 4 in the Supplementary Material). Since the 1998, harvest quotas and season closures have been set by the Atlantic States Marine Fisheries Commission (ASMFC 1998), a marine reserve has been established, and bait-saving devices have been used widely by commercial fishers. In recent years, reported bait landings ranged from 600,000 to 750,000 animals and more males have been harvested then females because states have established sex-specific restrictions designed to reduce harvest of females (ASMFC 2013).
Harvest of Horseshoe Crabs in the Gulf of México is regulated only at the state level. In Northeast Gulf region, harvesting of Horseshoe Crabs by shrimp trawlers began in the early 1980s (Rudloe 1982). In 1999 more than 110,000 Horseshoe Crabs were harvested from the west coast of Florida. In that year, fishermen were experiencing a bait shortage due to increased regulation of Horseshoe Crabs in Delaware Bay. Horseshoe Crabs can be an easy source of money for those willing to collect them off the beaches. Raffield Fisheries near Port St. Joe, FL estimated that they purchased approximately 99,000 Horseshoe Crabs in 44 days mostly from unemployed workers who had been encouraged to collect Horseshoe Crabs for bait (Wallace 1999). Since 2000 only 14,683 Horseshoe Crabs have been harvested for bait along the west coast of Florida based on data compiled from reported trip tickets (Gerhart 2007). Bait harvest in Florida is regulated and does not present a threat at this time.
Although Horseshoe Crab harvesting is illegal in México due to the species’ risk status (see the Conservation Actions section below), there are increasing reports of small-scale poaching of adults by local watermen who set shallow-water nets at the mouths of coastal lagoons during the incoming phase of the tidal cycle and hand-pick the animals (J. Zaldívar-Rae, Anáhuac Mayab University, pers. obs). In Chuburná, Yucatán, this activity coincides with the Horseshoe Crab spawning season (J. Zaldívar-Rae, Anáhuac Mayab University, pers. obs.), and anecdotal accounts suggest this harvest occurs in other localities. Poached Horseshoe Crabs are sold clandestinely and solely used as an alternative to commercial bait species (Libinia dubia and Cardisoma guanhumi crabs) in the artisanal octopus (Octopus maya) fishery of Campeche and Yucatán, which takes place between August and December. Locals state that poaching only occurs when octopus prices in international markets are low or commercial bait becomes scarce or too expensive. However, there are also anecdotal reports of a growing demand for large amounts of Horseshoe Crabs from medium-sized ship operators, who capture common octopus, Octopus vulgaris, in deeper waters during weeks-long trips. Poaching, selling and buying Horseshoe Crabs are Federal felonies under Mexican law and are punished with up to 12 years incarceration and fines of up to US$19,000 (Diario Oficial de la Federación 2014a).
Horseshoe Crabs are harvested by the biomedical industry for the manufacture of LAL, which is used to test for gram-negative bacterial contamination in injectable drugs and implantable medical devices. The LAL test was commercialized in the 1970s and is currently the global standard for screening medical equipment for bacterial contamination (Levin et al. 2003).
Blood from Horseshoe Crabs for LAL production is obtained by collecting adult crabs, extracting a portion of their blood, and releasing them alive. The Federal Drug Administration estimated medical usage increased from 130,000 crabs in 1989 to 260,000 in 1997 (D. Hochstein, Center for Biologics Evaluation and Research, pers. comm.) with a steady increase since that time. Five companies harvest Horseshoe Crab for blood to produce LAL: Associates of Cape Cod (Massachusetts); Limuli Labs (New Jersey); Lonza (formerly Cambrex Bioscience) (Virginia); Wako Chemicals (Virginia); and Charles River Laboratories (South Carolina).
The number of crabs collected and processed by the biomedical industry has increased over recent years. Based on a review of pertinent studies (Rudloe 1983, Kurz and James-Pirri 2002, Walls and Berkson 2003, Horton and Berkson 2006, Leschen and Correia 2010), the ASMFC assumes a 15% post-release, post-bleeding mortality with a range of 5 to 30% mortality depending on factors such as volume bled and handling stress. Under these assumptions, mortality of crabs processed for LAL for 2012 was 79,786 with a range of 31,189 to 152,681 crabs (ASMFC 2013).
Marine Life and Scientific Collection
Horseshoe Crabs are collected for marine life fishery (e.g., aquarium trade) and for scientific collection. Atlantic states are required to report all harvest, including harvest for marine life or scientific collection, to show compliance with the Fishery Management Plan (Marin Hawk, ASMFC, pers. comm). Monitoring program requirement A1 includes annual report of “The use and harvest of horseshoe crabs for scientific research, educational activities, and live trade should also be monitored and must be reported by all states.” (ASMFC 1998). The recent reporting from 2012 indicates that marine life or scientific collection not associated with biomedical harvest involves a few permits issued and relatively small numbers of animals kept (ASMFC 2013). For example in 2012, Massachusetts reported less than 1,000 collected; Connecticut reported that collections were for educational purposes and individuals were returned to open water alive; New Jersey reported a few hundred were collected and most were returned alive; Delaware reported less than 300 collected mostly for research and education; and North Carolina reported approximately 500 collected with half returned alive. The exception is Florida where the marine life fishery is substantial and may be expanding on the west coast, but may be declining on the east coast (Florida harvest data file compiled from trip tickets). On the east coast since 2008 (most recent five years) a mean of 109 trips have collected a mean of 4,938 animals per year (mean 45.3 animals per trip). Although these numbers are small and have declined substantially since 2004, the east coast populations of Horseshoe Crabs are small and could be affected significantly by this harvest. On the west coast since 2008 (most recent five years) a mean of 264 collecting trips have been made annually with a mean of 22,597 animals collected per year (mean 85.5 animals per trip). The magnitude of the threat from marine-life fishery is unknown because population size is unknown (Gerhart 2007). However, approximately half of reported marine-life landings of Horseshoe Crabs are from the Florida Keys (49%; FWC online survey) which have relatively low numbers of Horseshoe Crabs and where there is a relative dearth of suitable adult spawning habitat. If the current population abundance is indeed low, extensive removal of largely first or second-year juveniles due to marine-life landings could hamper the ability of the population to sustain itself (Gerhart 2007).
Historically, Horseshoe Crabs have been considered a non-target bycatch in commercial fisheries targeted at other species and returned to the water (Walls et al. 2002). However, injuries can occur during capture, and these injuries can lead to mortality or altered fitness. Horseshoe Crabs were the most abundant invertebrate bycatch species caught in shrimp trawls in Tampa Bay; 2,867 Horseshoe Crabs were caught during two sampling seasons with the largest catches in the fall (Steele et al. 2002). As part of a tagging study during which Horseshoe Crabs were caught using dredges (Smith et al. 2006), injury rate was 11% (4,459 out of 39,343; D. Smith, USGS, unpub. data). A subjective assessment was that 6% of the total catch (i.e., 2,542 out of 39,343) was severe enough to cause mortality. These injury and mortality rates would apply to bycatch when dredges are used to harvest whelk and to some extent when bottom trawls are used to harvest horseshoe crabs for LAL production. Many Horseshoe Crabs are damaged by hydraulic dredging for Surf Clams (Spisula solidissima) off the Atlantic coast of New Jersey (M. Botton, Fordham University, pers. obs). The significance to population viability depends on the magnitude of bycatch mortality relative to population size and natural mortality. As with any ancillary threat to Horseshoe Crabs, the importance will be greater for a small population restricted to a single embayment than for a large migratory population.
Horseshoe Crabs may have been a common bycatch species of shrimp trawlers in the southern Gulf of México, especially during the 1970s to 1980s, when this fishery experienced a boom in Campeche and few controls on bycatch were in place. However, in a study of bycatch composition among artisanal trawlers fishing Atlantic Seabob (Xiphopenaeus kroyeri) in areas within the Laguna de Términos where Horseshoe Crabs are common, they were not among the invertebrates caught with prawn (Wakida-Kusunoki 2005)
Undisturbed sandy beach is considered to be optimal spawning habitat (Botton et al. 1988), and the availability of optimal spawning habitat is considered to be a limiting factor on population growth (Rudloe 1982, ASMFC 1998). Botton et al. (1988) reported that only 10.6% of Delaware Bay shore on the New Jersey side was optimal spawning habitat. Beach erosion and human development are coast-wide concerns for conservation of beach habitat for Horseshoe Crabs (Jackson and Nordstrom 2009). Loss of sand to erosion exposes parent material, such as peat or fine-grained mud, which tend to be anoxic or low oxygen environments unsuitable for egg development (Botton et al. 1988, Penn and Brockmann 1994, Jackson et al. 2008). Human development per se is not necessarily a threat because Horseshoe Crabs will spawn on beaches in front of houses and do not avoid human activity. Some of the best beach habitat with the densest spawning occurs on sandy barriers associated with coastal development (Jackson and Nordstrom 2009). However, beach driving, which is permitted on some beaches, can result in crushing of stranded Horseshoe Crabs. Shoreline change is a function of both coastal geomorphology and human development, and the purpose of erosion control is mainly to protect human structures (Hapke et al. 2013). Hardening the shoreline as a means of erosion control can result in the loss of habitat suitable for Horseshoe Crab spawning and egg development. Protecting sandy barriers with hard structures, e.g., bulkheads and rip rap, can result in loss of habitat for spawning and egg development by truncating the beach foreshore and creating structures that trap spawning Horseshoe Crabs and increase stranding mortality. Jackson et al. (2015) reported that 40% of shoreline within five New Jersey spawning beaches was fragmented by bulkhead segments and enclaves. Further, between 20 to 100% of bulkheads intersected below the spring wrack line, which directly constricts spawning (Jackson et al. 2015). In contrast, protection or restoration of coastal ecosystems can serve the purpose of reducing risk to vulnerable property (Arkema et al. 2013). For example, beach replenishment can restore or maintain quality habitat (Jackson and Nordstrom 2009), if designed to match natural sediment characteristics (Jackson et al. 2005a, 2005b, and 2007), maintain sediment transport (Jackson et al. 2010), and avoid adverse effects on early life stages and on reproduction through project location and timing.
Impingement on Coastal Infrastructure
There has not been a comprehensive assessment of the extent of coastline with infrastructure at risk for impingement. Within Delaware Bay, Botton et al. (1988) estimated that 10% of New Jersey shoreline was severely affected by bulkheading, and more recent estimates indicate the influence of bulkheading along the New Jersey bay shore has increased (Jackson et al. 2015). Although bulkheading was eliminated from the Delaware shoreline, extensive impingement has been observed at breakwaters formed by rip rap and road overwash at Mispillion Harbor and Port Mahon (D. Smith, USGS, pers. obs). Mitigation measures at power plants have been successful at reducing impingement (see Connecticut’s example in the Conservation Actions section); however, not all power plants within horseshoe crab habitat have been assessed for impingement risk.
In a study of impingement at two power plants on the Indian River a total of 39,097 Horseshoe Crabs were trapped on the intake screens at the Florida Power and Light Cape Canaveral Plant (FPL) and 53,121 at the Orlando Utilities Commission Indian River Plant (OUC) over the 12-month period of the study (Applied Biology Inc., 1980). A previous study conducted in 1975 estimated 69,662 at FPL and 104,000 Horseshoe Crabs were retained annually at the FPL and OUC intakes. This level of mortality can be a major threat to a localized population (Ehlinger and Tankersley 2007).
Solutions to minimize entrapment and mortality have been engineered for some existing and new power plants. For example, through a federally approved National Pollution Discharge Elimination System permitting program pursuant to the section 316(b) of the federal Clean Water Act, the Connecticut Department of Energy and Environmental Protection has required the design and installation of Aquatic Organism Return Systems (AORS) in order to minimize the mortality of aquatic organisms, including horseshoe crabs (Mark Johnson, CT DEEP, pers. comm). Although the level of historical mortality is not known, from the 1970s through the 1990s it is likely that annual mortality of Horseshoe Crabs was in the low thousands. Horseshoe Crabs entered the cooling water intake forebays, climbed on to the intake travelling screens, and ultimately wound up in debris collection pits or baskets where many of them died. The retrofitted AORS returned the Horseshoe Crabs alive to the water body. Narrowing the spacing between the bars of trash racks in front of the intake was another measure that kept many Horseshoe Crabs from passing through to the intake screens. One power plant required periodic monitoring and removal of sediment accumulation near the intake structure to minimize trapping of Horseshoe Crabs. Collectively, these measures have largely eliminated Horseshoe Crab mortality and other negative effects at Connecticut’s coastal power plants.
Water Quality and Pollution Events
Botton and Itow (2009) reviewed studies on water quality and contaminant effects on Horseshoe Crab embryos and larvae. They concluded that current levels of contamination and water quality do not pose a population-level effect for L. polyphemus. They reached a different conclusion for Tachypleus tridentatus an Asian species in Japan where they believe pollution is contributing to population decline. Eutrophication due to excess nutrient loading, particularly nitrogen from anthropogenic sources on adjacent watersheds, is pervasive among coastal systems where Horseshoe Crabs reside. While nutrient enrichment and source shifts are known to affect Horseshoe Crab food web dynamics, these factors do not appear to have a major effect on Horseshoe Crab abundance and distribution where effects have been studied (O’Connell et al. 2003, Carmichael et al. 2004). As a result, unique dietary signatures (based on stable isotope values) have been useful to demonstrate that Horseshoe Crabs show loyalty to local habitat sites and associated food resources, regardless of level of eutrophication. Oil spills represent an acute threat, which depends on timing, magnitude, wind pattern, oil type, and factors that contribute to bioremediation (Venosa et al. 1996). Although there has been a history of oil spills in Delaware Bay, which is a major seaway for transport of oil (Botton and Itow 2009), the effect on the Horseshoe Crab population has not been evident largely because the timing and spatial extent of the spills have not overlapped with Horseshoe Crab spawning. However, an oil spill that coincided with spawning activity with oil washed onto spawning beaches could be catastrophic to a local population (Venosa et al. 1996). In addition to the obvious effect of oil coated animals, studies have examined effects of oil on growth and survival of eggs and early life stages. Laughlin and Neff (1977) observed reduced hatching success in horseshoe crab eggs exposed to 50% water-soluble fraction of No. 2 fuel oil and metabolic stress among 2nd instars at lower concentrations (5 to 10% water-soluble fraction). Oil that does not reach the beaches during spawning and is not collected will weather and lose volatile compounds (Strobel and Brenowitz 1981). The heavier oil that remains has been shown to affect development and survival with a minimum lethal dose of 2.25 mg/l in suspension (Strobel and Brenowitz 1981). A study lead by Ruth Carmichael (Dauphin Island Sea Lab, pers. comms) examined potential effects of the Deepwater Horizon oil spill (DWHOS) on young Horseshoe Crabs within the northern Gulf of Mexico (Estes et al. 2015). Comparison of molt patterns (size and timing) at Petit Bios Island, Mississippi before and following the DWHOS indicated no evidence of adverse effect to subadult survival. However, they lacked evidence to make inference about effects on spawning adults or population level effects.
“Possible reasons for the decline or small population size of horseshoe crabs in Indian River Lagoon (IRL) include changes in water quality, loss of habitat and increased human harvest. A 188% increase in the human population adjacent to the IRL from 1970 to 1990 resulted in significant decline in water quality, an increase in the prevalence of anoxic sediments and altered sediment composition along the shoreline (Woodward-Clyde 1994). The build-up of muck and anoxic sediments along the shoreline as a result of increased runoff may have reduced areas of optimal spawning habitat…Creation of impoundments around the IRL in the 1960s drastically reduced horizontal and vertical diversity of the shoreline suitable for spawning and nesting.“ (Ehlinger and Tankersley 2007).
Red tides are harmful algal blooms caused by abnormally high concentrations of dinoflagellates. Red tides caused by Karenia brevis are common in the nearshore areas of the Gulf of México, particularly southwest Florida and in the Yucatán Peninsula where Horseshoe Crabs are common. Periodic red tides occur along Florida’s west coast and young Horseshoe Crabs are one of the affected species (Galtsoff 1949). A major die-off of horseshoe crabs occurred in the Indian River Lagoon in July 1999. An estimated 100,000 adult L. polyphemus died in the northern part of the Indian River and the southern portion of Mosquito Lagoon (Scheidt and Lowers 2001). However, no other species were affected; thus, algal blooms or pollution leading to low oxygen or low water quality probably did not cause the die off.
In Yucatán, red tides are common, with the latest events taking place in 2003, 2008 and 2011. These last occurrences were due to blooms of Scripsiella trochoidea, Cylindrotheca clostridium and Nitzchia longissima, although other species were also detected (Ortegón et al. 2011, Herrera et al. 2010). There are reports of severe impacts of harmful algal blooms on commercially important fish and benthic organisms such as octopus, O. maya, and sea cucumbers, Isostichopus badionotus in the northern coast of Yucatán (Zetina et al. 2009), but effects on Horseshoe Crabs, although likely, have not been measured.
Limulus adults, as well as embryos and larvae, are eurythermal (Botton and Itow 2009), so direct mortally from rising water temperatures is probably less of a threat to the species than sea level rise. The obvious threat from climate change to coastal habitat is the loss of spawning habitat due to sea level rise and storms (Arkema et al. 2013). Sea-level rise will increase the rate at which these habitats disappear, and it will increase the likelihood that Horseshoe Crab spawning habit becomes compressed between the rising sea and existing housing and other infrastructure (Loveland and Botton 2015). Over the last century, sea level has risen by 20-40 cm depending on where you are on the coast, due to sea level rise and local sinking of land. Along the Florida shore, the sea level is rising 2.5 cm every 11-14 years. Other effects of climate change, such as increasing water temperatures and altered storm frequency and severity, could affect timing of spawning activity in some regions. Changes in timing of spawning activity would have uncertain consequences to Horseshoe Crab population viability, but could have ecosystem effects by creating mismatches in predator-prey dynamics, particularly those involving shorebirds and horseshoe crab eggs (McGowan et al. 2011a, Smith et al. 2011).
Atlantic States Marine Fisheries Commission Management Plan
As described above, L. polyphemus is harvested primarily for two purposes: (1) as bait for commercial fisheries and (2) for collection of their blood for use in the biomedical industry. The bait harvest along the Atlantic coast of the U.S. is regulated by the ASMFC. The mission of the ASMFC is to promote “better utilization of the fisheries, marine, shell and anadromous, of the Atlantic seaboard by the development of a joint program for the promotion and protection of such fisheries, and by the prevention of physical waste of the fisheries from any cause”. The ASMFC serves as the deliberative body that coordinates the conservation and management of the shared near-shore fishery resources, and is comprised of the 15 Atlantic coast states, as well as the U.S. Fish and Wildlife Service (FWS) and the National Marine Fisheries Service (NMFS). Each state is responsible for implementing management measures in its jurisdiction in a manner consistent with the regulations set forth in the ASMFC Interstate Fisheries Management Plan (IFMP; ASMFC 1998) and associated addendums, with the caveat that the States can always implement more conservative measures should they desire.
A management board exists for each of the species under management by the ASMFC and is responsible for developing and implementing a management plans for the species. To do so, the management board relies on input from technical committees and an advisory panel. A Horseshoe Crab Technical Committee and a Delaware Bay Ecosystem Technical Committee were formed to provide scientific advice to the Horseshoe Crab Management Board. These technical committees are composed of technical staff from states involved in the ASMFC, as well as representatives from NMFS, USFWS, and members of academia. They assess and interpret relevant data on Limulus and associated shorebirds, analyse the likely impacts of possible management actions, and make science-based recommendations to the Management Board.
The ASMFC Horseshoe Crab Management Board developed an IFMP for Limulus in October 1998 (ASMFC 1998), and seven Addenda have been approved since then to reflect improved understanding of exploitation and population dynamics. The Limulus management plan explicitly incorporates objectives for both sustainable prosecution of the fishery as well as continued function in the trophic ecology of coastal systems, i.e., use by migratory shorebirds and sea turtles. The migratory shorebirds that utilize Delaware Bay as a critical stopover includes the federally threatened red knot (FWS–R5–ES–2013–0097, http://www.fws.gov/northeast/redknot/pdf/2014_28338_fedregisterfinalrule.pdf). Beginning in 2000, harvest of Limulus was managed by a quota system for each Atlantic coast state, based on an across-the-board reduction from an established reference period landing (Table 4 in the Supplementary Material).
The harvest quotas established by ASMFC govern state-specific harvest regulations (Table 4 in the Supplementary Material), although individual states have the option of imposing more conservative measures. The Delaware Bay states have had the most complex regulatory history because of the link between horseshoe crabs and shorebirds within Delaware Bay. Delaware Bay harvest regulations are summarized following a bulleted summary of state-specific regulations for the non-Delaware Bay states.
Harvest quotas for the four Delaware Bay states in 2013 and 2014, as recommended by the ARM process and adopted by ASMFC, are provided in Table 5 in the Supplementary Material. A moratorium on harvest remains in effect in New Jersey.
In an effort to further ensure a sustainable Horseshoe Crab population in the mid-Atlantic region, NMFS established a 3,885 km² no-take zone in Federal waters outside the mouth of Delaware Bay (Figure 4 in the Supplementary Material). Harvest or possession of Horseshoe Crabs aboard vessels within the Carl N. Shuster Jr. Horseshoe Crab Reserve is prohibited and this area is known to have very large concentrations of Horseshoe Crabs (Botton and Haskin 1984, Botton and Ropes 1987). An exempted fishing permit for capture of crabs in the Reserve for biomedical purposes has been issued to Limuli Laboratories Inc. by NMFS since 2001. The permit allows the capture of up to 10,000 animals annually, and requires the permittee to collect demographic and morphometric data on the collected animals.
Northeast Gulf of México Region
Commercial harvest of Limulus in the Gulf of México waters of the U.S. is limited due to their relatively low abundance. Currently, horseshoe crab harvest in the Gulf of México is not addressed by the Gulf States Marine Fisheries Commission (GSMFC) although they have discussed the need for regulations. Florida‘s regulations, which apply to both the Atlantic and Gulf sides of the state, control harvest of Horseshoe Crabs for commercial use and the “marine life” trade via daily bag limits. Management of Horseshoe Crabs on the west coast involves the same regulations as those on the east coast (with no bait harvest quota). No other Gulf state regulates harvest of horseshoe crabs.
As for all species in the “en peligro de extinción” or “amenazada” protection categories (see below), harvesting of Horseshoe Crabs is forbidden by federal law in México, unless it is proven that a) harvesting quotas are below levels that allow the natural replenishment of the harvested wild population, or b) they are the result of controlled reproduction, in the case of captive organisms, or c) when the use of parts or tissues is involved, it will not negatively affect the population or modify the specimens’ life cycle, or d) when the collection of derivatives from specimens is involved, loss of these derivatives or the procedure used to collect them will not permanently harm specimens (Diario Oficial de la Federación 2014b). In addition, should harvested specimens or their parts and derivatives come from wild populations, it must be proven that controlled reproduction programs are in place to replenish these populations. In case harvested specimens come from captive populations, controlled reproduction of specimens in these populations must support governmental programs aimed at replenishing wild populations (Diario Oficial de la Federación 2014b). Currently, there are no legal Horseshoe Crab harvesting operations in México. However, as mentioned above, there is a seasonal clandestine market for Horseshoe Crabs as bait, fuelled by small-scale poaching. So far, this illegal trade has not been addressed by either local or Federal authorities.
Alternative Bait Strategies
Innovative efforts to reduce the quantity of Horseshoe Crabs required to meet the demand for the bait industry have produced some gains. Beginning in 1999, the fishing industry began to adopt the use of bait bags, wherein smaller portions of Horseshoe Crabs could be used as bait in a single conch pot, as opposed to a whole animal. This practice has expanded along the coast and resulted in more efficient use of Horseshoe Crabs as bait for the conch fishery.
An alternative bait, which chemically mimics the Horseshoe Crab, has been developed and was commercially marketed for the first time in 2013 (Wakefield 2013). Preparation of the bait by individuals is also possible via a published recipe. The product is a result of years of research by a team of researchers from the University of Delaware. While the product contains Horseshoe Crab tissue in its formulation, the amount is small enough such that widespread use of the artificial bait would significantly reduce bait harvest. Currently, it remains uncertain whether the fishing industry will widely adopt the artificial bait.
Increased prices and reduced availability of Limulus in the U.S. bait trade has motivated dealers to import Asian Horseshoe Crab species (Carcinscorpius rotundicauda, Tachypleus gigas, Tachypleus tridentatus) for use as bait in the domestic conch and eel fisheries. These importations are viewed as a significant threat to native Limulus populations due to possible introductions of harmful parasites and pathogens into U.S. waters (Appendix). In addition, C. rotundicauda is known to accumulate tetrodotoxin, and concerns that eel and whelk may accumulate this potentially lethal neurotoxin argue against continued importation of Asian species. For these reasons, many individual states have implemented regulations, or initiated the regulatory processes, to prohibit possession Asian species and/or their use as bait in their fisheries. The U.S. Congress is deliberating on legislation that would expand the reach of the federal government for designating non-native species as invasive or injurious and prohibiting their importation. The applicable statute (Lacey Act) presently applies to various taxa, including crustaceans, but not to chelicerates. While long-term harmful effects can result from relatively few or isolated introductions, it appears the threat posed by importation of Asian species is being addressed as expediently as possible.
Increased prices for Limulus in the bait market may also be responsible for increased incidences of illegal harvest. Charges were brought in two cases of illegal harvest in New York waters in the summer of 2013. The amount of illegal harvest in the mid-Atlantic region is unknown, although awareness by enforcement authorities is increasing (Goodman and Nir 2013).
In 1994, Horseshoe Crabs in México were assigned the status “en peligro de extinción” (literally, “in danger of extinction”), the highest risk category for extant species in the Mexican Official Standard for Mexican species at risk (NOM-059-ECOL-1994; SEDESOL 1994). Under that Standard, a species is assigned such status if “its distribution or population size have drastically decreased, putting its biological viability at risk throughout its range, as a result of the destruction or drastic modification of its habitat, severe restriction of its distribution, over-exploitation, disease, and predation, among other causes” (SEDESOL 1994). This risk category overlaps IUCN’s “Critically Endangered” and “Endangered” categories, and its definition is still in place (Sánchez et al. 2007).
It is unclear which of the “en peligro de extinción” criteria were deemed to be met in the case of Mexican Horseshoe Crabs, but it is known that in the 1994 NOM-059 exercise some species were included in the endangered species list purely on the basis of qualitative impressions and the individual interests of invited experts (Sánchez et al. 2007). There have been subsequent revisions of inclusion criteria to the NOM-059 standard (e.g. NOM-059-SEMARNAT-2001; NOM-059-SEMARNAT-2010; SEMARNAT 2002, 2010) and an objective risk assessment procedure (the “Method for the assessment of extinction risk of Mexican wildlife species”, MER; Sánchez et al. 2007) has been introduced, but to this day Mexican Horseshoe Crab populations have not been officially subjected to this procedure. Nevertheless, when the MER procedure was established it was decided that the status of no species in the NOM-059-ECOL-1994 list would be modified if data were not sufficient to apply the MER (Sánchez et al. 2007). Hence, L. polyphemus is still regarded as “en peligro de extinción”.
The MER includes four criteria with a variable number of levels, each affording 1 to 4 points to a general score: a) amplitude of the species’ distribution in México (4 levels, 1-4 points, respectively: wide, moderately restricted, narrow and very narrow); b) condition of the habitat (3 levels, 1-3 points: favourable, intermediate, hostile/very limiting); c) intrinsic species vulnerability on the basis of its biology and life history traits (3 levels, 1-3 points: low, medium, high); and d) impact of human activity on the species (3 levels, 2-4 points: low, medium and high) (SEMARNAT 2010; Sánchez et al. 2007). According to these criteria applied as an informal assessment, horseshoe crabs in México have a restricted distribution (5-15% of the Mexican Exclusive Economic Zone; 3 points); arguably, their habitats are in intermediate condition (coastal lagoons conditions vary widely from almost pristine to severely disturbed; 2 points); and human activities can exert medium to high impacts on these habitats, through water and sediment pollution and encroachment or modification of land use in sites adjacent to coastal lagoons (3 to 4 points). Thus far, information on the biological features of the species in México is not enough to establish how vulnerable Mexican populations of Horseshoe Crab may be. However, the only formal quantitative survey of spawning events carried out so far in a Mexican locality, has revealed that abundances of reproductive individuals are relatively low: spawning pairs do not exceed the tens of pairs in a 100 m transect on a peak high tide (J. Gutiérrez and J. Zaldívar-Rae, Anáhuac Mayab University, unpub. data) and reports by locals from other sites suggest that this may be the case throughout most of the distribution in México (J. Zaldívar-Rae, Anáhuac Mayab University, pers. obs). Moreover, spawning seems to be restricted to particular shore conditions within coastal lagoons, so the availability of suitable spawning habitat may also be limited. Based on these elements, the degree of vulnerability of the species may be regarded as medium to high (2 to 3 points). Adding up points in this informal assessment following MER criteria indicates that horseshoe crabs in México should either remain in the “en peligro de extinción” category (12-14 MER points) or at least in the “amenazada” (“threatened”) category (10-11 MER points).
Habitat-Based Conservation Actions
The creation, restoration, or protection of beach or nearshore habitat specifically for the benefit of Horseshoe Crab populations is not common throughout the Atlantic coast. The beach replenishment or fill activities, which occur in several states (e.g., Delaware, New Jersey, Maryland, Massachusetts), are generally justified and pursued for protection of communities and infrastructure, particularly for beaches damaged by storm erosion. A desired result is that these projects also augment horseshoe crab spawning habitat. One notable success is Maryland’s primary Horseshoe Crab spawning Island, which has been replenished with dredge spoils since 2010. This activity has almost doubled the available primary habitat for Horseshoe Crabs spawning in the state and thousands of Horseshoe Crabs have been documented spawning on the Island since that time. Both migratory shorebirds and horseshoe crabs have responded favourably to the replenishment action. Jackson and Nordstrom (2009) outline a management framework for conserving shoreline habitat for Horseshoe Crabs. Although the framework is designed for Delaware Bay, the general principles they follow apply range-wide.
There are no management programs specifically focused on Horseshoe Crabs along the U.S. Gulf of Mexico coast or in México. Neither are there habitat conservation programs explicitly aimed at the species. However, large portions of coastal habitats in the Yucatán Peninsula, including coastal lagoons where Horseshoe Crabs are common and have been reported to reproduce, are under the jurisdiction of both federal and state protected areas with different legal regimes. Although none of the management programs of these protected areas include actions to protect Horseshoe Crabs, protected area administrations pay particular attention and devote considerable efforts to the monitoring and preservation of mangrove forests. This monitoring effort is the result of an amendment to the Federal Law for Wildlife passed in 2007 that forbids and severely punishes any activities that may negatively affect mangrove forests and related ecosystems in Mexico (Diario Oficial de la Federación 2014a, b). In fact, 76.3%, 90.4% and 79% of mangrove forests in Yucatán, Campeche and Quintana Roo, respectively, have been estimated to be within the limits of a federal or state protected area (CONABIO 2009), and hence these ecosystems are subject to management programs. Given that all protected coastal lagoons in the Yucatán Peninsula harbour mangrove forests, it can be said that in México there is at least a legal framework and actions are being taken that incidentally conserve critical habitats for Horseshoe Crabs.
|Citation:||Smith, D.R., Beekey, M.A., Brockmann, H.J., King, T.L., Millard, M.J. & Zaldívar-Rae, J.A. 2016. Limulus polyphemus. The IUCN Red List of Threatened Species 2016: e.T11987A80159830.Downloaded on 28 June 2017.|
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