Malaclemys terrapin is found in brackish coastal waters (salinity range = 0 to 35). Typical habitats include coastal regions, estuaries, lagoons, tidal creeks, mangrove swamps, and salt marshes. In coastal estuaries supporting large watersheds, terrapins can occur far inland up to the terminal end of the salt wedge (e.g., Chesapeake Bay, Hudson River, and Pamlico Sound). Although M. terrapin is found in brackish water, periodic access to freshwater is necessary for long-term health, and populations in high salinity habitats actively seek and drink freshwater which is essential for neonate survival and growth (Dunson 1985; Holliday et al. 2009) and may be a contributing factor to the semi-terrestrial behavior of neonates to 3-year-olds. Hatchlings are exceptionally desiccation tolerant (Figueras et al. 2018).

The diet of M. terrapin consists largely of invertebrates, crustaceans, and mollusks (Tucker et al. 1995, 2018) and a variety of plant materials and algae (Tulipani 2015; Erazmus et al. 2018). In some populations terrapins are predators of the Salt Marsh Periwinkle (Littoraria irrorata), a snail that grazes on epiphytes that grow on Salt Marsh Cord Grass (Spartina alterniflora). Silliman and Bertness (2002) demonstrated through a series of experiments that when periwinkle predators were removed, periwinkles damage and kill cord grass, leaving a barren mudflat and increasing marsh erosion rates. Although terrapins were not removed or tested in this experiment, as a periwinkle predator they are frequently identified as a potential saltmarsh keystone species. This extrapolation suggests that healthy terrapin populations play a role in maintaining primary production in salt marsh ecosystems and stabilizing shoreline vegetation, thus reducing local erosion.

Malaclemys terrapin exhibits considerable variation in size, sexual size dimorphism, and reproductive output throughout its range. In general, along the Atlantic sea board there is a gradient of increasing body size, greater sexual dimorphism, later ages of maturity, and higher reproductive output with increasing latitude (Lovich et al. 2018). Gulf Coast populations do not conform to this latitudinal variation and are larger than their counterparts on the Atlantic Coast at the same latitude. Most comprehensive life history data come from long-term studies of populations in South Carolina, Maryland, and New York, with limited data from other parts of the range focused mostly on reproductive output and females. Nonetheless, the differences in morphological and life history traits contribute to different environmental stressors and management strategies relevant to terrapin populations throughout the range. Careful modeling is needed to compare reproductive traits among populations to assess how total annual reproductive output per individual varies along the latitudinal gradient.

Diamondback Terrapins under natural conditions exhibit high adult survivorship, low reproductive rates, and high egg/juvenile mortality characteristic of most temperate turtles (Mitro 2003). As a consequence of this suite of life history traits, increases in adult mortality result in declining populations (Congdon et al. 1993, 1994). Marked variation in life history traits can be attributed to a broad latitudinal distribution (Lovich et al. 2018). Females from northern populations mature later and at a relatively larger size than those from southern populations. Clutch size and body size appear to be somewhat correlated over the range; females from New York average 12.8 eggs per clutch (Burke et al. 2018), Maryland females may produce as many as 26 eggs per clutch (W.M. Roosenburg, unpubl. data), and smaller females from more southern populations produce only 4–6 eggs per clutch (Seigel 1984). Average annual reproductive output depends on the number of clutches produced per season. In Chesapeake Bay populations, the average female may produce 1–3 clutches of about 13 eggs in a single nesting season (13–40 eggs/year) and detailed nesting data suggest that terrapins rarely forgo nesting annually, but the number of clutches per year may vary (Roosenburg and Dunham 1997). For detailed reviews of reproductive data, see Roosenburg (1994), Butler et al. (2006a, 2006b), Brennessel (2006), Ernst and Lovich (2009), and Lovich et al. (2018).

Malaclemys terrapin demonstrates marked sexual dimorphism, most likely as a fecundity advantage (Carr 1952), with females being considerably larger, with maximum reported population mean female carapace length of 20.1 cm and male carapace length of 13.3 cm (reviewed in Lovich et al. 2018). Female terrapins can reach weights over 2.5 kg in northern populations, but males rarely exceed 0.5 kg throughout the range (W.M. Roosenburg, unpubl. data). The maximum recorded size for an adult female, 32 cm carapace length, is much greater than that of males, 16 cm, from the same population, and the discrepancy between male and female size increases with latitude (W.M. Roosenburg, unpubl. data). Female terrapins mature later and at a larger size than males. The resulting difference in body size is more pronounced in more northern populations. Size differences result in differences in jaw morphology and bite forces that likely contribute to resource partitioning (diet, habitat) between the sexes (Tucker et al. 1995; Roosenburg et al. 1999; Herrel et al. 2018). Throughout their geographic range, population means for hatchlings range from 25–38 mm in carapace length and 6.3–8.8 g in mass (reviewed in Lovich et al. 2018). Maturity is reached at 4–13 years of age in females, and 2–7 years in males, depending on location (see summary in Brennessel 2006). Longevity may be greater than 40 years for some individuals; marked adult females have been recaptured 30 years after their initial capture as adults (R.C. Wood, unpubl. data; W.M. Roosenburg, unpubl. data). Generation time has not been reported in the published literature; available information indicates that it ranges from 13 to 20 years and also increases with increasing latitude (Mitro 2003; W.M. Roosenburg, unpubl. data). Generation times in more southern populations are shorter, as maturity at lower latitudes is attained several years earlier (Lovich et al. 2018).

Nesting requirements vary among subpopulations (Roosenburg 1994, Butler et al. 2018). Malaclemys terrapin embryos have temperature-dependent sex determination (TSD), with females produced at higher incubation temperatures, generating important differences between populations (Jeyasuria et al. 1994; Burke and Calichio 2014). The challenges of climate change, combined with sea level rise and nesting habitat loss (Michener et al. 1997, Woodland et al. 2017) could potentially affect sex ratios in populations. The conservation challenges associated with climate change and TSD as it affects sex ratios generates new concerns for the effective management of terrapin populations throughout their range, particularly when integrated with the nesting habitat loss due to development (Seigel and Gibbons 1995).

Terrapin hatchlings in northern parts of the range also facultatively overwinter in the nest, suggesting that the nesting environment is critical habitat throughout the entire year (Baker et al. 2006, 2018; Graham 2009; Kitson 2016). Post-emergence neonate habitat is poorly known, but after emergence, some hatchlings move to shallow, self-excavated hibernacula in upland areas (Draud et al. 2004, Duncan and Burke 2016) before moving into intertidal marshes where they may remain semi-terrestrial well into their second growing season (Muldoon and Burke 2012). Adult terrapins hibernate communally in depressions in the mud at the bottom of tidal creeks and embayments protected from prevailing weather (Haramis et al. 2011), in undercut banks within the tidal zone, or burrow into creek banks (Yearicks et al. 1981). Southern subspecies (e.g., M. t. rhizophorarum) may be active throughout the winter.

Malaclemys terrapin inhabits the estuarine coastal waters of the United States along the Atlantic Ocean and the Gulf of Mexico, from Cape Cod, Massachusetts to Corpus Christi, Texas (Iverson 1992; TTWG 2017). A breeding subpopulation also is found in the Bermuda Islands (Davenport et al. 2005), which was considered native by Parham et al. (2008) and genetically attributed to subspecies M. t. centrata. The human role in the origin of the Bermuda subpopulation remains under consideration (Lovich and Hart 2018), but it currently is an established and protected population (Outerbridge 2014).

Subspecific distributional information for the morphologically-defined historically recognized subspecies is as follows (noting, however, that molecular phylogeography does not agree with this pattern, but that a revised taxonomy has not yet been formulated; see Taxonomy section):

M. terrapin terrapin: from Cape Cod, Massachusetts southward to the vicinity of Cape Hatteras, North Carolina, where it intergrades with M. terrapin centrata.

M. terrapin centrata: from the vicinity of Cape Hatteras, North Carolina southward to northern peninsular Florida, where it meets and intergrades with M. terrapin tequesta; also on Bermuda.

M. terrapin tequesta: Atlantic east coast of peninsular Florida to Biscayne Bay, where it intergrades with M. terrapin rhizophorarum.

M. terrapin rhizophorarum: Florida Keys westward to the Marquesas Keys.

M. terrapin macrospilota: Gulf coast of peninsular Florida, intergrading with M. terrapin pileata in the panhandle and with M. terrapin rhizophorarum in Florida Bay.

M. terrapin pileata: from the Florida panhandle, where it intergrades with M. terrapin macrospilota, to Louisiana, where it intergrades with M. terrapin littoralis.

M. terrapin littoralis: western Louisiana and the coast of Texas, at least as far south as Corpus Christi Bay.

Anecdotal reports that this species is found in Mexican waters (e.g., Carr 1952) have been disputed (Smith and Smith 1979); there is no credible information indicating occurrence in Mexico (Iverson 1992; Mexican Red List Workshop participants, Sept. 2005).

Range Size: The presence of a coastal species like M. terrapin is difficult to determine because of their cryptic nature and the inaccessibility of much of their habitat. Its distribution is discontinuous along the ~5,000 km of coastline between Cape Cod, Massachusetts, and Corpus Christi, Texas. However, it is unknown if the discontinuity can be attributed to natural biological boundaries or to barriers created by extensive urbanization and industrial development in portions of its coastal habitat. Terrapins inhabit salt marshes that form on the bay side of barrier islands, and the extent to which they move inland via estuaries varies considerably, but seems to be bounded by waters that are brackish for most of the year.  Nonetheless, terrapin habitat, salt marsh juvenile habitat, and nesting areas, have all diminished greatly because of salt marsh loss and the development of coastal and estuarine waterfront (Kennish 2001; Ner and Burke 2008). No range-wide population estimate is available, but data are consistent in that subpopulations are local, sensitive to anthropogenic perturbation, and many individually studied subpopulations are small, while others are more robust (summarized in Roosenburg and Kennedy 2018).

Diamondback Terrapins are protected from commercial exploitation in most states where they occur, but as of early 2018 commercial harvests remain permitted under regulations in Louisiana (Kennedy, 2018a). Personal possession is permitted under certain conditions in most states (Nanjappa and Conrad 2011; Kennedy, 2018a) but restricted in most cases to no more than two individuals and permits required for possession as pets. Malaclemys terrapin was included in CITES Appendix II in 2013, making its international trade subject to determinations and evaluation that traded volumes are not detrimental to the survival of wild populations.

Hackney (2010) used the standardized Natural Heritage Program to compare the conservation status of terrapins among range states. Of the 16 range states, one (Rhode Island) lists the species as S1, or critically imperiled. Four states list the species as S2, imperiled, while six states list it as S3, vulnerable. Four states consider the diamondback terrapin to be S4, apparently secure. South Carolina has not ranked terrapins according to this scale. State protective listing or harvest regulation in all range states is recommended. In Bermuda, diamondback terrapins have level 2 protection under the Protected Species Order 2012.

Diamondback Terrapins reside in coastal habitats that are intensively used and modified by humans (Kennish 2001, Ner and Burke 2008, Gedan et al. 2009). Safeguarding areas of optimal habitat, particularly foraging, nesting and hibernation sites, from development and other modification is essential for the long-term survival of this species. Terrapins occur in a substantial number of protected areas and their movement is limited (Converse et al. 2015). Mortality in crab pots and habitat fragmentation are suggested to reduce gene flow while signatures of the past commercial trade and movement of terrapins are still detectable in some populations (Converse et al. 2017, Converse and Kuchta, 2018). Terrapins use a variety of estuarine habitats throughout their range and local populations effectively exploit their native habitat. Populations can grow and respond quickly when provided with predator-free nesting areas, high quality juvenile habitat, and protection of adults, in particular reproductive females (Roosenburg et al. 2014, R.L. Burke, unpubl. data), suggesting that estuarine or marine preserves combined with nest predator control can provide an effective tool to restore terrapin populations in areas where numbers are reduced (Roosenburg 2018).

Bycatch reduction devices (BRD) can greatly reduce the number of terrapins that drown in crab traps (Wood 1997; Hoyle and Gibbons 2000; Roosenburg and Green 2000; Butler and Heinrich 2007). By installing BRDs in crab pots, most adult female terrapins are excluded, particular in mid-Atlantic, Northeast Atlantic, and the Gulf coast regions where body sizes are larger. BRDs are also effective in southern Atlantic regions and Florida and perhaps more important because adult females never outgrow the vulnerability to crab pots without BRDs (Grosse et al. 2011). Several states now require BRDs; however, regulations vary considerably across the terrapin’s range (Hackney 2010), as does the intensity of enforcement of regulations. Broad adoption of BRDs will significantly assist the species' prospects (Guillory and Prejean 1998; Rook et al. 2010; Radzio et al. 2013). Perhaps the greatest concern for terrapins is entrapment in derelict or ghost crab pots that are not regularly checked. Several cases have been recorded where >20 individuals have been recovered dead from crab pots that were abandoned, one with more than 100 individuals (Bishop 1983; Roosenburg 1991; Grosse et al. 2011).

Since 2003, research scientists and summer interns at The Wetlands Institute in New Jersey have deployed barrier fencing to reduce road mortality in their study area. Results to date suggest that barrier fencing can be an effective terrapin conservation tool (McLaughlin et al. 2012). Similarly, at JFK Airport in New York, personnel successfully use barrier fencing to keep terrapins off runways (Burke and Francoeur 2014). Combined with nesting habitat improvement adjacent to roadways, fencing can dramatically reduce roadway mortality and help direct terrapins to protected nesting sites (Crawford et al. 2013a,b; Maerz et al. 2018). Finally, roadway warning signals activated at times of peak nesting in high use areas can dramatically reduce nesting female road mortality. Impacts by watercraft also present a problem for terrapins in areas with high boat traffic (Roosenburg 1991; Lester et al. 2013). Furthermore, terrapins appear not to respond to the noise of approaching vessels, increasing the risk of boat impact; thus suggesting that marine speed limits near areas of terrapin congregation (e.g., nesting beaches) would be the only effective mechanism to reduce boat strikes (Lester et al. 2014, 2018).

Nest protection and head-starting programs exist at a variety of locations, and are subject to active research on their effective contribution to population dynamics and establishment success (Herlands et al. 2004; Smeenk 2010; Petrov 2014; Roosenburg 2018). High nest predation remains a key conservation problem for terrapins and control of mammalian nest predators increases nest success (Munscher et al. 2012) and the complete absence of nest predators can contribute to population recovery (Roosenburg et al. 2014). Head-starting can supplement population growth and head-started individuals can establish within native populations. Furthermore, head-started females can become reproductive and some remain in the local population where they were released (Roosenburg 2018). However, the low level of gene flow among populations (Converse et al. 2015) suggests that source animals should only come from the same population they will supplement or from the nearest population for extirpated areas.

One example of effective terrapin conservation and restoration is the Paul S. Sarbanes Environmental Restoration Project at Poplar Island, Chesapeake Bay, Maryland (Roosenburg et al. 2014, Roosenburg 2018). Poplar Island’s restoration has established predator-free nesting habitat and extensive salt marsh habitat for juveniles. There also is an active head-starting program on Poplar Island. The terrapin population on Poplar Island is growing rapidly due primarily to the high nest survival (Roosenburg et al. 2014) and high quality juvenile habitat that ensures high post-hatching juvenile survivorship (Roosenburg 2018). Although head-starting does contribute to population growth, its impacts are secondary to that of the habitat creation and protection (W.M. Roosenburg, unpubl. data). The Poplar Island example illustrates the basic requirements for population recovery which suggests that estuarine reserves or sanctuaries that include both nesting and juvenile habitat, combined with nest predator control, offer an excellent strategy for terrapin population recovery.

Further research on terrapin conservation is needed, particularly towards understanding individual and seasonal movement patterns, habitat use at different life stages, and refinement of the taxonomy. Although some populations and areas are well studied, other parts of the range have less data, particularly the Gulf Coast and the populations within Pamlico Sound and the estuarine waters of North Carolina, and southern New England. Most states have some form of ongoing terrapin research, but the expansive range and the need for boats to access key habitats limits both surveys and detailed population studies. Recent work also has suggested that climate change will dramatically affect habitat availability for terrapins, particularly the juvenile dependence on saltmarsh habitat, which is predicted to decline dramatically due to sea level rise (Woodland et al. 2017). Further study of the effects of sea level rise is needed to evaluate how habitat and resources will be affected.

Although these conservation targets have been known for years, declines in terrapin populations have not been fully addressed for a variety of issues. Conservation measures, such as establishing requirements for bycatch reduction devices in crab pots, have often been effectively resisted by commercial fishermen through political lobbying efforts. Furthermore, even with regulations in place, lack of compliance and enforcement results in ineffective use rates (Radzio et al. 2014). Finally, conflicts between the preservation of nesting habitat shoreline and the development and protection of high-value coastal real estate seem to preempt terrapin conservation measures. Despite the slow pace of terrapin conservation, some progress has been made to reduce road mortality in some areas (Crawford et al. 2013a,b) and population recovery and growth have been observed in at least one population (Roosenburg et al. 2014). Additionally, the technology to reduce terrapin capture in crab pots has made substantial progress (Roosenburg 2004, Chambers and Maerz 2018) as range-wide fine-tuning of bycatch reduction devices (BRDs) continues (Schwenter and Arndt, unpubl. data).