Author: Conor McGowan (CM), USGS Alabama Cooperative Fish and Wildlife Research Unit, School of Forestry and Wildlife Sciences, Auburn University, Auburn, AL.
Two year olds pair, court, defend territories, and excavate scrapes, but typically do not breed (Palmer 1967). In North Carolina 2 and 3 year old birds, banded as chicks returned to near their natal sites and were observed in groups of 2-6 birds (McGowan et al. 2005a) similar to the European Oystercatcher “aggressive club birds” described by Bruinzeel (2004). In the same area three known 3-year old birds (banded as a chicks) were observed incubating a clutch and one reared offspring to fledging (SS, JS).
Murphy (2010) reported that for 168 pairs monitored in Massachusetts in 2005 and 2006, they found nests for 166 pairs, and estimated nesting propensity (probability that a territorial pair will breed) as 0.99 (S.E. 0.008). In Massachusetts, clutch size ranged between 1–6 eggs/nest; for monogamous pairs: 1–4 eggs / nest with mean clutch size of 2.81 (R. Humphrey, unpublished data). In New York (including communal breeding associations with 5 or 6 eggs in some nests) mean clutch size was 3.26 (Zaradusky 1985) and in Virginia it was: 2.33 overall (n = 281), first clutches: 2.78 (n = 129), second clutches: 2.30 (n = 61; Nol et al. 1984). In Massachusetts, 67% of first clutches contained 3 eggs (n = 49) and 62% of second clutches contained 3 eggs (n = 8; R. Humphrey, unpublished data). In Virginia, 78% of first clutches had 3 eggs, 31% of second clutches had 3 eggs (Nol et al. 1984). In North Carolina, mean clutch size was 2.32 (n = 597) from 1998 – 2003 and ranged from 1 – 4 eggs (McGowan 2004). In Georgia mean clutch size was 2.5 eggs (n = 32, Sabine et al. 2006). In Massachusetts, mean clutch size for first nests was 2.82 eggs (S.E. = 0.043, n = 142 nests) 2.39 eggs for second nests (S.E. = 0.14, n = 23) and 2.00 eggs for third nests (S.E. = 0.00, n =5; Murphy 2010). There was no detectable relationship between food supply and clutch size (Nol 1989), at least within the range of food supply available in Virginia. In Argentina and the Galápagos, most clutches are 2 eggs (Nol 1984, A. J. Baker pers. comm.).
Number of repeat clutches is usually one but depends on type and timing of destruction. Three repeat clutches have been reported (Nol et al. 1984, McGowan and Simons unpublished data). In Virginia, most replacement clutches contained 2 eggs (60%, n = 62) in early years of study (Nol et al. 1984), 3 eggs in later years (Nol 1989). This shift may represent changing age structure of the population, with older birds more likely to lay 3-egg replacement clutches (Nol 1989). In North Carolina there were more repeat clutches in 2002 and 2003 in barrier island habitats than on non-traditional, dredge-spoil river islands. However, hatching rates were much higher (low nest predation) on river islands and replacement clutches were often not necessary (McGowan et al. 2005b). In Georgia in 2003 at least two nesting pairs laid three clutches after nest failure (Sabine et al. 2006). In Massachusetts 54% of pairs renested after losing their first clutch and 31% renested after losing their second clutch of the season (Murphy 2010). Murphy (2010) reported 6 instances of renesting if total brood mortality occurred within 1 – 2 days following hatching.
Nest success is highly spatially and annually variable. In North Carolina, Mayfield estimates of daily nest survival (DNS; Mayfield 1961) were significantly greater for 2 and 3 egg clutches compared to 1 egg clutches (CM). In North Carolina mean DNS rate was 0.928 and estimated hatching success rate was 0.133 in 1998 – 1999 (Davis et al. 2001). In an updated analysis, the DNS rate was 0.943 and estimated hatching probability was 0.204 from 1995 – 2003 (n = 852, McGowan 2004) Simons and Stocking (2011) reported DNS as 0.954 and estimated hatching probability as 0.278 from 1995-2010. Nest survival during incubation was 0.973 with a hatching success rate of 0.452 in 2003 – 2004 in Georgia (Sabine et al. 2006, Table 1). In New Jersey, over all DNS was 0.930 in 2005 and 2006 (n = 205, Virzi 2008, Table 1), and hatching probability was lower on barrier island habitats (0.058 S.E. = 0.001) than on isolated islands (salt marshes and dredge spoils; 0.37, S.E. 0.03; Virzi 2008). Similar patterns were reported in North Carolina where the estimated hatching success rate on dredge spoil islands was 0.45 (n = 97 nests) and 0.11 on barrier islands (n = 186; McGowan et al. 2005b). In Maryland in 2003, 44.7% of nests monitored hatched (n = 85 nests; Traut et al. 2006) but DNS estimates were not reported.
Mammalian predation (raccoons, foxes, feral cats, coyotes) has been identified as the primary cause of nest failure in Massachusetts (Murphy 2010), Virginia (Nol et al. 1984), North Carolina (Davis et al. 2001, McGowan et al. 2005b, McGowan and Simons 2006, Simons and Schulte 2010), and Georgia (Sabine et al. 2006, 2008). Overwash was identified as the primary cause of nest failure in Maryland (Traut et al. 2006) and the second biggest cause of failure in North Carolina (Simons and Schulte 2010).
Fledging success: 34% of chicks hatched in Massachusetts fledged (R. Humphrey, unpublished data) and between 54 – 80% in New York (Zaradusky 1985). In North Carolina, the proportion of hatched chicks that reached fledging (# fledged / # hatched) was lower on river islands (0.19) than on barrier islands (0.29); however, the ratio of fledglings per pair was similar between habitats (0.14 – 0.21; McGowan et al. 2005b). In Massachusetts the proportion of young hatched that fledged was 0.377 (S.E. 0.35) in 2005 – 2008 (Murphy 2010). In Florida in 2007, 35 pairs fledged 9 chicks (Hodgson et al. 2008). In North Carolina from 1996-2003, 996 monitored nests produced an observed 118 fledgings (0.118 fledglings per nest; McGowan 2004). McGowan (2004) reported that across all years (1998-2003) and monitoring sites (7 sites) mean number of fledglings per breeding pair was 0.12 (S.E. 0.012). Population production (total number of young) in Virginia varied considerably from 0 (1984: 11 nests) and 3 (1979: 21 nests), to 17 (1982: 41 nests) and 20 (1983: 41 nests; Nol 1989). Many states monitor and report the number of fledgings produced per breeding pair (Table 2), which can be used as an index to track productivity over time. However, these data and estimates do not account for detection rates and may be unreliable to compare productivity estimates. Cross state summary information indicates that the number of chicks fledged per breeding pair monitored ranged from 0.0 in Delaware (between 2002 – 2009, 9 pairs) and 0.53 in Virginia (2001 and 2009, 2416 pairs; Table 2).
More recent studies have attempted to estimate chick survival during the prefledging period and calculate the probability of survival to fledging. In Georgia, daily survival of chicks was reported as 0.991 and the probability of surviving to fledging was 0.329 in 2003 and 2004 (Sabine et al. 2006; Table 3). In Massachusetts the probability of survival to fledging was 0.463 (S.E. 0.046; Murphy 2010, Table 3). In North Carolina from 1999 to 2010, daily brood survival was 0.976 and the probability of survival to fledging (35 days after hatching) was 0.424 (Simons and Stocking 2011, Table 3)). From a radio tracking study in North Carolina, the probability of survival to fledging (35 days after hatching) was 0.438 (Simons and Schulte 2010). That study identified predation as the primary cause of mortality during the pre-fledging period and constituted 54.1% of observed mortality among hatchlings. Predators included Great Horned Owls (Bubo virginianus), Fish Crows (Corvus ossifragus), Feral Cats (Felis catus), Raccoons, (Procyon lotor), American Mink (Mustela vison), and Ghost Crabs (Ocypode quadrata); Simons and Schulte 2010). Mortality due to beach vehicle traffic and other human disturbance accounted for 16% of chick deaths (Simons and Schulte 2010).
Life time success is difficult to study as it requires long-term studies and marked populations. Nol (1989) reported that the number of young raised to independence by a single pair range from 0 over 6 years (7 pairs) to 5 over 3 years (1 pair) in Virginia. Nol (1989) estimated that mean success rate was one successful nesting attempt (rearing chicks to fledging) every 4 years (Nol 1989). Mean percentage life time success rate for all pairs in Virginia was approximately 1 successful nesting attempt in 14 to 21 nesting attempts, based on 1 to 2 renests/year (total number of nest years observed = 174; mea n = 4 year/nest; Nol 1989).
A bird banded as breeding adult (thus assumed to be at least 3 yr old) in 1978, was subsequently observed in 1992, as at least 17 years old (EN). A second bird banded in same study area banded as a breeding adult in 1981, was observed in the winter of 1993, so a minimum of 16 years old. In North Carolina seven birds banded as adults in 2000 and 2001 were observed as 10 or 11 year olds in 2008 (TS). Birds living at least 10 years appear to be common.
Annual survivorship (based on return rates to study area, averaged over 5 yr; n = 30 for each sex) was 85% for both sexes (Nol 1985), but quite variable among years (range: 50% to 90%; Nol 1989). However, at least two birds moved to different, distant sites (> 1 km), thus return rates confound survival with immigration and likely underestimate survivorship. In a reanalysis of these data using Cormack-Jolly-Seber approaches that account for detection probability, annual apparent survival was 0.886 (S.E. 0.040; Murphy 2010).
There have been few complete studies to estimate annual survival using mark/recapture type approaches. In Massachusetts between 2005 and 2008, using a Barker mark/recapture estimator that estimates “true” survival using trapping and resighting data from a specified sampling period and study site as well as from outside the sampling period and study site, Murphy (2010) estimated annual survival as 0.94 (S.E. 0.029, n = 97 birds). Murphy (2010) also reported that apparent survival estimates (estimates that do not account for permanent immigration from the study site) were 0.88 (S.E. 0.053). In North Carolina, apparent survival, estimated using a Cormack-Jolly-Seber model, was 0.89 (S.E. 0.013, no n reported; Simons and Schulte 2010).
Avian cholera was isolated from cultures of liver and heart of adult female found dead on Cape Romaine National Wildlife Refuge low levels of DDE found in tissues, not surprising for a mollusc eater (S. Carolina; Blus et al. 1978). All birds captured and evaluated (n = 101) in Georgia and South Carolina had Lice (identified as Saemundssonia haematopi; Carlson-Bremer et al. 2010). In the same study, no instance of West Nile Virus were dectected for the 34 birds tested (Carlson-Bremer et al. 2010). Other diseases: Chlamydophila psittaci was present in 24% of birds tested (n = 107), Aspergillus fumigatus was present in 29% of birds tested (n = 95; Carlson-Bremer et al. 2010). Mercury contamination was also present in 32 of 44 tested birds (73%; Carlson-Bremer et al. 2010).
McGowan et al. (2005a) reported on 8 sub-adult birds that were originally banded as chicks on their natal territories returning to with-in 42 km of their natal sites (range 2.4 – 42 km), apparently as “aggressive club birds” prospecting for breeding territories and foraging opportunities (Bruinzeel 2004). In North Carolina three birds returned to their natal areas and successfully acquired a territory and bred as a three-year olds (SS, JS). In Massachusetts, 17 of 127 birds banded as chicks between 2005 and 2010 were resighted between 2.28 and 68.80 km from their natal territories (SM). During the same time period 6 of the 127 chicks were resighted as breeders at distances of 0.51 – 49.96 km from their natal territories (SM).
A pair that stays together typically nests in the same territory, although not necessarily in the same scrape, for a number of consecutive years (Tomkins 1954). No new data have been published quantifying breeding site fidelity, though anecdotal evidence suggests it is very common for banded birds to retain the same breeding territory year after year as has been reported in North Carolina (CM), Virginia (AW), and Massachusetts (SM). Immatures begin dispersing from their natal site with parents in mid-August but some birds remain in the natal territory as late as mid-September. Birds from surrounding areas, and possibly farther away, form groups that may contain up to 100 individuals. These groups travel back and forth between different feeding areas, sometimes > 20 km (RH).
Pairs defend breeding territories that are often contiguous (Nol et al. 1984, McGowan et al. 2005b). Also defend feeding territories adjacent to breeding territories, separate from breeding territories, or may defend adjacent territories plus separate feeding territories (Cadman 1979) similar to European oystercatchers (“leap frog territories”; Ens 1992). Feeding territories not defended as vigorously because birds spend most of their time foraging. Feeding territories change in size with tide, thus often difficult to defend or define (EN). Feeding territories may be in excess of 1,600 m from breeding territories (R. Humphrey, unpublished data). Maximum distance observed travelling during breeding season in Massachusetts about 3 km (RH).
In the winter birds tend to gather in large roosting flocks, with many birds observed in Cape Romain National Wildlife Refuge, South Carolina, and in areas of coastal Florida and Georgia. Many banded birds (trapped as breeders in North Carolina) have been observed using the same roosting sites in successive winters in Florida (TS, P. Leary, and D. Leary unpublished data). Though some birds have been observed spending successive winters in greatly different locations; of the birds banded as breeders in North Carolina and seen or captured during winter months (n = 137 individuals), 48 times birds switched wintering locations either within season or across years (416 total observations; TS).
Historic: Density varies with habitat quality, food availability, and extent of human disturbance. From New York south to Georgia, occupies favorable habitat in coastal areas. Density depends on habitat type, with higher densities occurring on dredge spoil islands in areas where humans occupy nearby mainland (sand) beaches (Cadman 1979, Leck 1984, Lauro and Burger 1989, Lauro et al. 1992). From Georgia south, including the Gulf Coast, distribution is local and sparse (Burleigh 1958, Degange 1978). Historically the largest number of breeding pairs in Virginia occurred on the barrier islands but population trends over time are unclear, due to variation in counting effort and spatial coverage (Nol and Humphrey 1994, Davis et al. 2001). The number of breeding pairs counted in New York has doubled since 1986 (Humphrey 1990, Litwin et al. 1993). Earlier increases attributed to reduced hunting pressure and creation of sandy dredge spoil islands used for nesting (Zaradusky 1985).
Davis et al. (2001) compiled breeding pair count data from Atlantic coast states and reported 1624 pairs coast wide. Though previously extirpated from the Northeast region (Bent 1929), in recent decades populations have been expanding northward with the first successful recorded breeding attempt in Nova Scotia, Canada in 1997 (Mawhinney et al. 1999). Numbers of counted birds or breeding pairs have increased in Massachusetts and Nova Scotia (Mawhinney et al. 1999, Paton et al. 2005. Cross year mean breeding abundance from state counts and monitoring efforts between Florida and Maine was 2544 breeding pairs (Table 3). In Massachusetts in 2009, observers reported 382 adults (178.5 pairs) during the breeding season (Melvin 2010) up from 154 in 1991 (Meyers 1998 as cited by Davis et al. 2001). Wilke et al. (2005) counted 588 breeding pairs in Virginia on barrier islands, Atlantic coast bays and in the Chesapeake Bay, up from 255 pairs in 1999 (Davis et al. 2001). Wilke (2009) reported a further increase in Virginia to 731 breeding pairs in 2008, partly attributed to actual population increase but also to increased counting effort and coverage. In Maryland Traut et al. (2006) counted 108 breeding pairs of Oystercatchers with 57 pairs in the Chesapeake Bay and 51 on the Atlantic Coast; this represents a state-wide increase from 75 breeding pairs in 1998 (Davis et al. 2001). In North Carolina in 2010 there were 369 breeding pairs counted (SHS). In South Carolina, counts in 2001, 2002 and 2003 averaged 1105 adults, with 407 known breeding pairs in 2002 and 397 known breeding pairs in 2003 (Sanders et al. 2008). Total of state by state average abundance was 2544 breeding pairs and 4592 individuals (Table 4) over varying number of years in each state. These estimates come from unmarked populations and use varying counting methods across states and may not be useful for examining temporal trends or spatial patterns.
There are no recent breeding season population estimates published for Oystercatchers in the Gulf of Mexico states. They are uncommon in Louisiana, confined to secluded coastal islands (Lowery 1974). In Alabama they are uncommon breeders but permanent residents in coastal Mobile Co. (Imhof 1962). Considered a resident of Florida, where once common (Sprunt 1954), and Georgia since the first records were kept (Anon 1981). Listed as species of Special Concern in Florida and as a species of high conservation concern in Alabama (Mirarchi et al. 2004).
Historic counts exhibit large variability. Christmas Bird Count data indicate a low count of 1,106 in 1972 and a high count of 9,062 birds in 1970 (Nol and Humphrey 1994). Since 1975, numbers have been less variable (range 1,134–3,558; Nol and Humphrey 1994). Some winter population counts (particularly from Charleston, SC) were high in the late 1960s, but the total number of wintering birds appears to have stabilized around 2,000 individuals (Nol and Humphrey 1994). In 1999 a wintering count in the Southeastern states recorded 5,785 birds, with 1883 birds in Virginia, 567 birds in North Carolina, 3098 in South Carolina and 237 were counted in Georgia (Nol et al. 1999). A U.S. coast wide aerial count conducted between November 2002 and February 2003 estimated there were 10,971 (S.E. 298) wintering American Oystercatchers between Texas and New Jersey (Brown et al. 2005). Locally, South Carolina winter counts averaged 3136 individuals in the winter from 1999-2002, with the vast majority of birds using the Cape Romaine National Wildlife Refuge during the winter months (Sanders et al. 2004). Cross year mean wintering abundance from state counts and monitoring efforts between Florida and New Jersey was approximately 11,000 individuals (Table 4). These counts varied in counting methods and frequency across states.
Davis et al. (2001) summarized population status and counts, reporting on differential trends; the Northeast Atlantic coastal states (New Jersey – Maine) had increasing trends while Southeastern Atlantic states (Virginia – Florida) declining or stable trends in the Late 1990’s. There was substantial concern over the Virginia breeding population which had apparently declined from 619 pairs in 1979 to 255 pairs in 1999 (Davis et al. 2001). However, more recent counts estimated Virginia’s population at 588 in 2003 (Wilke et al. 2005) and 731 pairs in 2008. Wilke (2009) argue that the change in abundance between 2003 and 2008 was due to an actual population increase; however it is hard to determine if recently reported increases in breeding abundance across the range are due to changes in counting efforts and techniques or actual population increases. The change in estimated abundance could be due to detectability in count issues and a shift in habitat use from barrier islands to non-tradition breeding habitats such as salt mash islands in coastal bays. Earlier counts generally only looked at barrier island habitats. Similar patterns of shifting habitat use have been reported in New Jersey (Virzi 2008) and North Carolina (McGowan et al. 2005b).
Population simulation modeling indicates that population growth is most sensitive to adult and sub-adult survival rates (Davis 1999, Simons and Schulte 2010). Estimated adult survival is high (> 0.88 annually) and fairly stable across years and locations (Murphy 2010, Simons and Schulte 2010). Few studies have directly quantified the causes of adult mortality, though anecdotal reports indicate predation from raptors and possibly owls is at least one source of mortality (C. McGowan, personal observation).
Egg and chick predation, and storm and high tide wash out reduce breeding success (Johnsgard 1981, Nol et al. 1984, Lauro and Burger 1989, McGowan 2004, Sabine et al. 2005). McGowan (2004) reported that red fox control efforts were associated with increases in DNS in North Carolina however, and increases in population abundance following predator control in 2002 are confounded by major changes in the habitat due to a hurricane event in 2003 (Simons and Schulte 2010). Simons and Schulte (2010) reported major increases in nest survival and in the years immediately following Hurricane Isabelle in coastal North Carolina, compared to the years before. Simons and Schulte (2010) suggested a boom/bust population cycle where oystercatchers rely on major disturbance events every few years (i.e., hurricane storm surges) to create conditions for spikes in productivity to sustain the population through intervening years of low productivity.
Population growth may be regulated by availability of suitable nesting habitat and breeding territories. Nests on high ground tend to be most successful (Lauro and Burger 1989); in some locations high elevation nest sites appear to be limited, so pairs nest communally (Lauro et al. 1992). As communal nests are generally less successful, lack of nest sites probably limits population growth (E. Nol, unpublished data). Evidence of habitat shifts to non-traditional sites such as dredge spoil islands (McGowan et al. 2005b), marsh islands (Wilke et al. 2005, Virzi 2008) or even roof tops in Florida (Douglas et al. 2001) suggests that habitat may limit breeding opportunities for sub-adult birds. Growing issues such as continued urbanization or sea-level rise might lead to further nesting and foraging habitat limitation (Erwin et al. 2006). Human recreational disturbance may also indirectly affect productivity and spatial distribution of breeding territories (Novick 1996, Davis 1999, George 2002, McGowan 2004, McGowan and Simons 2006, Sabine et al. 2008).