Report on the Current Conditions for the Saltmarsh Sparrow (Ammospiza caudacuta)

U.S. Fish and Wildlife Service August 2020

1 Suggested citation: U.S. Fish and Wildlife Service. 2020. Report on the current conditions for the saltmarsh sparrow. August 2020. U.S. Fish and Wildlife Service, Northeast Region, Charlestown, RI. 106 pp.

Cover: Saltmarsh Sparrow (photo credit: Evan Lipton)

2 Executive Summary

This report describes the species needs, threats, and current conditions for the saltmarsh sparrow. It is intended to reinforce and support conservation planning for this species by the U.S. Fish and Wildlife Service (Service) and partners. As conservation measures are tested and implemented, the Service intends to expand this report to include an assessment of future conditions that reflects the effectiveness of on-the-ground implementation measures to slow or reverse saltmarsh sparrow population declines.

The saltmarsh sparrow (Ammospiza caudacuta) is a tidal marsh obligate songbird that occurs exclusively in salt marshes along the Atlantic and Gulf coasts of the United States. Its breeding range extends from Maine to Virginia including portions of 10 states. The wintering range includes some of the southern breeding states and extends as far south as Florida. Nests are constructed in the salt marsh grasses just above the mean high water level, and they require a minimum of a 23-day period where the tides do not reach a height that causes nest failure. Across its range, the saltmarsh sparrow is experiencing low reproductive success, due primarily to nest flooding and predation, resulting in rapid population declines. Forty-eight percent of nests across the breeding range failed to produce a single nestling from 2011 to 2015. Although it has not been quantified, there is strong evidence for range contraction at both the northern and southern limits of the breeding range. Furthermore, breeding individuals are not evenly distributed across the entire range, with approximately 78 percent of the breeding population breeding in marshes of the mid-Atlantic states.

At the northern end of their range, saltmarsh sparrows hybridize with the closely related Nelson’s sparrow (Ammospiza nelsoni subvirgata). Nelson’s sparrows are considered marsh generalists, occupying more brackish and inland marshes than the specialist saltmarsh sparrow, although both species can occupy the same patches in salt marshes in Maine and New Hampshire. Hybrid individuals may be able to occupy a wider range of habitat types that may not be impacted by sea-level rise at the same rate as tidal marshes. However, the extensive literature on hybrid populations suggests hybrid individuals typically have lower nest success than pure-saltmarsh sparrows, and that the two species are remaining distinct.

While the species still occupies the majority of its historical range, the number of individuals within the breeding range has significantly declined since 1998. Based on surveys in 2012 the population was estimated at 60,000 individuals, having declined at an average of 9 percent per year across the range since 1998. Projecting those declines through 2020 we estimate that the current populations is approximately 28,215 individuals. This represents a decline of 87 percent from the 212,000 individuals estimated in 1998.

Numerous threats have been identified as impacting the saltmarsh sparrow and/or its habitat. We assessed the following threats acting on the saltmarsh sparrow: (1) habitat loss, fragmentation, and degradation; (2) the effects of climate change; (3) hybridization; (4) predation; (5) contaminants; and (6) other factors influencing the species such as disease and altered food webs.

Tidal marshes across the Atlantic coast are being lost or degraded at a rapid rate due to a combination of historic and on-going direct anthropogenic impacts (e.g., development, ditching,

3 shoreline hardening, etc.) and the impacts of sea-level rise. Range-wide loss and degradation of tidal marsh habitat has resulted in smaller, more fragmented habitat patches, as well as a disproportionate loss in suitable nesting habitat for the saltmarsh sparrow. Site specific rates of habitat loss have been correlated with multiple localized factors, including nutrient input from agricultural and urban runoff, shoreline hardening, geomorphic setting, tidal range, and presence of invasive plants or animals. Although our understanding of how these variables interact, exacerbate, or ameliorate habitat loss is less understood, we have concluded that these factors do influence how resilient a marsh is or will be to the impacts of sea-level rise.

In addition to habitat loss, the frequency and duration of marsh flooding across the species’ range has also directly impacted reproductive success for the species. This has been attributed primarily to sea-level rise, but also increased precipitation and storm intensity during the breeding season as a result of climate change. While flooding impacts the breeding range uniformly, predation exerts a greater impact on the southern portion of the breeding range. We found little certainty for the potential additive impacts of hybridization, disease, altered food webs, environmental contaminants, adult or juvenile mortality during migration, or mortality on the wintering range. After review of the threats, we identified low reproductive success, and the loss of high marsh nesting habitat, as the two primary threats to the viability of the species. Both of these threats are driven primarily by sea level rise and exacerbated by historic and ongoing anthropogenic impacts.

We also summarize existing conservation policy and management practices which are currently affording some level of protection for the species. These include tidal marsh land ownership, federal and state laws, habitat restoration and protection activities, and conservation planning tools under development. We were not able to compile a comprehensive list of recent conservation actions to quantify the number of acres protected or restored. We also do not have an assessment of what would be required to achieve population level benefits to saltmarsh sparrow. There are efforts underway to compile this information but it was not available for this report.

The final section of this report evaluated the saltmarsh sparrow’s current condition based on our previously described evaluation of the species’ needs and the threats facing individuals and their habitat. This evaluation incorporated an assessment of the species’ population resiliency and species redundancy and representation. Resiliency metrics that were evaluated using the best available scientific information include habitat quality and quantity (i.e., total marsh habitat, marsh size, and availability of high marsh), population demographics (i.e., adult and juvenile survival, nest success and survival), and habitat alterations (i.e., human marsh alterations, development, and tidal restrictions). These factors were assessed within each of four analytical units across the species range. Three analytical units were identified in the breeding range (Northern, Central, and Southern) and one analytical unit in the wintering range. The units were defined based on variations in salt marsh habitat characteristics, tidal regimes, rates of sea-level rise, and other factors influencing saltmarsh sparrows and their habitat.

To look at redundancy, we evaluated the species’ distribution across its entire range. Although the species continues to generally occupy the majority of its historic range, the species is unevenly distributed with a large concentration of its breeding individuals in a relatively

4 small geographic area. During the winter months the population is also concentrated within a similarly compressed geographic area along the southern coast. As such, the population is susceptible to catastrophic events such as major storms or hurricanes, so we determined that the redundancy for the saltmarsh sparrow is moderate.

Representation for the saltmarsh sparrow is described as its ability to adapt to changing environmental conditions. There is some evidence that adult female saltmarsh sparrows will respond to nest loss by modifying the placement of subsequent nests to balance the trade-offs between flooding and predation (i.e. vary height within vegetation or proximity to the upland). Although saltmarsh sparrows exhibit this behavioral adaptation, the species has limited adaptive capacity with regard to nesting habitat characteristics. Nesting continues to be restricted to the high marsh area with few exceptions, and sea-level rise appears to be occurring at a rate that is faster than the species has been able to adapt based on the currently available information. We also conclude that hybridization is not presently providing any additional adaptive capacity for saltmarsh sparrows. As a result, we classified the saltmarsh sparrow as having low representation.

Resiliency for the saltmarsh sparrow is its ability to withstand stochastic events. Based on recent population declines and low nesting success as well as increased impacts to its habitat, we have determined that the resiliency of the species for the Southern and Central analytical units within the breeding range is low. Within the Northern breeding analytical unit, which supports 13 percent of the breeding population, and in the wintering range resiliency is moderate.

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6 Table of Contents

INTRODUCTION...... 9 SPECIES BIOLOGY, LIFE HISTORY AND INDIVIDUAL NEEDS ...... 9

2.1 Taxonomy 2.1.1 Species Delineation 2.1.2 Hybridization 2.2 Species Description 2.3 Range and Distribution 2.3.1 Historical range and distribution 2.3.2 Current range and distribution 2.3.3 Migration 2.4 Life History 2.4.1 Life stages 2.4.2 Diet 2.4.3 Site Fidelity 2.4.4 Breeding 2.4.5 Adult and Offspring Sex Ratios 2.5 Individual Resource Needs: Habitat 2.5.1 Breeding Range Habitat Characteristics and Use 2.5.2 Wintering Range Habitat Characteristics and Use 2.6 Summary of Needs for Individuals

FACTORS INFLUENCING THE VIABILITY OF THE SPECIES ...... 27 3.1 Habitat Loss, Fragmentation and Degradation 3.1.1 Urbanization and Marsh Habitat Loss 3.1.2 High Marsh Degradation, Conversion and Fragmentation 3.1.2.1 Invasive Species 3.1.2.2 Nutrient Enrichment 3.1.2.3 Altered Hydrology Tidal Restrictions and Impoundments Ditching 3.1.2.4 Suspended Sediment Concentrations 3.1.2.5 Shoreline Protective Measures 3.1.3 Human Disturbance and Management Activities 3.1.3.1 Fire 3.1.3.2 Open Marsh Water Management 3.2 Effects of Climate Change 3.2.1 Sea Level Rise and Marsh Subsidence

7 3.2.2 Increase in Storm Frequency, Duration and Severity 3.2.3 Increase in Flooding Events 3.3 Hybridization 3.4 Predation 3.5 Contaminants 3.5.1 Environmental Contaminants 3.5.1.1 Mercury 3.5.2 Oil and Chemical Spills 3.6 Other Factors Considered 3.6.1 Disease 3.6.2 Altered Food Webs 3.7 Conservation Efforts 3.7.1 Habitat Restoration 3.7.2 Applicable Laws 3.7.3 Conservation Status and Tools for Conservation 3.7.4 Conservation Partnerships 3.8 Summary of Factors Influencing Viability

POPULATION NEEDS AND CURRENT CONDITIONS ...... 53 4.1 Methods 4.2 Analytical Units 4.3 Resiliency 4.3.1 Habitat Quantity and Quality 4.3.1.1 Total Marsh Availability (Quantity) 4.3.1.2 Individual Marsh Size (Quality) 4.3.1.3 Proportion, Availability and Quality of High Marsh 4.3.2 Population Size, Growth Rate and Demographics 4.3.2.1 Abundance and Distribution 4.3.2.2 Range Contraction 4.3.2.3 Demographics 4.3.3 Human Alterations to Habitat 4.3.3.1 Degree of Human Alteration (within marsh alterations) 4.3.3.2 Urban and Suburban Development 4.3.3.3 Tidal Restrictions 4.4 Summary of 3 R’s - Current 4.4.1 Redundancy 4.4.2 Representation 4.4.3 Resiliency

BIBLIOGRAPHY ...... 83

8 INTRODUCTION

In this report, we summarize the saltmarsh sparrow life history and individual needs as well as requirements for population level survival and reproduction. We describe the threats influencing these resources as well as other factors affecting species viability, and then we evaluate current levels of population resiliency and species redundancy and representation using available metrics.

The report is intended to reinforce and support conservation planning for this species by the U.S. Fish and Wildlife Service (Service) and partners. In particular, the report can support the work of the Atlantic Coast Joint Venture (ACJV), a regional partnership comprised of state wildlife agencies, federal agencies, and other organizations. The ACJV’s primary focus is on coastal marshes and it has identified saltmarsh sparrow as one of three flagship species that represent this habitat. The ACJV recently completed (fall of 2019) a Salt Marsh Bird Conservation Plan for the Atlantic Coast (https://www.acjv.org/documents/ salt_marsh_bird_plan_final_web.pdf). They also worked with partners to develop complementary species action plans for Saltmarsh Sparrow and Black Rail, which will be published in 2020. The Saltmarsh Sparrow Conservation Plan builds on the Saltmarsh Bird Conservation Plan, including several implementation strategies specifically needed to address the threats to saltmarsh sparrow, as well as state-specific population and habitat objectives. Once these documents are finalized, ACJV staff and partners will focus on implementing the conservation actions called for in the plan; they expect to increase the collective investment of financial and organizational resources dedicated to Saltmarsh Sparrow habitat conservation over the next five years and beyond.

The report will also serve as the starting point for a Species Status Assessment (SSA) for saltmarsh sparrow. As conservation measures are tested and implemented by the ACJV and others, the report will be expanded to include an assessment of future conditions that reflects the effectiveness in on-the-ground implementation measures to slow or reverse saltmarsh sparrow population declines. The SSA will serve as the scientific analysis supporting the assessment of saltmarsh sparrow listing review under the Endangered Species Act (Act). The Service added the saltmarsh sparrow to our National Listing Workplan for the Act as a discretionary action in September 2016.

SPECIES BIOLOGY, LIFE HISTORY AND INDIVIDUAL NEEDS

In this chapter we provide biological information about the saltmarsh sparrow (Figure 1), including the resource needs of individuals and populations. For further general information about the saltmarsh sparrow refer to Greenlaw et al. (2018, entire).

2.1 Taxonomy 2.1.1 Species Delineation

For many years, experts considered the saltmarsh sparrow and Nelson’s sparrow to be one

9 species. The saltmarsh sparrow and Nelson’s sparrow were grouped together as “sharp-tailed sparrows” due to hybridization that occurs between the two species at the northern end of the saltmarsh sparrow’s range (Montagna 1942, pp. 107–120) (Figure 2) (see Hybridization section). However, prior to 1931, the two sparrows were considered distinct based on work that first recognized the differences in plumage (AOU 1899; AOU 1910; Norton 1897). In 1993, further research found that the two species differ in song, morphology, and habitat use, with only limited interbreeding (Greenlaw 1993, entire). As a result, in 1995, the American Ornithologists’ Union raised each subspecies to the species level as Ammodramus caudacutus and A. nelsoni (AOU 1995, p. 826). The AOU (recently changed to the American Ornithological Society (AOS)) is the scientific authority on avian classification in North America. The current basis for considering the saltmarsh and Nelson’s sparrows to be separate species is a combination of morphological (size, plumage), behavior (song), habitat differences, and genetics (Greenlaw 1993, p. 286; Rising and Avise 1993, p. 844; AOU 1995, p. 826; Greenlaw et al. 2018, (Systematics) unpaginated), with molecular differences also supporting this separation (Walsh et al. 2017a, p. 1250) (Fig. 2).

Figure 1. Adult saltmarsh sparrow in Rhode Island. Photo by Evan Lipton/Macaulay Library

Recent genetic analysis has found that the Ammodramus genus is polyphyletic, or an entity that is derived from more than one common evolutionary ancestor (Klicka et al. 2007, pp. 543– 545; Klicka et al. 2014, pp. 178–180). This called into question the genus designation for saltmarsh sparrow. In August 2018, after gaining acceptance from the scientific community, the AOS moved the genus of the saltmarsh sparrow to Ammospiza (Chesser et al. 2018, p. 808). As a result, the saltmarsh sparrow is now considered Ammospiza caudacuta. Based upon the preceding discussion regarding species, subspecies, and genetic delineations, and the acceptance

10 of the saltmarsh sparrow as a species by the scientific community, we also consider the saltmarsh sparrow a distinct species.

Figure 2. Saltmarsh sparrow (left) and Nelson’s sparrow (right) from Scarborough Marsh, Scarborough, Maine (photo by Kate Ruskin).

For many years the saltmarsh sparrow was thought to have two subspecies: Ammodramus caudacutus caudacutus breeding from Maine to New Jersey and Ammodramus caudacutus diversus breeding from New Jersey south to Virginia.

Recent work using molecular tools to compare the genetics of individuals across the range found weak differentiation exists between the two possible subspecies based on single nucleotide polymorphisms (SNPs) (Walsh et al. 2017a, p. 1247). This supports the view that the two subspecies populations are genetically indistinguishable, suggesting a single metapopulation of saltmarsh sparrow (Walsh et al. 2017a, p. 1248).

2.1.2 Hybridization

Hybridization with Nelson’s sparrow is thought to have occurred since the recession of the last continental ice (~11,000 years ago) (Beecher 1955, entire; Greenlaw 1993, p. 298; Rising and Avise 1993, p. 845; Walsh et al. 2016a, p. 3), has been widely accepted since at least 1942 (Montagna 1942, p. 115), and continues to be studied. Most recently, a series of detailed molecular studies have exhaustively assessed the extent and nature of the hybrid zone (Walsh et al. 2015a,b, entire; Walsh et al. 2016a,b, entire; Walsh et al. 2017b, entire). Pure species co-occur between South Thomaston, ME and Newburyport, Massachusetts (209 kilometer (km) (130 miles (mi)), with introgression (gene flow between the two species) extending 200 km (124 mi) to the north and south (Fig. 4; Walsh et al. 2017b, p. 461). Molecular and plumage evidence from resampling of sites first sampled in 1998 suggests that the area of introgression has doubled and its center has shifted 61 km (38 mi) to the south over a 15-year period (Walsh et al. 2017b, p. 461). Additionally, it was found that the relative frequency of Nelson’s sparrow genes has increased near the southern end of the hybrid zone in northern Massachusetts (Walsh et al. 11 2017b, p. 461). This shift in the hybrid zone may be a result of the range expansion of Nelson’s sparrows to the south or to differential population declines in saltmarsh and Nelson’s sparrows (Correll et al. 2016, p. 6; Walsh et al. 2017b, p. 462 (see section 2.7 Demographic Rates for more information on population declines). The lack of early generation (e.g., pure Saltmarsh x pure Nelson’s) hybrids, and support for maintained species boundaries, all lend evidence that the hybrid zone has been in existence since after the continental ice recession (~11,000 years ago), and that species recognition and reinforcement exist (Shriver et al. 2005, p. 103; Walsh et al. 2015b, p. 708; Walsh et al. 2016a, p. 3; Maxwell 2018, p. 21).

Figure 3. Map of range overlap for saltmarsh and Nelson’s sparrow, depicting the hybrid zone. Pure saltmarsh sparrow breeding range is highlighted in gray, pure Nelson’s sparrow breeding range is highlighted in black, and the hybrid zone is highlighted in yellow. Figure taken from Walsh et al. (2017).

2.2 Species Description The saltmarsh sparrow is a small songbird with a total length of 12–13 centimeters (cm) (4.7–5.1 inches (in)). Overall, the saltmarsh sparrow is olive-brown in color, with a white throat and belly, streaking on the breast, and distinctive yellow/orange coloration on the face (eyebrow and malar) (Greenlaw et al. 2018, (Appearance) unpaginated). Juveniles are similar to adults but have strong buffy tones above and below. The crown on juveniles is blackish with darker streaking and have a weakly developed medial stripe. The species appearance varies geographically and exhibits a weak clinal transition from paler browner birds in the north to

12 darker, blacker birds in the south (Greenlaw et al. 2018, (Appearance) unpaginated). For a more detailed species description of the saltmarsh sparrow, see Greenlaw et al. (2018, entire).

The saltmarsh sparrow is a tidal-marsh obligate that spends its entire life in coastal marshes along the eastern coast of the United States. The species feeds, breeds, nests, and raises its young within a narrow band of high salt marsh habitat, which is subject to tidal flooding. This species is one of several tidal-marsh endemic birds along the eastern coast of North America and is considered the most specialized of all tidal-marsh nesting bird species (Greenberg 2006, p. 3; Correll et al. 2016, p. 6). Within marshes in its breeding and wintering ranges, birds will feed in a range of microhabitats, but nesting is restricted to the highest elevation portions of the marsh, where flooding occurs only during peak spring tides. The species migrates between its breeding grounds which extend from Maine to Virginia and its wintering grounds which extend from Virginia to Florida. The species occupies the area between Massachusetts and Maryland year-round (see Figure 4).

2.3 Range and Distribution 2.3.1 Historical Range and Distribution

The historical distribution of saltmarsh sparrow is thought to have been broadly similar to the current distribution (Fig. 4). There is evidence for range contraction at the southern end of the breeding range, with the species now largely gone from the western shore of the Chesapeake Bay (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). Additionally, there is extensive documentation of habitat loss and degradation (Greenlaw et al. 2018, (Conservation and Management) unpaginated) (see Section 3.1 below) and recent population declines (Shriver et al. 2016, entire; Correll et al. 2017, p. 177). As such, saltmarsh sparrow are less numerous, though the extent of the current distribution is similar to what was reported historically (Correll et al. 2016, p. 6; Greenlaw et al. 2018, (Population Status) unpaginated) . Northward expansion of the species may have occurred in Maine, although increased sampling effort might explain this pattern (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated).

2.3.2 Current Range and Distribution

During the breeding season and immediate post-breeding molt (May-September) the entire population of saltmarsh sparrow occupies tidal marsh habitats in ten coastal states in northeastern North America (Maine, New Hampshire, Massachusetts, Rhode Island, Connecticut, New York, New Jersey, Delaware, Maryland, and Virginia). The breeding range lies from Knox County, Maine south to eastern Maryland and northeastern Virginia (Montagna 1942, entire; Hodgman et al. 2002, p. 39; Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). The core of the breeding range occurs in New Jersey, where saltmarsh sparrow abundance is highest (Fig. 5; Wiest et al. 2016, pp. 280–282).

13 Figure 4. Current Saltmarsh Sparrow range map (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated)

The range of saltmarsh sparrows is thought to be constricted at both ends by different factors. At the northern end of the range (Maine), tidal marshes are smaller and more isolated (Hodgman et al. 2002, p. 42), although proportionately they have more high marsh habitat than marshes at lower latitudes (Wiest et al. 2019, supplemental material). At the southern end of the breeding range suitable habitat is co-occupied by seaside sparrows (Ammospiza maritima). Although research has documented minimal aggression between the species (Greenlaw and Post 2006, entire) range contraction has been documented and we anticipate that as suitable nesting habitat becomes limited interspecific competition may confer an advantage to seaside sparrow (SHARP team, pers. comm.).

Following the completion of a complete body molt after the breeding season, individuals migrate south to wintering locations. The primary wintering range extends from Maryland/

14 Virginia south to Florida (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). Detailed information on occurrence and abundance across the wintering range is not available as there have been no comprehensive regional surveys; however, much of the population is thought to winter between South Carolina and northeast Florida (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). North Carolina and Virginia also support significant numbers of saltmarsh sparrows, especially during mild winters (Watts and Smith 2015, p. 390; Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). Low abundances occur on the Gulf Coast of Florida and potentially farther west (FL panhandle) (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). Winter records occur as far north as Massachusetts; however, north of Delaware, abundances become increasingly irregular and vary from year to year (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated).

2.3.3 Migration

Saltmarsh sparrows leave their breeding grounds between mid-September and late-October. Arrival to wintering grounds from North Carolina to Florida begins from September through October (Greenlaw et al. 2018, (Distribution) unpaginated), with some observations of arrivals as late as November in southeast North Carolina (Winder et al. 2012b, p. 425). Studies in North Carolina suggest that migration may occur into December (Michaelis 2009, p. 30; Winder et al. 2012b, p. 427). In South Carolina, the mean arrival date for saltmarsh sparrows was November 5, and the latest arrival was November 18 (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated).

In the spring, evidence suggests that departure dates vary by latitude. Saltmarsh sparrows wintering from North Carolina to Florida generally begin to depart the wintering grounds in March, though some individuals may remain into April and rarely into May (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). Departure dates for wintering birds in Virginia to New Jersey are not well known, but a few records suggest they leave by early to mid- June (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated).

Most individuals migrate, yet saltmarsh sparrows have been recorded in northern marshes (e.g., Cape Cod, Massachusetts) in mild winters. Migration primarily occurs at night along the coast. Preliminary results of migration studies using telemetry data suggest saltmarsh sparrows utilize stopover locations during their southern migration (i.e., depart from breeding grounds but do not fly continuously from breeding to wintering grounds) (Benvenuti et al. in prep.).

2.4 Life History This species is one of approximately 25 tidal-marsh endemic vertebrates (Greenberg 2006, p. 3) of North America and is considered the most specialized of all tidal marsh nesting bird species (Correll et al. 2016, p. 6). Within marshes in its breeding and wintering ranges, birds will feed in a range of microhabitats, but nesting is restricted to the highest elevation portions of the marsh where flooding occurs only during peak spring tides.

15 2.4.1 Life Stages

In this analysis, we consider the saltmarsh sparrow to have four life stages: egg, nestling, juvenile, and adult (Fig. 5). Saltmarsh sparrows are altricial (young are born featherless and require complete parental care) with limited mobility and eyes closed (Greenlaw et al. 2018, (Breeding) unpaginated). They spend the first 9 to 10 days developing in the nest, but are capable of fledging as early as day 8 (Gjerdrum et al. 2008a, p. 582; Greenlaw et al. 2018, (Breeding) unpaginated). Upon fledging, females continue parental care for approximately two weeks (Greenlaw et al. 2018, (Breeding) unpaginated). Collectively, the period from first egg laying to fledging takes from 23 to 27 days and chicks become independent from 15 to 20 days later (Greenlaw et al. 2018, (Breeding) unpaginated). Chicks are not capable of flight until around 22 days. The maximum recorded lifespan is 10 years for males and 6 years for females (Greenlaw et al. 2018, (Demography and Populations) unpaginated).

Figure 5. Annual life cycle of the saltmarsh sparrow. Darker shading per life stage indicates the core months during which each specific life stage occurs.

2.4.2 Diet

Adult diets differ by season. Insects (44 percent), amphipods (24 percent), and spiders (21 percent) make up the largest proportion of breeding season diet (Greenlaw et al. 2018, (Diet and Foraging) unpaginated). Nestlings strictly consume insects (Post and Greenlaw 2006, p. 768). Diet during the post-breeding period (September-October) includes seeds (30 percent), predominantly Spartina alterniflora (smooth cordgrass)(Greenlaw et al. 2018, (Diet and Foraging) unpaginated). Other elements of diet during the post-breeding season are similar to that of the breeding season. Adults and juvenile food resources during the wintering season are largely unknown. Saltmarsh sparrows are thought to consume seeds at high rates during the wintering season (Greenlaw et al. 2018, (Diet and Foraging) unpaginated).

16 2.4.3 Site Fidelity

Male and female saltmarsh sparrows are known to exhibit a high degree of site fidelity to both their breeding (DiQuinzio et al. 2001, p. 892) and wintering (Winder et al. 2012b, p. 427) grounds, both within and across breeding seasons (DiQuinzio et al. 2001, p. 892; Benvenuti et al. 2018a, p. 895). In Rhode Island, adult saltmarsh sparrows returned to the previous year's breeding site approximately 35 percent of the time (DiQuinzio et al. 2001, p. 892). In Maine and New Hampshire, 85 percent of females returned to nest within their home range from the previous year, regardless of the prior nest’s fate (Benvenuti et al. 2018a, p. 11). On the wintering grounds, 99 percent of banding recaptures were at the site of original banding, with a return rate of 10 percent across years (Winder et al. 2012b, p. 427).

A proportion of juvenile (6–30 percent) saltmarsh sparrows return to their natal marsh the following year to breed as adults (DiQuinzio et al. 2001, p. 895).

2.4.4 Breeding

Both sexes of saltmarsh sparrow begin to breed the summer after they hatch. The breeding range occurs from the northeast and mid-Atlantic regions of the United States including Maine, New Hampshire, Massachusetts, Connecticut, Rhode Island, New York, New Jersey, Delaware, Maryland, and Virginia (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). The breeding season of saltmarsh sparrows varies slightly by latitude, with nesting occurring from late-May through early-September (Ruskin et al. 2017b, p. 297).

Saltmarsh sparrows exhibit a breeding system that has evolved over several million years to utilize tidal marsh habitat (Greenlaw 1993, pp. 298–299). Their mating system is described as polygynous scramble competition, where males search for and attempt to mate with multiple receptive females (Hill et al. 2010, p. 303; Walsh et al. 2018, p. 6). Individuals are non territorial (Woolfenden 1956, p. 34; Shriver et al. 2010, p. 343) and do not maintain pair bonds (Greenlaw and Post 2012, pp. 256–259). The species is known to have a higher rate of female multiple-mating than nearly any other bird species (Hill et al. 2010, p. 303; Walsh et al. 2018, p. 6).

Female saltmarsh sparrows are responsible for all aspects of breeding success following mating, including nest construction, egg laying, incubation, and nestling and post-fledging care (Woolfenden 1956, p. 35; Post and Greenlaw 1982, p. 104). Early nests are often destroyed by peak spring tides, which results in synchronous renesting (DeRagon 1988, p. 58-61; Shriver et al. 2007, pp. 123–124) (Fig. 6), that is generally initiated within several days after nest loss. Although females typically only have one successful reproductive attempt per breeding season, if a nest fails, then a female will attempt to renest up to three times per breeding season (Shriver et al. 2007, pp. 123–124; Greenlaw et al. 2018, (Breeding) unpaginated).

To date, there is no lifetime reproductive success information on saltmarsh sparrows (Greenlaw et al. 2018, (Breeding) unpaginated). Historical data estimate annual reproductive success to range from 2.65 to 5.25 fledged young per female (Post and Greenlaw 1982, p. 104). Current, rangewide, estimates of annual reproductive success vary by site, with a mean of 1.27 (±0.45) fledged young per female each season (Ruskin et al. 2017b, p. 297). 17 Figure 6. Typical nesting cycle of saltmarsh sparrow in relation to the tidal cycle.

Nest Placement and Structure

Females construct nests in the vegetation on the high marsh platform, which by definition, floods during monthly, peak spring tides rather than daily, high tides. Saltmarsh sparrow nest construction is variable, but usually formed as a modified cup shape or bowl situated just a few inches above the marsh surface, typically in Spartina patens, Juncus gerardii, or S. alterniflora grasses (DeRagon 1988, p. 49-58; Gjerdrum et al. 2005, p. 858; Shriver et al. 2007, pp. 555– 556; Ruskin et al. 2015, p. 1642; Benvenuti et al. 2018b, p. 6; Greenlaw et al. 2018, (Breeding) unpaginated). Nests are supported by adjacent grass stems and anchored by grass strands. Nests are often placed under thickets of thatch (Gjerdrum et al. 2005, p. 858), and/or have woven grasses above the nest to form a canopy (Greenlaw et al. 2018, (Breeding) unpaginated; Benvenuti et al. 2018b, p. 2) (Fig. 7).

18 Figure 7. Simplified diagram of saltmarsh sparrow nest structural components (Benvenuti et al. 2018b, p. 2)

Saltmarsh sparrows construct their nests in areas of higher elevation within the marsh (Gjerdrum et al. 2005, p. 858; Shriver et al. 2007, p. 555; Benvenuti et al. 2018b, pp. 6–7), and in vegetation that is, on average, taller and more dense (Gjerdrum et al. 2005, p. 858).

Nests are typically randomly distributed across a marsh with respect to other nests (Bayard and Elphick 2010b, p. 491), but clustering has been observed due to similarity in nest-placement choices made by individual females (Bayard and Elphick 2010b, pp. 489, 491). Average distance between closest neighboring nests is 43.70 ± 3.01 meter (m) (143.3 ± 9.87 feet (ft) (Post and Greenlaw 1982, p. 103). Saltmarsh sparrows nest only 37–47 m (121–154 ft) apart from former nest sites in successive years (Post and Greenlaw 1982, p. 106; Benvenuti et al. 2018b, pp. 8–9), thus exhibiting site and nest placement fidelity.

Multiple studies have examined how nest structure impacts nest success, egg retention, and microclimate. Nest height, canopy cover, and marsh surface elevation are characteristics that have been found to contribute to nest success (Humphreys et al. 2007, p. 155–156, Benvenuti et al. 2018b, p. 6). Successful nests as well as depredated nests have been found to be constructed higher in the marsh vegetation, and in areas of higher surface elevation than those that fail due to flooding (Benvenuti et al. 2018b, p. 6). The presence of a nest canopy appears to aid in egg retention during periods of flooding, and provides concealment from predators (Humphreys et al. 2007, pp. 156–157), such that nests that are successful have a greater proportion of canopy cover than those that fail due to flooding or depredation (Benvenuti et al. 2018b, p. 6).

Egg laying

First nest initiation varies by latitude. In the southern portion of the breeding range, females begin nesting earlier than those nesting in the northern portion (Ruskin et al. 2017b, p. 297). Females typically lay a clutch of 4 eggs. Across 23 sites from Maine to New Jersey, the mean clutch size was estimated as 3.65, with variation among sites and slightly larger clutches at

19 higher latitudes (Ruskin et al. 2017b, p. 297). Females lay one egg per day until the clutch is complete (Ruskin et al. 2017b, p. 297; Greenlaw et al. 2018, (Breeding) unpaginated).

Incubation

Incubation begins following the laying of the last egg of a clutch (Greenlaw et al. 2018, (Breeding) unpaginated). The incubation period lasts 12.09 +/-0.98 days (Ruskin et al. 2017b, p. 297) with females completing all of the incubation (Post and Greenlaw 1982, p. 106). Females complete 98 percent of their trips away from the nest during the day, indicating that females incubate nearly continuously throughout the night. Females spend 33 percent of the daylight hours away from the nest, during trips that last 11.8 minutes on average (Gjerdrum et al. 2008a, p. 580). Females modify trips away from the nest in relation to ambient temperature. For example, they take shorter, more frequent trips away in cooler temperatures, but take longer, less frequent trips in warmer conditions (Gjerdrum et al. 2008a, p. 581).

Nestling stage

Hatching can occur synchronously or asynchronously, with all eggs typically hatching within 24–36 hours of each other (Greenlaw et al. 2018, (Breeding) unpaginated). Saltmarsh sparrow chicks are born featherless and blind with limited mobility (Greenlaw et al. 2018, (Breeding) unpaginated). They spend 9–10 days in the nest growing feathers, developing eyesight, and attaining mobility. Chicks typically fledge (leave the nest but still require parental care) after 10 or 11 days (Greenlaw et al. 2018, (Breeding) unpaginated), but are capable of fledging as early as 8 days if disturbed by flooding or a predator (Greenlaw et al. 2018, (Breeding) unpaginated) (Fig. 9). Lower nest temperature data indicate that nestlings often leave the nest after sunset near the high tide (Gjerdrum et al. 2008a, p. 582). Nest temperatures increased again between 1.5 and 8.5 hours after fledging, indicating that either nestlings or the adult female returned briefly to the nest (Gjerdrum et al. 2008a, p. 582).

Figure 8. (A) Saltmarsh sparrow nest with eggs (by Alison Kocek); (B) Day old saltmarsh sparrow chicks (by Bri Benvenuti); (C) Near fledging saltmarsh sparrow chicks (by Emma Shelly).

20 Nest failure and success

Nest Flooding - Across the range, flooding is the primary cause of nest failure, which is defined as the loss of eggs or nestling mortality. High tides often cause nest failure if they exceed the nest height (Ruskin et al. 2017a, p. 910). Tidal flooding of nests typically occurs during peak spring tides, and lasts, on average, 90 minutes during an inundation period (Gjerdrum et al. 2008a, p. 581; Bayard and Elphick 2011, p. 401). Nest failure due to flooding is also known to occur during severe storms/precipitation/wind events when tide heights are elevated (Bayard and Elphick 2011, p. 402; Ruskin et al. 2017a, p. 910). While flooding often causes complete nest failure, there are instances of partial failure due to the loss of eggs or nestlings (Humphreys et al. 2007, p. 155; Gjerdrum et al. 2008a, p. 582; Bayard and Elphick 2011, p. 397). However, in a study in Connecticut, only 15 percent of nests did not experience any flooding, and only 18 percent were successful overall (Bayard and Elphick 2011, p. 401).

Eggs are able to withstand flooding events if they are able to remain in the nest, thus incubation and egg development can resume when water levels recede (Humphreys et al. 2007; p. 156; Gjerdrum et al. 2008a, p. 582). Lower hatching success following periods of nest inundation has been observed, however (Walsh et al. 2016a, p. 900). Older eggs, which are closer to hatching, are likely to be more susceptible to failure due to the effects of cooling, which slows embryo development (Olson et al. 2006, p. 934; Gjerdrum et al. 2008a, p. 582).

In addition to eggs, nestlings are able to survive nest flooding events if they are large enough to tolerate cold temperatures due to nest inundation, and either keep their heads above water or climb high enough out of the nest to avoid drowning (Gjerdrum et al. 2008a, p. 582). Due to the mobility and body mass needed to survive flooding events, chicks younger than 5 days are typically unable to survive nest flooding (Fig. 9).

Figure 9. Examples of nest flooding (Left Jenna Mielcarek; Right Bri Benvenuti)

21 Nest Predation - A systematic study of nest predators has not been undertaken, but researchers across the breeding range believe the following species are likely predators: garter snake (Thamnophis sirtalis), mallard (Anas platyrhynchos), wading birds, northern harrier (Circus cyaneus), gulls, corvids (Corvus spp.), blackbirds, coyote (Canis latrans), raccoon (Procyon lotor), domestic cat (Felis domesticus), and rodents (Greenlaw et al. 2018, (Demography and Populations) unpaginated). The southern portion of the breeding range (New York, New Jersey) incurs higher predation rates than that of the more northern populations (Ruskin et al. 2017a, p. 909).

Nest Success - Nest success is best predicted by the timing of nest initiation. Nests that are initiated within three days of a peak spring tide are more likely to be successful by avoiding tidal flooding (Gjerdrum et al. 2005, p. 856; Shriver et al. 2007, p. 555). Timing of nest initiation in relation to peak spring tides strongly indicates that tide height and flooding frequency are important drivers of nest fate (Bayard and Elphick 2011, p. 393).

There is strong evidence of nest site selection preferences for characteristics that confer resistance to flooding (Gjerdrum et al. 2005, p. 859; Humphreys et al. 2007, p. 156; Shriver et al. 2007, p. 556; Ruskin et al. 2015, p. 1643). Nest sites typically have more Spartina patens, and are located in areas with deeper thatch and at higher elevation than random locations within a marsh (DeRagon 1988, p. 53-58; Gjerdrum et al. 2005, p. 854; Shriver et al. 2007, p. 555). Among nests, those that are successful or depredated have been found to be constructed higher in the vegetation, have increased canopy cover, and are in higher elevation areas of the marsh compared to nests that failed due to flooding (Benvenuti et al. 2018b, p. 6).

The minimum number of nesting attempts per season varies regionally from 1.06–1.36 and increases with latitude, showing that females are more likely to renest later into the season at northern sites (Ruskin et al. 2017b, p. 297). Females that are successful will regularly attempt to produce a second clutch, although information is lacking on females’ ability to produce two successful broods in a given year.

Post-fledging Period - Females with newly fledged offspring spend time in 0.14 to 1.06 ha (0.35 ac to 2.6 ac) surrounding the nest, typically moving no further than 100 m (328 ft) from the nest site (Hill 2008, p. 18). Furthermore, females bring food items to multiple different locations during the post-fledging period, suggesting that fledglings do not stay together after fledging (Hill 2008, p. 12).

There is no observed relationship between adult female distance from nest and time since fledging (Hill 2008, p. 19). Vegetation types associated with post-fledging movements include Spartina patens, S. alterniflora (tall and short), Distichilis spicata, and Juncus gerardii (Hill 2008, p. 39). A majority of saltmarsh sparrow females (75 percent) spent time closer to the marsh edge than expected while tending to fledglings (Hill 2008, p. 40); however, they did not move to different marshes nor did they cross elevated barriers such as roads, railroad tracks, or elevated berms within a marsh (Hill 2008, p. 22).

22 2.4.5 Adult and Offspring Sex Ratios

Adult sex ratios have been found to be uneven. Multiple mark-recapture studies report a male-biased adult sex ratio, with two or more males captured for every female in most cases (Gjerdrum et al. 2008a, p. 612; Greenlaw et al. 2018, (Behavior) unpaginated). In contrast, Benvenuti et al. (2018a, p. 13) found a more even adult sex ratio than that reported in other studies (1.05 to 2.01 males per female) (Post and Greenlaw 1982, p. 106). Present evidence shows no sex-based differences in adult annual survival (Borowske 2015, pp. 103–104; Field et al. 2018, p. 976), suggesting that the sex ratio differences cannot be attributed to sex-biased mortality. Therefore, there is uncertainty surrounding the causes of uneven adult sex ratios.

A study of four marshes across Maine, New Hampshire, and Massachusetts (n=370 nests, n= 990 offspring, n=210 females; 2011–2015) investigated site and population level sex ratios and tested the influence of site quality, year, ordinal date, nest initiation relative to peak high tides, and female condition on offspring sex ratio (Benvenuti et al. 2018a, p. 10). Across years and sites, they found an even offspring sex ratio (1.03:1), with inter-annual year and site variations (Benvenuti et al. 2018a, pp. 8–10). The authors completed the same analysis on nestlings that survived to fledgling and found the sex ratio remained even (1.10:1), suggesting there is no differential survival between male and female nestlings (Benvenuti et al. 2018a, pp. 9–10). Benvenuti et al. (2018a, p. 13) also found a pattern between offspring sex ratio of a current year and the adult sex ratio of the previous year, such that in years when the adult sex ratio was male- biased, fewer male nestlings were produced and vice versa. In Connecticut, the offspring sex ratio was found to be significantly skewed toward males (1.45:1) over a two year period (Hill et al. 2013, p. 414). Due to the smaller sample size and shorter time period of the Connecticut study, we give more weight to the first study and conclude that offspring sex ratios are likely to be even, and are not modified due to environmental factors or female condition, but may produce more of the rarer sex during a given year (Benvenuti et al. 2018a, pp. 11–13).

2.5 Individual Resource Needs: Habitat As discussed above, saltmarsh sparrows are habitat specialists dependent on tidal salt marsh along the east coast of the United States. Tidal salt marshes are dynamic systems that experience daily tidal inundations in accordance with the lunar schedule (Tiner 2013, p. 6), with the amount of the marsh flooded varying over the monthly cycle. The dominant vegetation in tidal marshes are grasses of varying salt tolerance levels. Typically marshes are divided into the low marsh and high marsh. The low marsh is inundated with salt water on a daily basis whereas the high marsh is generally inundated on a monthly basis. Typical low marsh plants include Spartina alterniflora (short and tall form) in the breeding range, and S. alterniflora and S. cynosuroides (big cordgrass) in the wintering range. Typical high marsh plants include S. patens, Juncus gerardii, and Distichilis spicata in the breeding range and J. roemerianus and D. spicata in the wintering range. In their breeding range, maritime shrublands are common at the upland border (Fig. 10).

23 Figure 10. Typical saltmarsh sparrow breeding habitat (top left: John H. Chafee National Wildlife Refuge (NWR), RI (Photo: Suzanne Paton); top right: Rachel Carson NWR, ME (Photo: Bri Benvenuti), and wintering habitat (bottom: Kiawah Island, NC (Photo: Adam Smith).

2.5.1 Breeding Range Habitat Characteristics and Use

Saltmarsh sparrows utilize each area of the marsh for different purposes, with nesting occurring on the high marsh platform (Greenlaw et al. 2018, (Distribution, Migration and Habitat) unpaginated). Foraging individuals and fledglings spend time in low marsh channel edges and salt pannes (Hill 2008, p. 39). Adults and juveniles will often take cover in terrestrially-bordered shrubs, or patches of Phragmites, and this may be especially true during extreme weather and high tide events (Greenlaw, pers. comm.).

Due to the lack of pair-bonds and territorial boundaries, home ranges of both males and females commonly overlap (Shriver et al. 2010, p. 342). The average home range of adult saltmarsh sparrows is 53 ha (131 ac) for males and 28 ha (69 ac) for females, with a single “core area” of 10 ha (25 ac) and 5 ha (12 ac), respectively (Shriver et al. 2010, p. 342). Male home ranges have greater cover of Spartina alterniflora less S. patens than random locations; female home ranges had greater cover of Juncus gerardii than did random locations (Post and Greenlaw 1982, p. 342). Home ranges for both sexes had more vegetated salt panne cover (water-retaining depressions) than did random locations (Shriver et al. 2010, p. 343).

24 The density of nesting saltmarsh sparrows is positively correlated with marsh area (Shriver and Vickery 2001, p. 8; Benoit and Askins 2002, p. 317; Meiman et al. 2012, p. 733) and more connected marshes (Meiman et al. 2012, p. 733; Shriver et al. 2004, p. 549). Numerous studies have further documented the association between the percent cover of high marsh vegetation and suitable nesting habitat (Meiman and Elphick 2012, p. 857), saltmarsh sparrow home ranges (Shriver et al. 2010, p. 342), and nesting abundance (DiQuinzio et al. 2002, pp. 183–183; Gjerdrum et al. 2005, pp. 854–855; Shriver et al. 2007, p. 556).

2.5.2 Wintering Range Habitat Characteristics and Use

Saltmarsh sparrow habitat use in the wintering range is not as well described in the literature. Generally, saltmarsh sparrows are found in Spartina marshes. Studies along the Gulf Coast of Florida suggest they prefer patches of S. alterniflora and appear to avoid large monocultures of Juncus roemerianus (black needle rush) (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). Declines in saltmarsh sparrow prevalence at Shell Key, Florida have been attributed to the intrusion of salt marshes by mangroves (Cox et al. 2012, p. 8).

Saltmarsh sparrows will roost in tall vegetation (e.g., tall cordgrass, black needlerush) during high tides, at night, or when disturbed (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). During high spring tides they will leave the wetter, low marsh areas, and instead use vegetation along the landward edges of the marsh in drier grass zones, spoil banks, dune scrub, hummocks, or berms (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). Saltmarsh sparrows have also been observed taking refuge in stands of Myrica cerifera (southern bayberry) and Baccharis spp. (aster) (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). There is anecdotal evidence to suggest that saltmarsh sparrows may prefer to roost in interior portions of the salt marsh rather than along upland edges (Borowske 2015, p. 157).

Saltmarsh sparrows generally forage in Spartina alterniflora during both high and low tides and sometimes on the ground in Juncus roemerianus and along the interface of J. roemerianus and S. alterniflora (Greenlaw et al. 2018, (Diet and Foraging) unpaginated). Foraging locations are generally within 300 m (984 ft) of roost sites (Greenlaw et al. 2018, (Diet and Foraging) unpaginated).

25 2.6 Summary of Needs for Individuals Table 1. Summary of individual resource needs (habitat) for saltmarsh sparrow to complete each life stage.

Life Stage Resource Needs (Habitat) Select References

Egg ● High marsh platform Greenlaw et al. 2018 ● Nest site above reach of ≅80% of high tides Humphreys et al. 2007 ● Dense vegetation cover composed of different Gjerdrum et al. 2005 species, structure, and thatch Shriver et al. 2007 ● Nest cup, with canopy Ruskin et al. 2017b

Nestling ● High marsh platform Greenlaw et al. 2018 ● Nest site above reach of ≅80% of high tides Humphreys et al. 2007 ● Dense vegetation cover composed of different Gjerdrum et al. 2005 species, structure, and thatch Shriver et al. 2007 ● Nest cup, with canopy Ruskin et al. 2017b ● Food (insects) Post and Greenlaw 2006

Juvenile/ ● High and low marsh habitat with limited Greenlaw et al. 2018 Adult fragmentation and alterations Ruskin et al. 2017b ● Nest site above reach of ≅80% of high tides Hill et al. 2008 ● Dense vegetation cover composed of different Benoit and Askins 2002 species, structure, and thatch Marshall 2017 ● Taller vegetation for storm refugia SHARP team 2017, pers. ● Food (insects, amphipods, spiders, seeds) comm.

26 FACTORS INFLUENCING THE VIABILITY OF THE SPECIES

In this section we discuss the threats that may influence the current condition of the saltmarsh sparrow as well as those conservation measures in place which may ameliorate those threats. We use the term “threat” to refer in general to actions or conditions that may, or are reasonably likely to, negatively affect individuals of a species. The term “threat” includes actions or conditions that have a direct impact on individuals (direct impacts), as well as those that affect individuals through alteration of their habitat or required resources (stressors). The term “threat” may encompass—either together or separately—the source of the action or condition, or the action or condition itself. We assessed the following threats (and associated stressors where applicable) acting on the saltmarsh sparrow: (1) habitat loss, fragmentation, and degradation; (2) the effects of climate change; (3) hybridization; (4) predation; (5) contaminants; and, (6) other factors influencing the species such as disease and altered food webs.

As a result of assessing the magnitude and degree of impact of these threats, we identified flooding from sea level rise (effects of climate change) as the threat most likely influencing the current condition of the saltmarsh sparrow and potentially driving population- or species-level impacts on the species. The increased frequency and duration of flooding directly impacting nest success is also the primary driver of the degradation and loss of suitable nesting habitat. While predation is responsible for a significant proportion of nest losses, and contaminants may be having sublethal impacts to the species, we considered these to be impacting the population on a localized- or individual-level, and not a primary driver of the population trend overall for the species. Hybridization with Nelson’s sparrow at the northern extent of the range is considered an environmental baseline natural to the areas where the species’ overlap and part of their life history that does not appear to be influencing the species negatively or positively at this time.

3.1 Habitat Loss, Fragmentation, and Degradation Salt marshes have evolved in the dynamic interface between terrestrial and marine ecosystems and historically have been capable of maintaining themselves through natural ecological processes of sediment capture and organic matter accumulation both above and below ground (Cahoon et al. 2009, pp. 59–62) (Fig. 11). Variations in sediment supply, hydrologic functioning, geologic activity, and nutrient supply can support or undermine the natural ecological processes that maintain salt marsh habitat and influence the ability of salt marshes to respond to natural environmental or anthropogenic changes to sea-level (Kirwan et al. 2010, p. 2; Burdick and Roman 2012, p. 374; Deegan et al. 2012, p. 1; Weston 2014, entire; Wigand et al. 2014, p. 645; Haaf et al. 2015, pp. 5–8; Roman 2017, entire; Ganju et al. 2016, entire; Mitchell et al. 2017, p. 9). Alterations to marshes can also influence their ability to respond to any individual stressor or combination of stressors (Smith et al. 2017, pp. 33–34).

In this section, we summarize the literature with regard to different stressors impacting tidal marsh habitats throughout the range of saltmarsh sparrow. We have not attempted to represent the scope and breadth of the literature attributed to tidal marsh ecological processes but have focused on the stressors that have been discussed in recent publications as impacting marsh vulnerability to sea-level rise. Comprehensive reviews of tidal marsh processes and threats are

27 also provided in Tiner (2013, entire), Roman and Burdick (2012, entire) and Silliman et al. (2009, entire) among others.

Figure 11. Graphic illustration of tidal wetland processes and factors that may influence natural salt marsh ecosystem dynamics (from Cahoon et al. 2009)

3.1.1 Urbanization and Marsh Habitat Loss

Coastal areas throughout the United States (U.S.) support a disproportionate number of people per capita and as such have experienced increased development pressure over time. According to the National Oceanic and Atmospheric Administration (NOAA) (2018a, entire), coastal areas representing 10 percent of the U.S. land mass supported 39 percent of the U.S. population in 2010. Additionally, among the 31 States with coastal communities, 8 of the top 15 most densely populated States (persons/mi2) are within the breeding range of saltmarsh sparrow, with only New Hampshire and Maine ranking lower. Between 1780 and 1980, about 50 percent of all salt marsh habitats in the U.S. were lost or converted by filling and dredging activities (Dahl 1990, p. 1). A 2005 assessment of the New England States estimated 37 percent of salt marsh has been lost since 1851, with the highest percentages lost in Rhode Island (53 percent) and Massachusetts (41 percent); and the lowest in New Hampshire and Maine at 18 percent and less than 1 percent, respectively (Bromberg and Bertness 2005, p. 828).

Much of these salt marsh losses represent direct conversion of wetlands to upland habitats prior to the protections afforded under the Clean Water Act (CWA) of 1972. Marshes,

28 particularly in the mid-Atlantic States, were ditched and diked to create suitable farmland or grazing habitat (Roman et al. 2000, p. 749-750), to provide non-tidal shallow water or mudflat habitats for waterfowl, or in later years to reduce mosquito breeding. In addition, shorelines were armored (rocked or otherwise protected for erosion control purposes). Over time, development continued to expand on land adjacent to marshes resulting in the loss of vegetated buffers and introduction of associated stressors (i.e., nutrients, contaminants, invasive species, and predators). Although the degree of habitat loss and alteration varies throughout the range of the saltmarsh sparrow, the severity and extent of urbanization is correlated with marsh loss as well as marsh health and resilience (Roman et al. 2000, entire; Bertness et al. 2009b, p. 140-141). Although marsh loss has slowed since the 1970s, an evaluation of marsh loss between 1970 and 2009 in the Chesapeake Bay watershed (York River estuary) showed a total loss of 2.7 percent of the marsh (Mitchell et al. 2017, p. 6). The most important predictor of marsh loss was surrounding development, including a high correlation with length of riprap, bulkheads, and shoreline hardening. Low elevation sites with extensive stabilization lost between 13 and 31 percent of the marsh attributed to the inability of marshes to migrate landward and a reduction in sediment availability. Sites without shoreline hardening saw the most gains in marsh extent, (Mitchell et al. 2017, pp. 6–9) with some areas maintaining or expanding their historic extent through landward migration (Schieder et al. 2018, pp. 944–945).

An evaluation of 36 marshes in Rhode Island calculated 104 ha (257 ac) of marsh loss between 1972 and 2011 (3.3 percent) with one marsh expanding in extent and the rest experiencing losses up to as much as 37 percent (Watson et al. 2017b, p. 665). Massachusetts lost 30 percent of its salt marsh habitats by 1995, with the marshes on Cape Cod losing slightly less at 21 percent (Carlisle et al. 2005, p. 13). In the nearby Westport River, an average of 48 percent of the marsh islands were lost between 1934 and 2016, with rates accelerating in the past 15 years (Costa and Weiner 2017, p. 16). In Jamaica Bay, New York, 38 percent of the vegetated marsh area was lost between 1974 and 1999 (Hartig et al. 2002, p. 79).

3.1.2 High Marsh Degradation, Conversion, and Fragmentation

As stated above, salt marsh habitat is dynamic in nature and its position on the landscape is a function of its interaction with both the ocean and upland components and processes. If these interactions are disrupted (e.g., by changes in hydrology), salt marsh habitat may be restricted, degraded, or otherwise converted to habitat types not used by or less suitable for saltmarsh sparrows. The amount of high marsh within the larger marsh landscape is an important predictor of saltmarsh sparrow nesting habitat suitability. In this section, we discuss in more detail the current information published with regard to the direct loss, fragmentation, degradation, and conversion of high marsh to other habitat types, including primarily low marsh habitat which does not provide suitable nesting habitat. As nesting habitat is the most limiting component of the saltmarsh sparrow habitat needs, an assessment of the current condition of high marsh is important to our assessment of habitat condition overall.

The degradation and loss of vegetated marsh has been characterized generally in four categories: (1) shoreline erosion; (2) loss at the bay head region of back barrier lagoons and

29 estuaries; (3) widening and headward erosion of tidal channels; and (4) development and expansion of interior depressions or ponds (Kearney et al. 1988, pp. 217–219; Hartig et al. 2002, pp. 85–87; Kearney et al. 2002, p. 175; Smith 2009, p. 189; Watson et al. 2017b, p. 669). Both high marsh and low marsh are affected by these losses, but numerous studies have documented a net transition to more flood tolerant low marsh vegetation or a complete loss of vegetation. Specific examples include the conversion of high marsh to low marsh (Donnelly and Bertness 2001, pp. 14222–14223 [Rhode Island]; Raposa et al. 2017b, pp. 643–644 [Rhode Island]), a shoreward migration of S. alterniflora (Warren and Niering 1993, pp. 99–102 [Connecticut, New York]; Field et al. 2016b, 365–367 [Connecticut]; Hartig et al. 2002, 85–87 [New York]; Bertness et al. 2002, pp. 1395–1398 [Rhode Island]; Smith 2009, p. 189; Smith 2015, pp. 132–134 [Massachusetts]; Mitchell et al. 2017, pp. 5–12 [Maryland, Virginia]), mudflat expansion (Tiner et al. 2006, entire [Connecticut]) and standing water on the marsh at low tide (Basso et al. 2015, pp. 16–17 [Connecticut, New York]).

Multiple factors contribute to degradation of tidal salt marsh habitat and the loss of high marsh habitat in particular. These factors include sea level rise, invasive species, nutrient enrichment, altered hydrology, suspended sediment concentrations, shoreline protective measures, and human disturbance and management activities, which will each be discussed below. Following is a summary of the data available on the resulting loss of high marsh habitat that has been documented in recent years.

At Cape Cod National Seashore (CCNS), Massachusetts there has been a documented shift in the high marsh to low marsh boundary by 100 m (328 ft) upslope between 1984 and 2000 in the ‘Gut’ at Wellfleet, and a loss of 46 percent of the high marsh between 1947 and 2005. Another site, Middle Meadow, lost 37 percent of the high marsh during the same time period. Loss patterns vary between sites, suggesting that it cannot be attributed to a single climatic event or sudden wetland dieback, although crab herbivory on degraded marsh edges seems to compound the losses (Smith 2009, pp. 197, 200). A follow up study comparing aerial imagery on CCNS from 1984 to 2013 at six marshes (approximately 1,700 ac (688 ha)) found that 190 ac (77 ha) (70 percent) of high marsh was lost while low marsh increased overall by 131 ac (53 ha) (18 percent) during the same time period. Marsh transgression into upland sites was documented, as was burial of high marsh due to sand over-wash from a dune. (Smith 2015, pp. 132–133).

An assessment of marshes on Long Island, New York estimated overall marsh loss from 11 to 23 percent between 1974 and either 2005 or 2008 (2,750 ac (1,113 ha) overall). During the same time period high marsh declined by an average of 27 percent across sites (range 11.3–71.8 percent), representing a loss of over 2,000 ac (809 ha) of high marsh on Long Island over a 31 to 34 year period (Cameron Engineering & Associates 2015, pp. 23, 27).

On two Rhode Island marshes, the occurrence of Spartina patens (i.e., high marsh) declined by 16 to 40 percent between 1995 and 1999 with S. alterniflora (i.e., low marsh) becoming 5 times more abundant and migrating landward into high marsh areas (Donnelly and Bertness 2001, p. 1). There was also a direct correlation between marsh elevation and the rate of loss such that as these marshes become lower in relation to sea-level, the rate of loss accelerated (Watson et al. 2017b, p. 670). Statistically significant decreases in S. patens cover and increases in S. alterniflora cover were documented between 2000 and 2013 (Raposa et al. 2017b, pp. 644–645).

30 In Virginia, marshes also shifted from diverse marsh habitats to a monotypic marsh dominated by S. alterniflora (Mitchell 2016, unpublished presentation).

The loss and conversion of high marsh habitat within the overall marsh effectively fragments the availability of suitable nesting habitat into smaller and potentially disjunct patches.

3.1.2.1 Invasive Species

Invasive species cause direct and indirect negative effects to saltmarsh sparrows and their habitat. Invasion of salt marshes by Phragmites australis (common reed) has the potential to negatively impact saltmarsh sparrows through a displacement of native high marsh nesting habitat (Roman et al. 1984, p. 146; Chambers et al. 2003, p. 399; Greenberg et al. 2006, p. 682). Invasion of the marsh by Phragmites may be aided by human activities such as creating tidal restrictions that alter tidal flow, or altering salinity or nutrient flow to marshes (Bertness et al. 2002, p. 1397; Roman et al. 1984, pp. 144–145). Marshes used by saltmarsh sparrows have a lower percent cover of Phragmites than those not used by the species (Shriver and Vickery 2001, p. 10). Nesting saltmarsh sparrows used Phragmites less frequently than salt meadow habitat (DiQuinzio et al. 2002, p. 184) or were absent from tall or moderate heights of Phragmites at Galilee Marsh, Rhode Island (Myshrall et al. 2000, entire).

In addition to causing a direct loss of habitat, the spread of Phragmites may also have indirect effects on saltmarsh sparrows, in making the habitat more suitable for generalist species over salt marsh specialists (Benoit and Askins, 1999, p. 198). Species that have been observed destroying saltmarsh sparrow nests, including marsh wren and red-winged blackbirds, utilize Phragmites habitat for their own nests (Greenberg et al. 2006, p. 682).

There is evidence that Phragmites invasion can result in higher sedimentation rates in Atlantic Coast marshes due to greater below ground production (Rooth and Stevenson 2000, pp. 180–181). This has the potential to provide appropriate elevation for high marsh habitat to outcompete Phragmites and persist under rapid sea-level rise, but any potential for long term benefits are currently under evaluation and have not been confirmed (Raposa, 2018 [pers. comm.]) Currently the loss of habitat to Phragmites invasion negatively affects saltmarsh sparrow.

Additional invasive plant species, including Lepidium latifolium (perennial pepperweed) and Lythrum salicaria (purple loosestrife) have also been introduced to salt marsh areas in the saltmarsh sparrow’s range. Although the effects of these invasive plants have not been specifically evaluated relative to saltmarsh sparrows, they have been documented as: invading tidal marshes and displacing native vegetation, altering the decomposition rates and nutrient cycling of wetlands, reducing wetland plant diversity, impacting pollination and seed output of the native plants, altering the structure of marshes, causing shifts in food availability for species that eat insects, and reducing habitat suitability for specialized wetland bird species such as black tern, least bittern, pied-billed grebe, and marsh wren (Blossey et al. 2001, pp. 1787–1807; Zedler and Kercher 2004, pp. 431–452; Tobias et. al. 2016, pp. 411–418).

31 3.1.2.2 Nutrient Enrichment

The amount of development adjacent to tidal marsh habitats is positively correlated with increases in nitrogen and phosphorus discharged into waterways from sewage treatment plants or leached into groundwater from individual cesspool and septic systems (Tasdighi et al. 2017, pp. 116–117). Farm waste and / or the use of fertilizer on crop lands, residential lawns, and golf courses in coastal watersheds can also result in elevated nitrogen levels in ground or surface waters (Bertness et al. 2002, p. 1397). The sources and amounts vary from Maine to Virginia, and also include atmospheric deposition, with more southern marshes generally more impaired than those to the north (Roman et al. 2000; 759-760).

Extensive research of marsh plants exposed to excess nutrients has quantified changes to marsh community composition and diversity, as well as individual plant growth. Generally plants respond with an increase in above ground growth, a decrease in below ground biomass, and an increase in decomposition rates of organic material driven by microbial activity (Deegan et al. 2007, p. S43; Selman et al. 2008, entire; Wigand et al. 2009, p. 960; Tiner 2013, p. 262). These changes in turn result in weakened soil strength and decreased integrity of marsh edges and banks. The combined effects are susceptibility to erosion and creek widening, as well as subsidence of the entire marsh platform (Deegan et al. 2012, pp. 388–389). These effects have been associated with loss of low marsh habitat at several sites occupied by the saltmarsh sparrow, including Jamaica Bay, New York (Hartig et al. 2002, pp. 82–87), at six nutrient enriched estuaries in Long Island Sound (Deegan et al. 2012, Figure 3b. p. 391), and at islands in the Westport River, Massachusetts (Buzzards Bay Coalition 2017, entire). Marshes that are stressed by elevated levels of nutrients and concomitant elevation deficits will be less resilient to the impacts of sea-level rise and are expected to experience loss rates that exceed what would be expected by either stressor alone (Tiner 2013, p. 101; Watson et al. 2014, entire).

Nutrient availability in marshes can also influence the amount and distribution of high marsh vegetation by providing a competitive advantage to the low marsh (Levine et al. 1998, p. 291; Bertness et al. 2009b, p. 141) or invasive species at the upland edge (Bertness et al. 2009b, p. 141). At sites with elevated nitrogen in Narragansett Bay, Rhode Island, the transition zone between the high marsh and low marsh was as much as 0.5 m (1.6 ft) higher in elevation than sites without enrichment, indicating that the low marsh vegetation was outcompeting the high marsh plants at the same relative elevation (Bertness et al. 2002, p. 1396). Marshes with increased nitrogen input from adjacent development also have more extensive invasion of Phragmites into the high marsh zone (Bertness et al. 2002, pp. 1395–1398). The resulting loss of high marsh vegetation directly impacts the amount of suitable habitat available for nesting by saltmarsh sparrows.

3.1.2.3 Altered Hydrology

Tidal Restrictions & Impoundments - Infrastructure such as dikes, roads, railroads, and associated culverts or tide gates have negatively affected salt marshes by creating tidal restrictions. Generally these activities have resulted in negative impacts on salt marsh ecology through disruption and limitation of natural tidal flow (i.e., flooding and drainage). This results in decreases in sediment accretion rates, soil water salinity, nutrient exchange, and use by fish and certain saltmarsh dependent bird species (Roman et al. 1984, p. 141; Brawley et al. 1998, pp.

32 629–632; Gedan et al. 2009, p. 127; Correll et al. 2017, pp. 178–179). The decrease in salinity also creates more optimal conditions for Phragmites (see Invasive Species section above), which can invade suitable nesting habitat for saltmarsh sparrow. On the New Jersey shore of the Delaware Bay Estuary, it is estimated that a minimum of 50 percent of the marshes (18,562 ha (45,868 ac)) were impounded to promote the farming of marsh hay between 1931 and 2008.

Upon reintroduction of tidal flow this resulted in the conversion of 4,000 ha (9,884 ac) of formerly impounded marsh to open water. Similar elevation deficits at impounded sites have been documented in other parts of the range (Roman et al. 1984 [Connecticut], Portnoy and Giblin 1997, entire [Massachusetts]), indicating that these sites would be unable to support marsh vegetation if tidal flow were restored without restoration of the marsh elevation (Smith et al. 2017, pp. 37–38).

Historically, impoundments may have created conditions that supported high marsh communities. However, nest success would have likely been reduced relative to natural marshes due to haying and grazing activities, and documented increased predation rates in restricted marshes (DiQuinzio et al. 2002, pp. 183-184). Nests are also vulnerable to flooding during breeding season rain events that trap freshwater on the marsh surface. Across the range of the species, there is a strong negative correlation between tidally restricted marshes and saltmarsh sparrow population trends, with saltmarsh sparrow populations declining annually at 9 percent from 1998 to 2012 at tidally restricted sites (Correll et al. 2017, p. 177).

Ditching - Construction of ditches in salt marshes was a common practice along the Atlantic seaboard during the 1800s and early 1900s. It is estimated that over 90 percent of tidal marshes in New England experienced ditching before the 1940s (Bourn and Cottam 1950, entire) and numbers are similar for other areas along the east coast. Ditching was intended to drain the marshes to allow for grazing, hay production, and later the reduction of mosquito breeding.

Researchers have hypothesized that marsh ditching may have initially benefited saltmarsh sparrow in that the practice increased the area of high marsh. However, recent studies have found that the altered water table, reduction to natural sedimentation, and trapping of water on the marsh following high tide or precipitation events as a result of ditching has resulted in compaction, degradation, and a net loss of high marsh habitat (Vincent et al. 2013, pp. 618–623; Raposa et al. 2017a, p. 394; Raposa et al. 2017b, p. 647) over time. These losses of interior marsh habitat exceed what has been observed at natural (un-ditched) marshes (Adamowicz 2018 [Aug. 22, 2018 [pers. comm.]]). Natural drainages within marshes also have more sloping banks, and support significantly greater vegetative cover and biomass as compared to ditched habitat. The sloping banks promote sediment accretion and flushing allowing marshes to keep pace with sea-level rise through self-maintenance (Vincent et al. 2013, p. 623). In ditched marshes, habitat is subject to subsidence which creates wetter conditions and alters the vegetation community composition, sediment trapping, and soil organic content (Vincent et al. 2013, pp. 618–623). Marsh vegetation subjected to hydrologic stress also has a reduced ability to withstand or recover from sea-level rise and other stressors (Silliman et al. 2009, entire; Haaf et al. 2015, entire).

33 3.1.2.4 Suspended Sediment Concentrations

Suspended sediment concentrations are the concentrations of the fine sediments of a stream (fluvial) or bay (estuarine & marine) sediment load which are normally carried in suspension. The amount and size of sediment can vary significantly according to the flow of the water source and sediment availability. Sediment accretion leading to heightened elevation in salt marshes involves dynamic feedback loops. Tidal inundation provides a mechanism for suspended sediments to be delivered to the marsh surface while also promoting plant growth. Increased plant biomass in return increases the potential for the trapping and deposition of suspended sediments such that both mineral and organic material promote the vertical gain in marsh elevation (Tiner 2013, p. 398-400). Marshes in different portions of the saltmarsh sparrow range and in different geomorphic settings rely differentially on the presence and abundance of either marine or river derived sediments. The sources of sediment can be from off-site or within site locations. Assessing changes in the availability of sediments from all sources is difficult, but is a critical component to understanding marsh persistence.

Within the range of the saltmarsh sparrow, the mid-Atlantic coast was experiencing the greatest reductions in the amount of suspended sediment concentrations with smaller declines in New England and the Southeast coast (Weston 2014, p. 11). Evaluated as percent decline annually over the past 20 to 60 years both mid-Atlantic and New England rivers experienced the greatest losses (> 1 percent per year). Sediment supply in general is related to watershed size, land use, and soil erodibility, but declines over time were most closely correlated with increased retention behind dams (Weston 2014, p. 13). An assessment of tidal inlets between Maine and New York also highlights alterations to the distribution and availability of marine sediments with 70% of tidal inlets artificially modified and directly affecting sediment transport (Rice 2015, entire). When combined with the higher rates of relative sea-level rise along the Northeast coast, the decline in suspended sediment concentrations was a factor in increasing vulnerability to loss of marsh in the mid-Atlantic and to a lesser extent in New England (Weston 2014, p. 21, Figure 10).

3.1.2.5 Shoreline Protective Measures

Approximately 14% of the coastline of the United States has been armored, with rates of shoreline hardening positively correlated with housing density, gross domestic product, storm frequency and wave height (Gittman et al 2015, entire). In many developed areas along the east coast shorelines have been modified through the addition of sea walls, jetties, bulkheads, riprap, revetments, and other hardened structures intended to deflect wave energy, control erosion, and protect infrastructure. The often unintended consequence is increased erosion in adjacent areas due to the increase in deflected wave energy, interruption of the landward migration of marshes (Lentz et al. 2016, p. 1; Mitchell et al. 2017, pp. 11–12), and disruption of the longshore current which would be a source of sediment transported to the marsh surface during high tides (Wilson et al. 2007, p. 126). Shoreline hardening is not unique to this portion of the range of saltmarsh sparrow, and each state has different coastal zone management laws and regulations (see Conservation Efforts). Shoreline hardening associated with development within 1.5 km (0.9 mi) from a marsh edge is, in fact, the most important predictor of marsh loss in the Chesapeake Bay (Mitchell et al. 2017, p. 11). Marsh loss then directly impacts the amount of suitable nesting habitat .

34 available for saltmarsh sparrow. Additionally, artificial shorelines constructed along tidal marsh edges often block efficient drainage of water after high tides or storm surges, effectively creating tidal restrictions that increase the risk of nest flooding and the direct loss of eggs or chicks, while also potentially altering habitat composition within the marsh

Comprehensive summaries were not found for each state within the range of the saltmarsh sparrow, preventing us from quantifying the impacts throughout the entire range, but data were available for portions of the mid-Atlantic. Between 1978 and 1997, for example, close to 500 km (311 mi) of the Maryland shore was armored (Titus 1998, pp. 1281–82, Appendix II; Wilson et al. 2007, p. 126). In Virginia, 400 km (249 mi) of shoreline was armored between 1993 and 2004 (Hardaway and Byrne 1999, p, 24), resulting in 40 percent of the marshes blocked from landward transgression and vulnerable to sea level rise induced losses (Bilkovic et al. 2009, p. 42). Within the Chesapeake Bay, 18 percent of the tidal shoreline has been hardened (Bilkovic and Mitchell 2017, p. 294; Virginia Institute of Marine Science 2018, entire) while 17 percent of the New Jersey coastline has been hardened and approximately 29 percent of the tidal marsh retreat area is presently limited by development and roads (Lathrop and Love, 2007, p. 11-13). Factors that influence a landowner’s decision to select hardened structures are variable but are positively influenced by pre-existing hardened shorelines (Field et al. 2017). Additional research evaluating the impacts of shoreline hardening on coastal habitats and species is summarized in a new edition of Estuaries and Coasts (Prosser et al. 2018, entire).

3.1.3 Human Disturbance and Management Activities 3.1.3.1 Fire

The use of fire through prescribed burning is a frequent practice in the mid-Atlantic and Gulf coast regions (Nyman and Chabreck 1995, p. 134). Fire has been used to enhance wildlife (primarily muskrat and waterfowl) habitat, promote habitat for rare and threatened plant species, control invasive plants, and reduce hazardous build-up of fuel which may lead to wildfire conditions (Nyman and Chabreck 1995, pp. 135–136; Mitsch and Gosselink 2000, entire; Stevenson et al. 2002, p. 721).

Little data is available on the direct effects of prescribed burning on saltmarsh sparrows. One study (Kern 2010, entire) evaluated the impact of four prescribed burn treatments on vegetation cover and breeding tidal marsh birds, including saltmarsh and seaside sparrows, at Blackwater National Wildlife Refuge and Fishing Bay Wildlife Management Area in Maryland. This study indicated that prescribed burning had neither a discernably positive nor negative impact on saltmarsh sparrows. Burn treatment was not a strong predictor of saltmarsh sparrow occupancy (for adult birds) compared to site-specific factors (Kern 2010, p. 50-69). The 0 years-since-burn treatment had significantly less thatch (dead vegetation) than the other treatments (Kern 2010, p. 34), yet more than twice as many saltmarsh sparrow nests were observed there (n=15) than any other treatment (Kern 2010, p. 70). This result was unexpected given the association of nest sites with thatch elsewhere in the range (Gjerdrum et al. 2005, p. 858; Shriver et al. 2007, p 555); however, it is likely that this finding was simply spurious (R. Longenecker, pers. com). First, the higher number of nests in the 0 years-since-burn treatment may have been an artifact of observability, as the lack of thatch increased nest visibility to researchers (R. Longenecker, pers.com.). Furthermore, there was a relatively small sample size of saltmarsh sparrow nests in this study (n=30 across all treatments; Kern 2010, p. 70), limiting the ability to draw statistically rigorous conclusions about the effects of burning on nest-site selection or nest success. Lastly,

35 none of the burn treatments were completely unsuitable for nesting, as saltmarsh sparrow nests were observed in all four types (Kern 2010, p. 70).

Despite the inconclusive findings of Kern (2010), prescribed fire may impact saltmarsh sparrow nesting and overwintering habitat by removing thatch, altering vegetation species composition, and reducing vegetative cover. The use of prescribed fire reduces thatch (Kern 2010, p. 34; Flores et al. 2011, p. 38, 42), which is associated with nest sites. Thus, frequent burning of breeding habitat may reduce habitat suitability or nest success. Fire can alter plant species dominance, depending on the season in which it is applied. For example, burning in fall can promote Scirpus olneyi, while burning in spring can promote Spartina patens (Chabreck 1981, p. 209). Fire frequency can also affect plant species dominance, as annual burning was found to increase cover of Distichlis spicata compared to areas under a three-year burn rotation (Flores et al. 2011, p. 39-42). However, vegetative response to prescribed fire may be dependent on salinity, precipitation, temperature and other environmental factors (Flores et al. 2011, p. 42; Mitchell et al. 2006, p. 158). Although no studies have evaluated the impact of prescribed fire on overwintering saltmarsh sparrows, fire reduced habitat suitability for overwintering seaside and Nelson’s sparrows in Louisiana. Seaside and Nelson’s sparrows were absent from burned areas for several months post-fire when vegetation cover was reduced, but returned the following winter after vegetation regrowth (Gabrey et al. 1999, p 599-603). Assuming saltmarsh sparrows use the same types of overwintering habitat as these congeners, we can conclude that prescribed fire may reduce habitat suitability temporarily.

Fire may also cause direct or indirect mortality, depending when it occurs during the saltmarsh sparrow’s annual cycle. Although prescribed burns are uncommon during breeding season, natural fires can occur and may cause mortality of eggs, chicks, and/or juvenile birds (Nyman and Chabreck, p. 135, 138). Individuals that are molting may also fail to escape a fire. Winter prescribed burns will not result in mortality of eggs, chicks, or juveniles, but have been known to directly displace marsh birds (Gabrey et al 1999, p. 599-603) or cause direct or indirect mortality of marsh birds. Mortality of black rails (Laterallus jamaicensis jamaicensis), a secretive marsh bird, was observed following an intense burn in Florida (Legare et al. 1998, p. 114).

Indirect mortality can also occur; smoke from fires can attract raptors, which have been observed preying upon black rails as they escape (Grace et. al. 2005, p. 6). General recommendations for reducing direct and indirect mortality of wildlife during prescribed burns include providing escape cover by leaving unburned refugia, buringin in staggered rotations, and reducing the frequency of burns (Block et. al. 2016, p. 16; Grace et. al. 2005, p. 35; Legare et. al. 1998, p. 114).

3.1.3.2 Open Marsh Water Management

Open Marsh Water Management (OMWM) has been utilized at marshes throughout the Northeast since the late 1970’s as a strategy to reduce mosquito densities and minimize the use of pesticides on tidal marshes. The strategy involved creation of deep pools and a series of associated ditches that would support fish throughout all tidal stages. This was an effective strategy at reducing mosquito breeding (James-Pirri et al. 2009, p. 1394) and was common practice in New Jersey and Delaware in particular (Mitchell et al. 2006, p. 167). There was an increase in the number of nesting sparrows at the OMWM site after one growing season in a New Jersey study, with numbers returning to pre-treatment levels by year two (Meredith and Lesser 2007, p. 64-65). A similar study in Massachusetts found birds using the site one year

36 post OMWM in lower numbers that pre-restoration, but at higher abundances two years post management (Brush et al. 1986, p. 192-193). In Delaware, a comparison of limited to extensive OMWM found relative abundance of tidal marsh obligate birds (saltmarsh sparrow, seaside sparrow, clapper rail) were higher at sites following the limited OMWM treatment as compared to extensive OMWM (Pepper 2008, p. 12).

3.2 Effects of Climate Change

Within the range of the saltmarsh sparrow, climate change is influencing the rates of sea-level rise as well as storm frequency and severity. The capacity of a tidal marsh to maintain itself is tied directly to the elevation of the marsh surface relative to local mean high water and the associated frequency and duration of flooding. Increased flooding events impact the natural dynamic ecosystem processes of marshes and can limit their ability to recover from stochastic or catastrophic events. Below we discuss sea-level rise, increased storm frequency and severity, and the impacts of flooding on the saltmarsh sparrow and its habitat.

3.2.1 Sea Level Rise and Marsh Subsidence Globally, sea-level increased an average of 1.7 (1.5 to 1.9) millimeter (mm)/year (yr) (0.06 to 0.07 in/yr) between 1901 and 2010, but during more recent years rates have been increasing such that between 1971 and 2010 the global average was 2.0 (1.7 to 2.3) mm/yr (0.07 to 0.09 in/yr) and between 1993 and 2010 it was 3.2 (2.8 to 3.6) mm/yr (0.11 to 0.14 in/yr)(IPCC 2014, p. 42). Rates of sea level rise in specific regions can be several times larger or smaller than these global mean numbers due variations in ocean circulation (IPCC 2014, p. 42). Between 1950 and 2009, for example, marshes between Cape Hatteras, North Carolina and Boston, Massachusetts experienced accelerated sea-level rise that exceeded the global average by 3 to 4 times primarily driven by a slowdown in the Atlantic Meridional Overturning Current (AMOC) related to changes in temperature, salinity and circulation (Sallenger et al. 2012, entire). Throughout the range of the saltmarsh sparrow, there have been assessments of how sea- 2000 to 2013 it was 7.5 mm/yr. (0.30 in/yr) (Raposa et al. 2017a, p. 393). A detailed assessment of 49 SETs across five marshes in Rhode Island found a mean rate of elevation gain of 1.4 mm/ yr (0.05 in/yr). None of the sites were keeping pace with sea-level rise, and all are currently below the elevations where maximum productivity would occur for marsh plants (Raposa et al. 2017a, p. 394). An evaluation of fourteen marshes in Connecticut and New York documented that they were accreting at a faster pace in 2012 (3.6 mm/yr. (0.14 in/yr)) than they were historically (1.0 mm/yr. (0.04 in/yr) in 1900), presumably in response to sea-level rise. Nonetheless, the average accretion rate of 2.3 mm/yr (0.09 in/yr) over a 112 year (1900–2012) time period is less than the documented sea-level rise rates over that same time frame (averaged 3.1mm (0.12 in/yr)+/- 0.1 mm/yr. (0.004 in/yr)(Hill and Anisfeld 2015, pp. 188). This is leading to a decline in marsh surface elevation relative to sea level (Hill and Anisfeld 2015, pp. 4–6). In the Chesapeake Bay, median sea-level rise between 1969 and 2014 was 4.1 mm/yr (0.16 in/yr) with a range of 3.24 to 5.11 mm/yr (0.13 to 0.20 in/yr), and the rate accelerating at 0.08 to 0.22 mm/yr (0.003 to 0.009 in/yr) (Boon and Mitchell 2015, pp. 1299–1301).

37 These documented declines in marsh elevation relative to sea-level rise have resulted in localized marsh loss ranging from 2 to 78 percent over the past 40 years between Cape Cod and Chesapeake Bay (New York - Hartig et al. 2002, pp. 77–82; Chesapeake and Delaware Bays - Kearney et al. 2002, pp. 174–176; Cape Cod - Smith 2009, p. 189; Long Island - NYDEC 2015, entire; Connecticut & New York - Hill and Ainsfield 2015, p. 5; Smith 2015, pp. 132–134; Beckett et al. 2016, p. 2; Mitchell et al. 2017, pp. 5–9; Rhode Island - Raposa et al. 2017a, p. 393).

In Southern New England, marshes with an existing marsh elevation below local mean high water are experiencing greater declines in marsh area than marshes at elevations above mean high water. Areas above mean high water have plants that experience optimal growth of above and below ground biomass leading to more organic matter accumulation and associated marsh elevation gain (Watson et al. 2014, Fig. 1A & 1B; Watson et al. 2017b, p. 670). Modelling of tidal inundation and plant biomass showed that under high tides there was a negative feedback loop with a decrease in biomass leading to reduced peat formation that, in turn, resulted in further elevation loss, and the decline of Spartina patens (high marsh) in marshes (Watson et al. 2016, p. 117). The majority (87 percent) of marshes from a sample of 38 across New York, Connecticut, Rhode Island, and Massachusetts marshes were below optimum elevation for plant biomass production (Watson et al. 2014, p. 505; Figure 1 D) which puts them at increased risk for continued loss (Raposa et al. 2017a, p. 394)

Although sea-level rise is the primary driver for marsh loss across the region, variation between sites is attributed to additional stressors acting synergistically and can include herbivory, fungal pathogens, poor water quality, altered hydrology, as well as sediment availability, soil composition, temperature, and other specific local factors (Hughes 2004, pp. 23–26; Smith 2009, pp. 200–205; Kirwan and Megonigal 2013, pp. 53–59; Smith 2015, pp. 131– 135; Raposa et al. 2017a, p. 394; Watson et al. 2017b, pp. 676–677).

Marshes in the mid-Atlantic are also losing elevation relative to sea level, with the effects of sea- level rise exacerbated by deep subsidence related to geological processes which result in 1.4 mm/ yr. (0.06 in/yr) of additional elevation loss in New Jersey (Kemp et al . 2013, p. 98). A study of three marshes in the Delaware Bay and three in Barneget Bay, NJ tracked accretion from 2010 to 2015 and found that marshes were not keeping pace with sea level rise at 10 of 18 stations and losses were documented at each of the six marshes (Haaf et al. 2017, p. 19)

Marshes in the Nanticoke sub-estuary of the Chesapeake Bay are losing elevation at an average of 1.8 mm/yr (0.07 in/yr) relative to sea-level despite the fact that they are accreting material at the surface as high as 19 mm/yr in some areas (range 7 to 19 mm/yr (0.28 to 0.75 in/ yr)) (Beckett et al. 2016, p. 6). Because the loss was not correlated with position in the estuary or salinity, the researchers suggested that it may be driven by subsurface and surface processes (from adjacent development and agriculture) such as groundwater withdrawal (Miller et al., 2013 p. 13) or increased nitrogen inputs (from agriculture or development) which are driving microbial decomposition of belowground organic matter (Beckett et al. 2016, p. 8). In other portions of the Chesapeake Bay, marshes have been able to maintain their extent with relative sea-level rise through landward transgression where development is limited (Schieder et al. 2018, pp. 947– 949). Marshes in the sediment-rich upper reaches of the Delaware Bay may also

38 be keeping pace with sea-level rise with accretion rates of 25-27 mm/yr or more in these tidal freshwater systems, while surface elevation of salt marshes in the lower bay are only increasing by 1.58 +/- 0.99mm/yr (Kreeger et al. 2012, pp. 37-47). Combined with erosive wave energy and increased tidal volumes, this is resulting in the loss of approximately 0.4 ha (1.0 ac)/day in lower parts of the estuary (Miller et. al. 2012, p. 139; Kreeger et. al. 2016, entire).

3.2.2 Increase in Storm Frequency, Duration, and Severity

In the Northeast, there has been a 71 percent increase in heavy rainfall and flooding events with regard to both the frequency and intensity of events between 1958 and 2012 (Walsh et al. 2014, pp. 36–40). During the breeding season, increased frequency and intensity of precipitation can lead to direct losses of eggs and chicks due to flooding, with losses most pronounced when the tidal cycle is closest to the peak spring tides. Nests placed in areas where tidal restrictions prevent adequate drainage of water from the marsh, or where impervious surfaces around the marsh lead to excessive surface water flow from developed areas onto the marsh, would also be expected to experience higher rates of nest loss as a result of prolonged inundation of nests. Saltmarsh sparrow trends at marshes in Maine were negatively impacted by severe precipitation events that resulted in nest failure during June and July (Shriver et al. 2016, pp. 193 and 197). Although flooding of the marsh surface related to precipitation during the summer has the potential to increase nest failure, the impacts to marsh vegetation are less clear. A study of habitat change on Rhode Island marshes noted that increased duration of tidal inundation led to significant impacts to high marsh vegetation while various precipitation scenarios did not (Watson et al. 2016, p. 117).

Two evaluations of vegetative impacts to New Jersey marshes following Hurricane Sandy, which occurred during the post-breeding season (October 2012), documented no widespread change in vegetative composition or cover as a result of the storm (Elsey-Quirk 2016, p. 247; Longenecker et al. 2018, pp. 569-575). These studies indicate that high marsh vegetation was largely resilient to Hurricane Sandy over the short-term (1-2 years pre-storm compared with 1 year post-storm) and that it may be resilient to future hurricanes. However, no direct comparisons of saltmarsh sparrow nest success before and after Hurricane Sandy were made. Longenecker et al. evaluated predation rates on artificial sparrow nests, designed to simulate saltmarsh sparrow nests, and found no change in artificial nest predation rates before and after Hurricane Sandy (Longenecker et al. 2018, p. 568-569). Thus, they suggested that saltmarsh sparrow nest predation rates may have been similarly unaffected by the Hurricane, but caution against assuming that it had no impacts on reproductive success at all. It is possible that Hurricane Sandy altered other resources necessary for saltmarsh sparrow reproductive success, such as thatch availability or invertebrate food sources, but those impacts have not be measured (Longenecker et al. 2018, pp. 571-572).

3.2.3 Increase in Flooding Events

Flooding has been documented as a primary cause of nest failure, is reducing overall nest survival rates for saltmarsh sparrow (Gjerdrum et al. 2005, p. 582; Shriver et al. 2007, p. 556; Bayard and Elphick 2011, pp. 397, 401), and has been correlated with recent population declines (Shriver et al. 2016, p. 7). As discussed above, saltmarsh sparrows require approximately 26 days between flooding events to allow their young to reach fledgling stage. However, at some 39 sites as many as 85 percent of nests experience flooding at least once during incubation (Bayard and Elphick 2011, p. 401). Normal flooding of the marsh surface from tidal actions is compounded by precipitation events or storm driven tides which can lead to nest flooding in between peak spring tide periods; however, sea-level rise is the major driver for increased flooding of marshes. Saltmarsh sparrow eggs and nestlings are vulnerable to flooding from flooding events that result in prolonged inundation, repeated flooding over multiple days, or extreme flooding outside of the expected tidal cycle (Ruskin et al. 2017a, p. 910). Saltmarsh sparrows, as an apparent adaptive strategy to avoid nest flooding, build nests higher off the ground in the vegetation, or build nests at higher elevations within the marsh. Both these strategies increase the susceptibility to predation as nest are often more visible or are in areas more easily accessed by predators (see Predation section below).

Marsh vegetation potentially can respond to this new flooding regime by moving up slope to the appropriate elevation relative to the new mean high water. If the marsh vegetation and sediment accretion rates do not respond at a pace equivalent to sea-level rise or areas are not available for landward transgression (due to development or other alterations) the high marsh habitat used by the saltmarsh sparrows for nesting will transition to low marsh and will no longer be able to support nesting.

3.3 Hybridization The saltmarsh sparrow is known to hybridize with the closely related Nelson’s sparrow within an area of range overlap (hybrid zone) (Shriver et al. 2005, entire; Walsh et al. 2015b, entire; Walsh et al. 2016b, entire; Walsh et al. 2017b, entire; Maxwell 2018, entire; See also section 2.1 Taxonomy ). Hybridization can lead to reduced reproductive success, and disrupt long-term adaptations of native populations (Allendorf et al. 2004, p. 1203), but can also be beneficial in terms of increased adaptive capacity through increased genetic diversity (Fitzpatrick et al. 2015, pp. 43–46).

Where the saltmarsh and Nelson’s sparrow ranges overlap, despite the presence of hybrid individuals, there are very few first generation hybrids (pure Saltmarsh x pure Nelson’s) on the landscape (Shriver et al. 2005, p. 103; Walsh et al. 2015b, p. 708; Walsh et al. 2016a, p. 3; Maxwell 2018, p. 21). This suggests that genetic swamping, where the local genotypes are replaced by hybrids, is not occurring and that the species are remaining distinct, despite the expansion of the hybrid zone over the past 15 years (Walsh et al. 2017b, p. 463). However, within the center of the hybrid zone (see Figure 3; Taxonomy), researchers found local, site-specific characteristics influenced the extent of hybridization across a habitat gradient with higher numbers of hybrids and asymmetrical backcrossing (hybrids breeding with pure individuals) towards Nelson’s sparrows at a riverine, fringing salt marsh compared to a coastal, back-barrier salt marsh (Maxwell 2018, p. 2). This may be explained by the more abundant Nelson’s sparrow in these riverine systems and the absence of saltmarsh sparrows, which are saltmarsh specialists.

While there is no evidence that supports reduced fertility in hybrid individuals, researchers found nearly twice as many hybrid female nestlings than adult female hybrids, suggesting hybrid females may have reduced survival between their nestling to adult life stage (Maxwell 2018, pp.

40 23–24). There is also evidence for differential fitness (i.e., nest success) between saltmarsh sparrows and Nelson’s sparrows, suggesting that even though the genetic composition of a female is an important predictor of nesting success, it is likely secondary to other predictors of nest failure, such that saltmarsh sparrows will have higher nesting success than Nelson’s/hybrid sparrows during drought years, but under sea-level rise conditions there is too much flooding for either species o be successful (Maxwell 2018, p. 69).

Walsh et al. (2017b, entire) conducted a formal assessment of the conservation implications of hybridization between saltmarsh and Nelson’s sparrows following methods developed by Jackiw et al. (2015, entire). This study resulted in the conclusion that despite potential increasing levels of gene flow between the two species, hybridization appears to have minimal consequences at the population level relative to other imminent threats, such as sea-level rise and habitat degradation.

3.4 Predation As discussed above, predation and nest flooding rates can be interrelated at any given marsh (see Increase in flooding events above). More broadly, flooding impacts the saltmarsh sparrow throughout the breeding range while predation exerts a greater impact on the southern portion of the breeding range where approximately 33 percent of the saltmarsh sparrows population breeds. Across the breeding range, the following species are likely predators of saltmarsh sparrow: garter snake (Thamnophis sirtalis), fish crow (Corvus ossifragus), long-legged waders, northern harrier (Circus hudsonius), gulls (Charadriiformes), short-eared owl (Asio flammeus), blackbirds (Icteridae), coyote (Canis latrans), raccoon (Procyon lotor), domestic cat (Felis domesticus), rats (Rattus spp.) and meadow vole (Microtus pennsylvanicus) (Greenlaw et al. 2018, (Demography and Populations) unpaginated). River otter (Lontra canadensis), red fox (Vulpes vulpes), mink (Neovison vison), and marsh wren (Cistothorus palustris) are also potential nest predators (Roberts 2016, p. 61).

Rangewide, from 2011 to 2015, 15 percent of saltmarsh sparrow nests were lost due to predation (Greenlaw et al. 2018, (Demography and Populations) unpaginated), the second highest cause of failure after flooding. Predation rates are not consistent across the range, with daily predation probabilities varying from less than 0.001 to 0.087 (Ruskin et al. 2017a, p. 909) and predation probability increasing in the more southern parts of the range (Ruskin et al. 2017a, p. 909). Researchers found that 29 percent of nests were lost to predation at Forsythe NWR, New Jersey from 2011 to 2015 (Roberts et al. 2017, p. 123). This was the leading cause of nest loss in this study area and may be a driver of population declines at the site (Roberts et al. 2017, p. 127). Only a few other studies have evaluated nest predation rates. A study in New York observed a 20–52 percent loss of active nests to predation from 1977 to 1980 (Greenlaw et al. 2018, (Demography and Populations) unpaginated); in Rhode Island, 50–100 percent of failed nests were attributed to predation (DiQuinzio et al. 2002, p. 182); and in Connecticut, 31 percent of failed nests were due to predation (Gjerdrum et al. 2005, p. 855).

Characteristics of a particular nest site or marsh likely influence the risk of nest predation. For example, depredated nests have been higher off the ground in the vegetation and in higher elevation areas (Benvenuti et al. 2018b, pp. 6–7). Likewise, predation risk may influence nest

41 placement as females whose first nest was depredated constructed the following nest at lower elevation sites and lower down in vegetation (Benvenuti et al. 2018b, pp. 11–12). In Connecticut, the density of breeding adults was related to the distance to the marsh edge, suggesting that birds may avoid areas where mammalian predators are likely to be more common (Gjerdrum et al. 2008b, pp. 612–614). A follow-up study did not find evidence that distance to the marsh edge was especially important (Meiman et al. 2012, p. 734). Neither study examined whether characteristics of the marsh edge (e.g., habitat type, developed, etc.) or the presence of anthropogenic structures (e.g., docks extending into the marsh) affect either predator behavior or nest success. In urban areas, introduced predators (e.g., cats) or high densities of mesopredators (e.g., raccoons, skunks, snakes) supported by ancillary food provided by humans may have local effects, although this hypothesis remains to be fully tested.

Predation on the wintering grounds has not been studied. However, one study of winter survival found that survival rates to be relatively high (Borowske 2015, p. 104), suggesting that predation on the wintering grounds may not be a significant concern.

3.5 Contaminants 3.5.1 Environmental Contaminants

Tidal marshes have experienced and continue to be vulnerable to exposure to pollutants in addition to fertilizer and other nutrient sources introduced through groundwater or surface water flow. Accumulation of mercury, polychlorinated biphenyls (PCBs), dichlorodiphenyltrichloroethane (DDT), selenium, trace heavy metals, and other contaminants are also present in marshes where they can bind with the soil and bioaccumulate in organisms and wildlife. New generation, broad spectrum pesticides are also applied in many areas to control mosquitoes (Greenberg et al. 2006, p. 683).

3.5.1.1 Mercury

Numerous avian studies have documented the deleterious impacts of methylmercury (MeHg) exposure on birds. Impacts range from changes in bird behavior and health to altered reproduction. Many of these changes (especially at levels classified as sub-lethal) are subtle, and various species have different tolerance levels for exposure. A review paper by Whitney and Cristol (2017, entire) documents the current state of knowledge about sub-lethal effects of MeHg in birds and classifies birds as low, medium or high sensitivity species. The saltmarsh sparrow is considered as a medium sensitivity bird (Whitney and Cristol 2017, p. 8). The classification is supported by the collection and analysis of brain tissue from a saltmarsh sparrow fledgling from a high mercury site (mean blood mercury level for all sparrows at site 1.5 ± 0.32 parts per million (ppm) (Scoville and Lane 2013, p. 617). The sparrow had brain abnormalities consistent with those caused by exposure to toxic levels of mercury (Scoville and Lane 2013, p. 619).

Two historical studies have found elevated levels of MeHg in saltmarsh sparrows, as well as geographic, site specific, and annual differences in MeHg exposure (Shriver et al. 2006, p. 132; Lane et al. 2011, p. 1987). In both studies saltmarsh sparrows were found to have almost two times the amount of mercury in their blood as Nelson’s sparrows even though they both co-occurred within the same habitat. Blood collected from southern Maine saltmarsh sparrows

42 (n=53) had values ranging from 0.342 to 1.260 ppm with an average of 0.69 ppm (Shriver et al. 2006, p. 131). Lane et al. (2011, p. 1,987) found 27 percent of the saltmarsh sparrow blood samples (n=653) collected from Maine, Massachusetts, New Hampshire, Connecticut, Rhode Island, and New York marshes had mercury concentrations greater than 1.0 ppm and the study found males had a tendency towards lower levels of Hg (p=0.054). Lowest mercury values were found in Connecticut, while the highest levels (over all mean 1.32 ppm, highest annual mean 1.8 ppm) were found at Parker River NWR in Newburyport, Massachusetts. Summer blood mercury levels for saltmarsh sparrows in the Delaware Bay had average concentrations between 0.40 and 0.54 ppm (Warner et al. 2010, p. 674).

The mercury exposure profile appears to change for the species during the non-breeding season. Sparrows were able to reduce their mercury burdens while on their wintering grounds, limiting the exposure to high levels of mercury to time spent during the 12-week breeding season (Cristol et al. 2011, p. 1777). Winder and Emslie (2012, p. 331) found a similar pattern, with November blood levels averaging 0.361 ± 0.033 ppm.

Recent research conducted by Ruskin et al. (in prep., entire) on saltmarsh and Nelson’s sparrows did not find an impact of female blood mercury levels (geometric mean 0.352 ± 0.019 (SE) µg/g) (1 ppm = 1 µg/g) on any reproductive outputs (clutch size, brood size, daily nest survival).

Several researchers have posited that sub-lethal impacts, if they exist, are difficult to measure and might only be observed at saltmarsh sparrow blood levels around 1.0 to 4.0 µg/g (Lane et al. 2011, p. 1989). It is known that negative impacts from mercury exposure are not consistently expressed across species, or at degrees of specific levels of exposure. Nonetheless, Winder (2012, p. 332) proposed that greater than 58 percent of saltmarsh sparrows are at risk of negative effects from mercury based on the number of birds with mercury concentrations above the threshold for moderate to high reproductive risk in other species (i.e., wren and loon). Indeed, there may be impacts to the species at these higher levels; however, the deleterious impacts are likely swamped out by other factors such as tidal flooding and predation (Ruskin et al. in prep.).

As mercury blood concentrations in saltmarsh sparrows vary dramatically by geography and watershed, impacts of mercury exposure will also vary. Determining the percentage of the population which could potentially be at risk is difficult, and further study at sites with higher mercury loads would be warranted.

3.5.2 Oil and Chemical Spills

Throughout the range of the saltmarsh sparrow there are shipping lanes that support the transport of oil and gas in offshore and nearshore waters. This has resulted in spills that have impacted shorelines over the past 50 years, with notable examples being the North Cape oil spill in Rhode Island in 1996, the Bouchard oil spill in Buzzards Bay, Massachusetts in 2003. Depending on the timing of the spill relative to the breeding season of the saltmarsh sparrow, the proximity to shore and saltmarsh habitats, and the ability of agencies to respond, there is the potential for impacts to all life stages of saltmarsh sparrows as well as their habitat and their prey base.

43 3.6 Other Factors Considered

3.6.1 Disease

Little information exists and no systematic studies have been conducted on saltmarsh sparrow disease. Bird banding efforts throughout the species range have not observed any known diseases impacting other sparrow or bird species in the saltmarsh sparrow (i.e., avian pox or other specific diseases) (Greenlaw et al. 2018, p. 45). West Nile virus has not been specifically detected in saltmarsh sparrows, but has been detected in other sparrow species whose ranges overlap with the saltmarsh sparrow (i.e., chipping sparrow, field sparrow, fox sparrow, Lincoln’s sparrow, savannah sparrow, song sparrow, swamp sparrow, and white- crowned sparrow) (Center for Disease Control (CDC) 2017, entire). Regardless, we currently are not aware of any known diseases that have been identified which impact saltmarsh sparrows at this point in time.

3.6.2 Altered Food Webs

Indirect impacts from locally abundant geese grazing can lead to a loss of vegetative cover in tidal marshes that are targeted repeatedly, but these overgrazed areas tend to represent a small proportion of the marshes available in the mid-Atlantic and the impacts are focused on the low marsh habitat (Gauthier et al. 2006, p. 110). The loss of low marsh vegetation can lead to an increase in local erosion and exacerbate marsh loss. This reduces available foraging habitat for saltmarsh sparrow, but does not directly impact nesting habitat. There are several crab species which have also been implicated in marsh dieback events. Purple marsh crab (Sesarma reticulatum) populations, in particular, have increased in the Northeast due to the potential depletion of coastal predators. Without the predation pressure, this species is likely causing low marsh die off through overgrazing (Coverdale et al. 2013, p. 1042) of Spartina alterniflora where the soft soil permits burrow creation (Bertness et al. 2009a, pp. 2111–2112). This in turn exacerbates erosion of the marsh edge, with locally driven losses from marsh crabs appearing to interact synergistically with climate change impacts to result in three times more loss than if the individual impacts were simply additive (Crotty et al. 2017, pp. 4–6). Additionally, while the peat associated with high marsh zones limits the ability of crabs to burrow and graze (Bertness et al. 2014, pp. 3–4), increased rates of tidal inundation of high marsh softens the substrate enabling increased crab colonization (Crotty et al. 2017, pp. 7–8).

3.7 Conservation Efforts In addition to our evaluation of threats (Sections 3.1-3.6), we looked at existing conservation efforts and opportunities. This includes opportunities for habitat restoration, the current state of knowledge with regard to best management practices for marsh restoration and any evaluation of effectiveness in restoring nesting habitat for this species. Additionally, we summarize existing laws that provide protection for tidal marsh habitats or saltmarsh sparrow and conservation tools that are currently available or under development that may facilitate the prioritization, funding and implementation of additional conservation actions.

44 3.7.1 Habitat Restoration

In response to the dramatic decline in tidal marshes, and the more recent recognition of their ecological values, tidal marsh restoration has been a focus of state and federal agencies over the past 30 years. The primary objective has been the reversal of past actions that converted former wetlands to uplands through filling or restriction of tidal flow, which resulted in invasion by Phragmites. As such, restoration efforts have been focused on removal of Phragmites through application of herbicide, mowing, and burning; plugging or otherwise experimenting with ditch alterations; removal of tidal restrictions; or excavating sediment to restore suitable intertidal elevations.

The assumption has been that reversing past alterations would restore the function and values of tidal marshes. Studies designed to evaluate this have found that although marshes will transition to native vegetation following restoration and may support species that forage or nest in low marsh habitats (e.g., wading birds, rails) the restoration practice has not benefited saltmarsh sparrow in the short-term (DiQuinzio 2002, pp. 183–184; Elphick et al. 2015, pp. 442– 443). There has been little attention to targeting elevations that would create high marsh. It is now known that in areas where there have been tidal restrictions, the resultant lack of flooding results in a compaction and oxidation of the marsh peat (Roman et al. 1984, p. 147; Burdick et al. 1997, pp. 139–140; Crooks et al. 2002, p. 591) such that when flow is returned, the elevations are too low to drain at low tide. Where flow was restored in New Jersey marshes, the formerly impounded sites had significantly less high marsh vegetation and significantly more mudflat, open water, and Phragmites (Smith et al. 2017, p. 37). The net result of these restoration strategies has been to create expanses of low marsh habitat with little to no suitable nesting habitat for saltmarsh sparrow.

A study of 14 marshes in Connecticut with restored tidal flow did not support saltmarsh sparrow nesting as compared to 19 reference sites (Elphick et al. 2015, pp. 441–444). Alternately, in the same study the authors found that the removal of Phragmites at seven sites without tidal restrictions resulted in the creation of suitable nesting habitat (Elphick et al. 2015, p. 444). Studies in Long Island Sound showed that it can take from 5 to 21 years for the full restoration of ecological functions and up to 15 years for breeding saltmarsh sparrows to recolonize restored habitat (Warren et al 2002, pp. 497, 507) where tidal flow was restored. Restoration of tidal flow at a previously impounded salt marsh at Barn Island, Stonington, Connecticut, resulted in the dominant vegetation transitioning from Typha angustifolia (cattail) and Phragmites to Spartina alterniflora and S. patens, which provided breeding habitat for saltmarsh sparrows within 10 years of the restoration, although still at vegetation densities below adjacent unditched natural marsh (Brawley et al. 1998, pp. 629–632). Additional studies also confirmed that sparrows returned to sites following restoration, but many of these were not suitable for nesting and resulted in dramatic declines in nest success initially (DeQuinzio et al. 2002, p. 182), cautioning that the presence of the species does not necessarily confirm suitable nesting habitat (Meiman et al. 2012, p. 732). Nonetheless, at this and other sites, if the habitat supports foraging and roosting individuals initially there is some evidence that it will eventually transition to suitable nesting habitat within 10 to 15 years if the conditions support sufficient marsh accretion and high marsh development (Slavin and Shisler 1983, pp. 283–284; Brawley et al. 1998, pp. 629–632; Warren et al. 2002, pp. 497, 507).

45 Ongoing tidal marsh restoration is focused on four primary strategies that are not mutually exclusive and depend on site-specific conditions. These consist of: (1) protection andmanagement of land where marsh migration can be allowed and/or facilitated; (2) marsh restoration focused on building elevation; (3) marsh restoration focused on restoring or improving hydrology; and (4) removal of invasive species.

In the Chesapeake Bay region, there is evidence that marsh transgression into forested areas will continue (http://www.slammview.org), but it represents a small area of potential gain due to the extensive amount of hardened shoreline (Mitchell et al. 2017, p. 10). Other studies provide evidence that marsh expansion into undeveloped and agricultural areas could offset marsh loss and result in a net gain in marsh extent in some areas (Schieder et al. 2018, p. 944). This would be quite extensive in areas like the Delaware Bay if it were not for extensive tidal restrictions which have prevented the natural transgression of marsh into adjacent uplands (Smith et al. 2017, p. 38).

An evaluation of marsh transgression into wooded habitats in southern New England found that transgression into the upland is not advancing quickly enough to offset marsh loss overall, or losses of high marsh in particular (Field et al. 2016b, pp. 366–367 [Connecticut]). Wooded sites also tend to transition to Phragmites or bare ground whereas Spartina patens is more likely to colonize open areas (Anisfeld et al. 2016, pp. 5–9 [New York]). Marshes in more Northern landscapes are often either adjacent to steep slopes that would impede transgression, surrounded by developed land, or predominantly forested and require active management to facilitate transgression at rates that would be necessary to offset losses.

Land managers restoring existing marshes are focusing on restoring the natural hydrology and/or enhancement of elevation if one or both of those actions would be expected to support the colonization and maintenance of high marsh vegetation (Elphick et al. 2015, p. 444). There have been marsh restoration and creation projects focused on enhancing marsh surface elevation through thin layer sediment placement at sites in New York, Delaware, and Rhode Island within the breeding range of the saltmarsh sparrow. These initial attempts are being evaluated to determine what variables (i.e., sediment depth, grain size, local conditions, hydrology, etc.) influence marsh recovery time and will help to inform future restoration strategies (Weston 2014, pp. 15–22; Wigand et al. 2016, pp. 15–21). Additionally, SHARP researchers have established monitoring plots to assess the response of the avian community to these restoration practices (Elphick et al. 2018, pp. 2–3).

Alternatively, various techniques are the subject of experimentation to determine their effectiveness as restoration practices. This includes modifications to creeks, channels, ditches, or other topographic features of the marsh which could influence the frequency and duration of marsh flooding. Since grid ditches experience higher sedimentation rates than the adjacent marsh surface (Corman et al. 2012, pp. 876–878), one strategy has been to fill or plug ditches. However, there is a high degree of uncertainty associated with this technique given that other research has documented an increase in the water table following ditch plugging (Adamowicz and Roman 2002, pp. 15–17; James-Pirri et al. 2008, pp. 168–170). Another technique that is being utilized to improve drainage of heavily ditched sites involves digging of minor runnels (narrow channels) to improve drainage of water from the marsh surface following flooding

46 events. There is some evidence that this results in the recovery of marsh vegetation and an increase in accretion rates (D. Curson, 2019. pers. comm; W. Ferguson, 2019. pers. comm) but the technique has not been tested extensively at this time.

How marshes respond to targeted restoration actions requires an understanding of the primary stressors acting on the specific marsh. Having this understanding leads to the development of appropriate restoration strategies that are site specific (Roman 2016; entire). Even in fragmented urban areas, there is evidence that if less than 50 percent of the border is developed it may be beneficial to restore even small marshes if a network of sites can be maintained (Kocek 2016, pp. 82–83). Researchers are also beginning to evaluate the use of site- specific flood protection strategies during the peak spring tides to protect saltmarsh sparrow nests, for example the use of tide gates to restrict the full range of tides during the new or full moon.

3.7.2 Applicable Laws

The Migratory Bird Treaty Act of 1918 (16 U.S.C. 703 et seq.) is the Federal law providing specific protection for the saltmarsh sparrow due to its status as a migratory bird. The Migratory Bird Treaty Act prohibits the following actions, unless permitted by Federal regulation: to “pursue, hunt, take, capture, kill, attempt to take, capture or kill, possess, offer for sale, sell, offer to purchase, purchase, deliver for shipment, ship, cause to be shipped, deliver for transportation, transport, cause to be transported, carry, or cause to be carried by any means whatever, receive for shipment, transportation or carriage, or export, at any time, or in any manner, any migratory bird...or any part, nest, or egg of any such bird.” Through issuance of Migratory Bird Scientific Collecting and bird banding permits, the Service, USGS and the state wildlife agencies ensure that best practices are implemented for the careful capture and handling of saltmarsh sparrow, eggs, nests and nestlings during banding operations and other research activities.

The Coastal Zone Management Act of 1972 (P.L. 92-583) (86 Stat. 1280; 16 U.S.C. 1451–1464) provides Federal funding to implement the states’ federally approved Coastal Zone Management Plans. All coastal states in the saltmarsh sparrow range have approved Coastal Zone Management Plans, which guide and regulate development and other activities within the designated coastal zone of each state (NOAA Office for Coastal Management, 2018b, entire). The Federal Consistency provision of the Coastal Zone Management Act requires Federal action agencies to ensure that the activities they fund or authorize are consistent, to the maximum extent practicable with the enforceable policies of that state’s federally approved coastal management program (16 U.S.C. 1456).

The Living Shoreline Protection Act (2008) in Maryland requires living shorelines be the first defense against shoreline erosion, but also recognizes that this is most effective in low energy sites and where the ‘fetch’ is less than 5.6 mi (9.0 km) (Tiner 2013, p.230). Policies in other states are less well known, but there have been attempts to identify the most publicly acceptable recommendations for compensatory mitigation to protect marsh acres or implement restoration to mitigate for losses at other sites (Bauer et al. 2004, entire).

The Clean Water Act (CWA) (33 U.S.C. §1251 et seq.) and the Rivers and Harbors Act (33 U.S.C. 403; Chapter 425, March 3, 1899; 30 Stat. 1151) have sections 404 and 10 respectively, that contain provisions for the protection of jurisdictional wetlands from filling

47 activities. The U.S. Army Corps of Engineers in conjunction with the U. S. Environmental Protection Agency administers permits that consider avoidance, minimization, and compensation for projects affecting wetlands. Projects that cannot avoid impacts to wetlands must mitigate their impacts through a restoration action for the equivalent functional loss. Mitigation banks are often used which tend to centralize restorative actions at a specific location for impacts in a considerably wider service area. Exact wetland types are not always restored and there is considerable uncertainty that current mitigation practices support the presence of saltmarsh sparrow. The status of geographically isolated wetlands under the CWA has fluctuated with different court cases and rulings (Rains et al. 2016, entire; Haukos and Smith 2003, p. 582–586).

3.7.3 Conservation Status and Tools For Conservation

Approximately fifty-seven percent of the salt marsh (162,580 ha (401,746 ac)) within the breeding range of saltmarsh sparrows is conserved in some way, whether through fee title ownership or conservation easement. Among the conserved lands, the largest percentage is in state ownership (49 percent) followed by Federal lands (20 percent), and privately-owned conserved lands (15 percent; Table 2).

Saltmarsh sparrow are not currently listed as threatened or endangered on any State lists where breeding occurs, but Connecticut identifies them as a Species of Concern and Maryland includes them as a Tier 3 Species defined as “In Need of Conservation.” All of the States within the breeding range include the species in their State Wildlife Action Plan (SWAP) as a “Species of Greatest Conservation Need” and assign a Natural Heritage Rank (NatureServe 2017, unpaginated).

The NWR System administered by the Service manages close to 10 percent of all of the tidal wetlands within the breeding range of saltmarsh sparrow (Maine to Virginia), representing about 16 percent of all of the conserved lands (26,725 ha (66,038 ac); Table 2). The National Wildlife Refuge System Improvement Act of 1997 (16 U.S.C. 668dd et seq.) establishes the protection of biodiversity as the primary purpose of the NWR system; recreational and other uses of a NWR may only be approved if the Service finds such uses to be compatible with the purposes of that individual NWR and the purposes of the NWR system.

Approximately 5,113 ha (12,634 ac) of the tidal wetlands, representing two percent of the tidal marsh habitat within the breeding range for saltmarsh sparrow, is managed by the National Park Service (NPS) as a Park or National Seashore (Table 2). The NPS must balance visitation and recreation with the protection of natural resources like the saltmarsh sparrow and its habitat. The NPS Organic Act of 1916, as amended (39 Stat. 535, 16 U.S.C. 1), states that the NPS “shall promote and regulate the use of [NPS units]...to conserve the scenery and the national and historical objects and the wildlife therein and to provide for the enjoyment of the same in such manner and by such means as will leave them unimpaired for the enjoyment of future generations.” In addition to the NPS Organic Act, the saltmarsh sparrow may benefit from a 2010 non-regulatory Memorandum of Understanding (MOU) between the NPS and the Service regarding migratory birds that was executed pursuant to Executive Order 13186; section F.4. of the MOU states that the NPS will identify and protect natural habitats of migratory bird species within park boundaries.

48 Table 2. Land Ownership (USGS Gap Analysis Program 2012; unpaginated, USGS Gap Analysis Program 2013, unpaginated) Ownership Acres % of conserved land % of all marsh

Tribal 87 0.02% 0.01% FWS 66,038 16.44% 9.45% NPS 12,634 3.14% 1.81% Other FED 5,226 1.30% 0.75% State 195,040 48.55% 27.90% other public 18,888 4.70% 2.70% Audubon 1,294 0.32% 0.19% TNC 21,427 5.33% 3.07% Land trust, Private non- profit 11,334 2.82% 1.62% Private 59,708 14.86% 8.54% Other conserved 10,066 2.51% 1.44% All conserved 401,746 100.00% 57.47% Not conserved 297,277 42.53% Total acres 699,023

Figure 13. Ownership of occupied tidal marsh and the estimated distribution of individuals from data collected 2011–2012 (Wiest et al. 2016). Total area represented = 282,887 ha (699,028 ac). Total number of individual saltmarsh sparrows represented = 61,829.

49 Conservation properties managed by State, Municipalities, and other public entities account for approximately 31 percent of the tidal wetlands within the breeding range for saltmarsh sparrow (Table 2), representing the largest proportion of conserved lands. Protected lands such as Wildlife Management Areas, State Parks, State Natural Areas, and Preserves typically have rules that protect wildlife and prohibit the collection, destruction, or disturbance of plants and nongame animals. These lands are often managed for a suite of wildlife species while providing outdoor recreation opportunities to the public.

Additional conserved lands include about eight percent of the tidal marsh acres protected by The Nature Conservancy, Audubon, and local conservation organizations and another 15 percent under conservation easement or otherwise protected by private landowners (Table 2). An additional assessment that estimates the number of saltmarsh sparrows that are supported within each of the landowner categories suggests that Federal lands support more individuals than would be expected based on acres owned, while State lands support less than would be expected. (Wiest et. al, 2016; Figure 13).

Federal conservation land owners have undertaken restoration actions to improve habitat resiliency and overall health (e.g., Blackwater NWR, Maryland; Prime Hook NWR, Delaware; Jamaica Bay, New York; Sachuest Point NWR, Rhode Island; Parker River NWR, Massachusetts) although we do not have a comprehensive summary of these actions or the total acres that have been impacted. Tidal marsh restoration has also been supported on non-Federal lands through technical and financial assistance provided by the Service’s Coastal and Partners for Wildlife Programs, with the Coastal Program establishing high marsh restoration and protection that would benefit saltmarsh sparrow as a high priority in their strategic plan (2017– 2021). NOAA also administers a grant program focused on restoration of coastal habitats, including tidal marshes, and they have funded several large scale restoration projects in recent years. The Natural Resources Conservation Service (NRCS) - Wetland Reserve Program (WRP) has identified saltmarsh sparrow as a species that will be factored into the prioritization and selection of private lands to receive funding under their WRP easement / land protection program. This is available throughout the states with suitable nesting and wintering habitat, but the states will each need to define the Geographic Area Rate Caps (GARC) and a ranking process for selection of eligible sites. To our knowledge this has already happened in Rhode Island, with saltmarsh sparrow distribution being incorporated into the ranking. We do not know how many other States have begun to implement this practice. All of these actions by Federal agencies on non-Federal lands have the potential to help maintain and expand habitat for saltmarsh sparrow but more research is needed to quantify benefits to the species.

50 The Service also oversees the implementation of two large land conservation grants that could support the conservation of coastal marsh habitats. The National Coastal Wetlands Conservation Grant Program, for example, had $17 million in funding nationwide in 2019 and the North American Wetland Conservation Act (NAWCA) grant funds approximately $37 million annually and in 2018 supported projects in Virginia and Maine. Both of these national, competitive wetland grant programs have great potential for land protection and habitat restoration that may benefit saltmarsh sparrow. Over five fiscal years (2014-2018), NAWCA grants protected (in fee or easement) 157,661 acres and restored or enhanced 28,399 acres, at a cost (including grant, match, and federal or non-match funding) of $227,373,215. That equates to $1,222 per acre conserved.

3.7.4 Conservation Partnerships

Since 2016, the Atlantic Coast Joint Venture (ACJV) has increased its emphasis on coastal marsh habitats and adopted three flagship species--one being the saltmarsh sparrow--on which to focus their coordinated conservation efforts. As part of this initiative the ACJV recently completed (fall of 2019) a Salt Marsh Bird Conservation Plan for the Atlantic Coast (https:// www.acjv.org/documents/salt_marsh_bird_plan_final_web.pdf). They also worked with partners to develop complementary species action plans for Saltmarsh Sparrow and Black Rail, which will be published in 2020. The Saltmarsh Sparrow Conservation Plan builds on the Saltmarsh Bird Conservation Plan, including several implementation strategies specifically needed to address the threats to saltmarsh sparrow, as well as state-specific population and habitat objectives. Now that these documents are finalized, ACJV staff and partners are focused on implementing the conservation actions called for in the plan. The decision support tool (DST) that the ACJV developed prioritizes patches of salt marsh habitat (ranging in size from a few to thousands of acres) for conservation attention within the breeding range of the saltmarsh sparrow. The DST incorporates several variables in a model, including the distribution and abundance of saltmarsh sparrow based on 2011/2012 rangewide surveys (Weist et al. 2018, pp. 2–7). An initial version of the DST is available at this link: https://arcg.is/1e4rK8.

Additional planning tools have been developed by The Nature Conservancy and local conservation organizations, including the Partnership for the Delaware Estuary, the Barnegat Bay Partnership, and other non-governmental organizations. These include a Salt Marsh Assessment and Restoration Tool (SMART) developed to guide tidal marsh restoration efforts (Konisky 2012, entire), the Conservation Assessment and Prioritization System (CAPS) tool created by the University of Massachusetts with the support of the North Atlantic Landscape Conservation Cooperative (LCC) (https://www.umass.edu/landeco/research/caps/data/iei/iei.html), and the Coastal Resilience Evaluation and Siting Tool (CREST) https://resilientcoasts.org/#Home developed by the National Fish and Wildlife Foundation (NFWF) with NOAA and other partner organizations to help prioritize restoration actions.

Researchers recently conducted an analysis to identify priority marsh patches for tidal marsh bird conservation given the abundance of five tidal marsh specialists bird species (saltmarsh sparrow, seaside sparrow, Nelson’s sparrow, clapper rail (Rallus crepitans), and willet (Tringa semipalmata)) on a patch and its cost to conserve (Klingbeil et al. 2018, entire). They found the saltmarsh sparrow to be the most representative focal species to prioritize land for protection, as funding focused on protection and management of this species would also benefit the other species (Klingbeil et al. 2018, p. 880).

51 3.8 Summary of Factors Influencing Viability

The total amount of tidal marsh habitat across the range of the saltmarsh sparrow has declined over the past 200 years, primarily as a result of human population growth and development. Over the past 40 years these threats have continued and are exacerbated by accelerated sea-level rise which is currently outpacing the natural ability of most marshes to maintain their vegetative structure through disruption of the natural ecosystem processes of salt marsh development. Under current flooding frequencies across much of the species range, high marsh vegetation is rapidly converting to mudflat and low marsh habitats that are not suitable for saltmarsh sparrow nesting. Multiple stressors exacerbate marsh vulnerability to flooding, including development of the adjacent uplands, historic alterations to marsh hydrology, disruption of processes of sedimentation, increased nutrient inputs, and the presence of invasive species. The result has been the degradation and fragmentation of suitable nesting habitat and barriers to the natural creation of new marsh habitat along upland marsh zones.

Marsh flooding and the subsequent flooding of nests is the primary cause of reproductive failure across the range of the species and is closely correlated with dramatic declines in saltmarsh sparrow populations. Predation causes additional nest losses, which can exceed flooding during some years, particularly in the southern portion of the breeding range or where high marsh has been fragmented into smaller patches or pushed closer to the upland edge. There is also some evidence that methyl mercury concentrations in prey could be influencing reproductive success, particularly at more northern latitudes. At the southern end of their range competition with seaside sparrows may limit access to habitat. At the northern end of the range saltmarsh sparrows hybridize with Nelson’s sparrow, but the overall impacts of this are unknown and this is not currently considered a significant threat when compared to the other stressors. Both Nelson’s and seaside sparrows are more generalized in their habitat needs and utilize brackish or low marsh vegetation for nesting. As a result, they are not currently experiencing habitat loss at the same rate as saltmarsh sparrow.

Numerous agencies and organizations are engaged in salt marsh protection and restoration. Efforts have evolved over time and are beginning to focus more on approaches that will result in high marsh habitat that would provide suitable nesting habitat for saltmarsh sparrow. Many of these techniques have not been implemented at large geographic scales or have not been in place long enough to evaluate their effectiveness.

52 POPULATION NEEDS AND CURRENT CONDITION

4.1 Methods To evaluate and determine the species’ current condition, we used the wealth of literature published on the saltmarsh sparrow and its habitat over the past forty years. In order to determine the current condition of the saltmarsh sparrow, which occupies areas from Maine to Florida, we divided the species’ range into four analytical units (see 4.2 Analytical Units below). Within each analytical unit, we evaluated the species’ current condition by assessing its resiliency, redundancy, and representation after taking into consideration the threats that are impacting them. We evaluated the current condition of each analytical unit individually. By evaluating the current condition across each analytical unit, we were able to determine the overall current condition for the species.

We evaluated resiliency using a resiliency matrix designed by our team a priori, incorporating metrics identified as the most important factors influencing resiliency of the saltmarsh sparrow. For each metric, we created three threshold values representing high, moderate, or low resiliency. The metrics threshold values were identified by using the best information available including a combination of team member knowledge, consultation with species or habitat experts, and scientific literature reviews.

We chose to evaluate redundancy and representation by describing relevant information from peer-reviewed literature. For redundancy, we referred to data representing the number of individuals across the range of the species and the size of the species range. We qualitatively assessed how these data may make the species more or less at risk should a catastrophic event occur. We then addressed whether this may represent a lower or higher redundancy value. For representation, we referred to peer-reviewed literature on the saltmarsh sparrow’s ability to adapt to change and took into consideration any behavioral differences or other differences within each analytical unit into our analysis. We then qualitatively assigned a threshold value for representation.

4.2 Analytical Units The saltmarsh sparrow does not have distinct populations (Walsh et al. 2017a, p. 1247). Therefore, we used geographic characteristics (marsh size) and geographic-specific stressors (competition with other species, tidal regime, and rate of sea-level rise) in order to evaluate and take into consideration any potential variability across the species range and to more specifically evaluate any potential localized conditions. This resulted in the identification of three analytical units within the breeding range (Northern, Central, and Southern Analytical Units).. The wintering range does not appear to have geographic-specific stressors, and therefore was consolidated into one unit (Wintering Analytical Unit)(Fig. 15).

The Northern Analytical Unit extends from the northernmost known breeding site in Maine (South Thomaston; Hodgman et al. 2002, p. 39) south to the north shore of Cape Cod, Massachusetts. This area includes tidal marsh in Maine, New Hampshire, and Massachusetts and is characterized by a large tidal range and lower rates of marsh loss (Kennish 2001, pp.

53 742–743; Goodman et al. 2007, pp. 117–119) when compared to other portions of the range. Additional considerations in this unit include hybridization and potential interspecific competition with the Nelson’s sparrow (Walsh et al. 2016a, pp. 6–7). The Central Analytical Unit is located in between the Northern and Southern Breeding Analytical Units. The Central Analytical Unit continues from the south facing shores of Cape Cod, Massachusetts to the Long Island Sound and Peconic Bay shores of Long Island, New York. This unit contains tidal marshes on the mainland and islands of Massachusetts, Rhode Island, Connecticut, and New York. It is defined by the biotic and physical stressors of accelerated rates of sea-level rise, marsh loss and degradation, and extensive development in the landscape (Hartig et al. 2002, p. 79; Carlisle et al. 2005, p. 13; Smith 2015, pp. 132–134; Raposa et al. 2017a, p. 393; Watson et al. 2017b, p. 665). The Southern Analytical Unit incorporates marshes on the south shore of Long Island, New York continuing south along the Atlantic coast to the southern-most known breeding site in Virginia (Gloucester County; Greenlaw et al. 2018, p. 12). This unit contains tidal marshes in New York, New Jersey, Delaware, Maryland, and Virginia. The biotic and physical stressors in this unit are high rates of predation and interspecific competition with seaside sparrows (Ruskin et al. 2017a, p. 910; Roberts et al. 2017, p. 123) combined with high rates of habitat loss and conversion associated with extensive development in the landscape (Mitchell et al. 2017, entire).

The Wintering Analytical Unit ranges from North Carolina through Florida encompassing areas where saltmarsh sparrows most commonly winter but that do not support nesting. Saltmarsh sparrows will also overwinter from North Carolina to as far North as Massachusetts although it is less common further north except during warmer winters (Greenlaw et al. 2018, (Distribution, Migration and Habitat) unpaginated) (Figure 15).

54 Figure 15. Map of saltmarsh sparrow analytical units including core parts of the wintering range. Range polygons extend further inland than is accurate because tidal marshes are too narrow to view on a map of the entire range.

55 4.3 Resiliency Resiliency is defined as a species’ ability to withstand stochastic events. We analyzed saltmarsh sparrow resiliency at the level of individuals and breeding populations. Wintering population data were incorporated when known and applicable. Resiliency metrics evaluated include measures of habitat quantity and quality; demographic metrics (adult survival and juvenile survival and fecundity) and those that affect fecundity (predation and tidal flooding); and metrics designed to assess the degree of human alteration on the landscape (Table 3).

More specifically, we summarized the total amount of saltmarsh that is available in each analytical unit to assess habitat quantity and then evaluated individual marsh patch size and nesting habitat availability to evaluate habitat quality across each analytical unit.

There are multiple peer-reviewed papers published by the SHARP team that calculated overall population growth, adult survival, nest success, nest predation, and nest flooding rates across 21 marsh locations throughout the majority of the saltmarsh sparrow breeding range, which were used in this evaluation. Although these data are mostly based on a relatively short window of time (2012–2017), the data were collected across the range of the species and are presumed to be representative of the population as a whole.

Human alterations to the landscape were included to reflect the degree to which the saltmarshes in the different analytical units are experiencing additional stressors that will impact their ability to persist into the future. Each of these metrics and our assessment of what would constitute a high, moderate or low resilience score for them are described in Table 3.

Table 3. Resiliency table for saltmarsh sparrow. Confidence in measurement and/or High Resiliency Moderate Low Confidence threshold** Citations Resiliency Resiliency in metric*

Habitat Quantity and Quality

Total marsh Increasing in Stable amount of Experiencing High Moderate habitat availability across habitat across the declines across the availability the range and range of the species range and through through time and through time time

Low; hard to Benoit and tell what Askins Individual > 55 ha 10 – 55 ha < 10 ha habitat size Low 2002; Marsh Sizes the species SHARP historically team pers. had comm. 11 Dec. 2017

56 Meiman and Elphick Proportion Greater than 40 Between 40 and 20 Very little high High Low - 2012; and percent high marsh percent high marsh marsh (less than 20 Moderate DiQuinzio et Availability of percent) al. 2002; High Marsh Gjerdrum et al. 2005; Shriver et al. 2007

Unable to assess this metric SHARP Proximity to High Low personal next marsh communicati on

Demographics

Population Stable to Declining (<0) High High N/A Growth Rate Increasing (>0)

Adult annual Greater than 50 Between 35 and 50 Less than 35 Moderate High Field et al. survival percent adult percent adult annual percent adult 2018; annual survival survival annual survival DiQuinzio et al. 2001; Winder et al. 2012b; Post & Greenlaw 1982

Juvenile Greater than 25 Between 25 and 10 Less than 10 Moderate Low DiQuinzio et annual percent juvenile percent juvenile percent juvenile al. 2001; survival annual survival annual survival annual survival Post & Greenlaw 1982

Nest success Each female Each female Each female High High Post & produces 2–5 produces 1–3 produces 0–2 Greenlaw chicks/season chicks/season chicks/season 1982; and/or >45 percent and/or 35 to 45 and/or <35 percent DeRagon nest success percent nest success nest success 1988; DiQuinzio et al. 2002; Gjerdrum et al. 2005; Greenburg et al. 2006; Shriver et al. 2007; Roberts et al. 2017, Ruskin et al. 2017

57 Flooding regime Flooding regime Flooding regime provides >75 provides 50 provides <50 Ruskin et al. percent of nests percent

Alterations

Degree of Natural marsh free Minimal history of Extensive historic Moderate Low Bourn and human of historic degradation human caused Cottam alteration degradation degradation that 1950; might include tidal Vincent et restrictions, al. 2013; ditching, nutrient Raposa et al input, that has 2017a & b; made the marsh etc. more susceptible to future threats

Development Less than 15 Between 15 and 40 Greater than 40 Moderate Low - Mitchell et within 1000m percent percent percent developed Moderate al. 2017, of marsh developmen development p.8; Wiest patch t et al 2018 landward edge

Tidal No tidal # tidal restrictions High Moderate Correll et al. Restrictions restrictions > 0 2017a

Degree of Unable to quantify the degree of shoreline hardening in each Moderate Mitchell et shoreline analytical unit for this assessment al. 2017 hardening *Confidence in Metric - constitutes the SSA teams assessment of how much evidence supports this metric as one which is important to the resiliency of the species; **Confidence in measurement and / or threshold - constitutes the SSA teams assessment of how confident they are that the thresholds we established between High/Moderate/Low resiliency rankings are well established in the literature and / or applicable for the species; ***Use of 23 days represents the minimum amount of time considered necessary to potentially fledge at least one chick. 58 4.3.1 Habitat Quantity and Quality

4.3.1.1 Total Salt Marsh Availability (Quantity) Across the breeding range of the saltmarsh sparrow there is approximately 277,711 ha (686,236 ac) of tidal marsh habitat. Divided among the three analytical units in the breeding range, 87 percent of marsh habitat falls within the Southern Analytical Unit, while 9 percent is in the Northern Analytical Unit, and 4 percent of the habitat is in the Central Analytical Unit (12,328 ha (30,488 ac) (Fig. 16). As described in section 3.1, habitat has been reduced and fragmented, even though the range of the species has been stable. During the winter months, the majority of the species migrates to marshes from Virginia through Florida, with individuals continuing to utilize marshes as far north as Massachusetts during mild winters. There is extensive literature documenting the decline in marsh habitat across the range of the species so we assigned low resiliency score to all analytical units for this metric.

Figure 16. Total marsh area (ha), and percentage available by breeding analytical unit based on land cover data (spatial data layer from Wiest et al. 2019 used to calculate values).

4.3.1.2 Salt Marsh Size (Quality) The majority of the literature on saltmarsh sparrow habitat has been collected throughout the breeding range. The smallest salt marsh size with breeding saltmarsh sparrows detected during a study in Connecticut was 10 ha (24.7 ac) (Benoit and Askins 2002, p. 317), although there are records of nesting at smaller sites in New York (Kocek 2016, p. 57-58, 81). Across the range, marsh size is positively correlated with saltmarsh sparrow presence (Shriver and Vickery 2001, 59 p. 6; Benoit and Askins 2002, p. 317; Meiman et al. 2012, p. 733). Saltmarsh sparrow occurrence in Maine is higher in larger more connected marshes and sites within 1 km (0.6 mi) of another marsh are preferred (Shriver et al. 2004, p. 550).

Saltmarsh sparrow abundance is also higher in more open patches (i.e., not surrounded by trees/buildings). Marsh area tends to be correlated with openness, but angle to the maximum horizon is a better indicator of openness (i.e. an angle of 0 degrees to the nearest object on the horizon indicates a very open marsh) (Marshall 2017, p. 4). Abundance of saltmarsh sparrows is greatest between 0 and 13 degrees to the maximum horizon, suggesting that marshes that are surrounded by fields/low shrubs are preferred over those bordered by trees/buildings (Marshall 2017, p. 12). To assess marsh size across the breeding analytical units, we estimated individual marsh patch size using a wetland cover/community type Geographical Information System (GIS) layer (Wiest et al. 2019, entire). Patches were defined using the National Wetlands Inventory (NWI) estuarine emergent marsh (NWI code: E2EM) polygons with a 50 m (125 ft) buffer. Due to the inherent error in use of aerial imagery; patch buffers that intersect each other were considered the same patch (see Wiest et al. 2019, entire, for a full description of the data and its limitations). We evaluated each analytical unit and determined the proportion of salt marsh patches that fit the criteria of high, medium, or low resiliency (per Table 3), and gave an overall assessment for each analytical unit by selecting the metric associated with the highest percentage of habitat. For each analytical unit, 76 to 85 percent of the patches were smaller than 10 ha (25 ac) and classified as having low resiliency, while 10 to 16 percent were moderate resiliency, and 4 to 8 percent of the marshes were classified as high resiliency (Fig. 17). Despite the fact that 85 percent of the marsh patches are smaller than 10 ha (25 ac) in the Southern Analytical Unit, the median marsh size is 49.8 ha (123.0 ac) demonstrating that there are some much larger marshes. This contrasts with the Central Analytical Unit where the median marsh size is 10.38 ha (25.65 ac). The Northern Analytical Unit is intermediate with a median marsh size of 27.21 ha (67.24 ac).

60 Figure 17: Marsh Resiliency Based on Size - Proportion of marsh patches within different size categories assigned scores of high (>55 ha), moderate (10-55 ha) or low (<10 ha).

Wintering range: As described in section 2.5.2 saltmarsh sparrow utilize salt marsh habitats throughout the wintering range for foraging and roosting. We believe that the species is likely most limited at this time of year by the amount and distribution of tall vegetation within the marshes. This provides critical habitat for perching above water levels in the marsh during flooding associated with high tides, rain events, or tropical storms and hurricanes. There is evidence that the birds prefer to roost within the marsh and will avoid the adjacent upland. As such, an assessment of the distribution of tall marsh vegetation would help to quantify the quality of habitat that is currently available within the wintering range, but we did not have these data available to us. We do not believe that food or other resources are limited within the wintering range.

4.3.1.3 Proportion, Availability and Quality of High Marsh (Quality)

Although saltmarsh sparrows utilize all parts of the marsh for different aspects of their life history, the quantity of high marsh habitat is the best assessment of suitable nesting habitat for the species (Meiman and Elphick 2012, p. 857). In the breeding range, stem density and mean percent cover of Spartina patens and Juncus gerardii are higher on marsh plots that contain saltmarsh sparrow adults and nests than on marsh plots that do not contain individuals or nests (Meiman et al. 2012, p. 732). For plots with saltmarsh sparrow nests, the amount of high marsh cover was higher than plots without nests (Gjerdrum et al. 2005, p. 854; Meiman et al. 2012, p. 732). In habitats bordering Long Island Sound, saltmarsh sparrow presence was positively correlated with the amount of native vegetation (Shriver et al. 2004, p. 550), with vegetative 61 structure being the best predictor of suitable nesting habitat at the local scale, and the area of high marsh at a site being the most reliable predictor of suitable nesting habitat at a regional scale (Meiman and Elphick 2012, p. 857). This research substantiates the notion that availability of high marsh habitat will limit the distribution of breeding females.

Throughout the breeding range there was approximately 79,760 ha (197,091 ac) of high marsh habitat in 2013 (Wiest et al. 2019, spatial data), which represented approximately 29 percent of the total marsh area being available for nesting on average. The majority of high marsh habitat (77 percent; Fig. 18) occurs within the Southern Analytical Unit. The Northern Analytical Unit, however, hosts the largest proportion of high marsh per hectare (ac) of marsh, and provides more nesting habitat on a per hectare (ac) basis (59 percent; Fig. 19).

Figure 18. The total amount of high marsh estimated to occur within the breeding range (79,760 ha) which is available within each breeding analytical unit and the percentage of all high marsh which this represents (data from 2013).

62 Figure 19. The proportion of marsh that is high marsh in each breeding analytical unit.

To assign a resiliency score for habitat quality across the breeding range, we calculated the amount of high marsh in each marsh patch for each analytical unit of the breeding range using the Wiest et al. (2019, entire) GIS layer. Any marsh patch that contained ≥40 percent high marsh was considered to have high resiliency, while a patch with >20 percent but <40 percent high marsh has moderate resiliency and a marsh with <20 percent high marsh has low resiliency (Table 3). The Northern Analytical Unit had the highest proportion of marsh patches that were considered to have high resiliency (42.7 percent) with the proportion of highly resilient marsh patches declining through the Central and Southern Analytical Units. Across the breeding range the proportion of marshes that had moderate resiliency with regard to the amount of high marsh remained relatively constant (10 to 13.7 percent). Low resilience marshes were more common in the Southern Analytical Unit (68.9 percent of the marsh patches) and became less common through the Central and Northern Analytical Units (Fig. 20).

63 Figure 20. The proportion of marsh patches in each analytical unit that were classified as having low, moderate or high resiliency based on our habitat quality metric. Habitat quality was defined as the proportion of high marsh in the patch.

4.3.2 Population Size, Growth Rate and Demographics

4.3.2.1 Abundance and Distribution

Abundance. The total population of saltmarsh sparrows declined by 75 percent between 1998 and 2012 (Greenlaw et al. 2018, (Demography and Populations) unpaginated) based on a population estimate of 53,000 individuals (95 percent confidence interval (CI): 37,000–69,000; Wiest et al. 2016, p. 281). An updated analysis using a more detailed, spatially explicit model calculated a similar, range-wide population size estimate of 60,000 individuals, 95 percent CI: 40,000–80,000) (Wiest et al. 2019, p. 117). Both of these global population estimates (Wiest et al. 2016 and Wiest et al. 2019) used data collected in the breeding seasons (April – July) of 2011 and 2012 at 1,780 survey points throughout the majority of the species’ breeding range (excluding only the western Chesapeake Bay; Wiest et al. 2016, pp. 275-277). This updated number of 60,000 individuals (95 percent CI: 40,000-80,000; Wiest et al. 2019, p. 117) is the best available data value, and is what we use for our estimates of current population size.

A comparison between surveys conducted in the 1990s, 2011, and 2012 determined that the population is exhibiting a negative growth rate across the breeding range. Using an analysis of 14 datasets from across the breeding range, including 3,195 survey points and >170,000 individual observations, Correll et al. (2017) found that the population declined an average of 9 64 percent per year across the range from 1998 to 2012, with some portions experiencing even higher rates of decline (e.g. 12.2 percent per year in New England; Correll et al. 2017, p. 177). Applying the 9 percent rate of decline to the mean of the most recent population estimate of 60,000 with a 95 percent CI of 40,000-80,000, we estimate that the current population size (2020) of the saltmarsh sparrow is 28,215 individuals (95 percent CI = 22,715–45,430). This represents a decline of 87 percent (79–89 percent range) from 1998 levels which were estimated at 212,000 individuals. Distribution. During the 2011 and 2012 surveys, the majority of the breeding population (78 percent or 46,800 individuals) were nesting in the Southern Analytical Unit , while only 13 percent (7,800 individuals) nested in the Northern Analytical Unit and 9 percent (5,400 individuals) were nesting in the Central Analytical Unit (Figure 21; Wiest et al. 2016, entire). Assuming that nesting individuals are currently distributed across the landscape as they were in 2012, the current population estimate for each analytical unit (2018) is 4,429 individuals in the Northern Analytical Unit, 3,066 individuals in the Central Analytical Unit, and 26,576 individuals in the Southern Analytical Unit.

Figure 21. Saltmarsh sparrow population distribution across the breeding range based on data collected in 2011–2012 (Wiest et al. 2016, p. 281).

Since the majority of the salt marsh habitat and the largest proportion of the population are in the Southern Analytical Unit, we were interested in understanding the density of nesting individuals throughout their range. Point count surveys conducted in 2011 and 2012 indicate that the mean density of nesting saltmarsh sparrow across the breeding range is 0.17 birds/ha (0.42/ac) (standard error (SE) = 0.02) with the lowest densities in Maine (0.02 birds/ha (0.05/ac) (SE = 0.01) and the highest densities in New York (0.31 birds/ha (0.76/ac), SE= 0.07). All other areas within the breeding range have an intermediate density of birds but generally marshes from New Jersey to the north have more birds/ha (birds/ac) of marsh than the three southern states. Although this could be explained by the variability in the proportion of high marsh across the range, this is currently the best assessment of density available (Figure 22) (Wiest 2016, entire). 65 Figure 22. Estimated percentage of the saltmarsh sparrow population as it compares to the percentage of total marsh area and the percentage of high marsh area by state (North to South) as an indication of rough density estimates. Based on surveys of 1,780 point locations in 6,917 saltmarsh patches conducted in 2011 and 2012 (from Wiest et al. 2016, p. 281). *Estimates for ME missing some salt marsh patches and do not include unidentified sharp-tailed sparrows (i.e. sparrows that could not be identified to species (Nelson’s vs. Saltmarsh)). **Estimates for MD and VA only include marshes east of the Chesapeake Bay.

4.3.2.2 Range Contraction

In addition to an estimated annual population decline since at least the late 1990s (Correll et al. 2017, p.177), the saltmarsh sparrow has experienced range contraction at the southern end of its breeding range. Historically, breeding extended onto the western shore of the Chesapeake Bay (Virginia and Maryland), but by the 1980s saltmarsh sparrows were absent from several sites with only one recorded nest in 1992 along the western shore. By the early 2000’s nesting had ceased at some sites on the eastern shore as well (Greenlaw et al. 2018, (Distribution, Migration, and Habitat) unpaginated). At the northern extent of the species’ range, data suggests some range expansion may have occurred in Maine within the hybrid zone, but the potential for misidentification between saltmarsh, Nelson’s, and hybrid sparrows at the date of this historical inference makes this expansion uncertain (Greenlaw 1993, 298–300; Hodgman et al. 2002. pp. 42–43). Further, throughout the northern portion of the range saltmarsh sparrow are declining

66 relative to Nelson’s sparrows at 66 percent of the sites monitored where the two species overlap (Shriver et al. 2016, pp. 193-195; Walsh et al. 2017b, p. 462; Correll et al. 2016, p. 7). Thus the overall effect appears to be a contraction of the range at both ends of the breeding range. Current wintering range extent relative to what has been documented in the literature is unknown, but based on the available literature, we expect it is similar to the current extent.

4.3.2.3 Demographics The most comprehensive study of saltmarsh sparrow survival and reproduction was conducted at 23 breeding sites across seven states from 2011 to 2014 by the SHARP team in collaboration with many partners. These 23 sites were distributed across the three breeding analytical units (Figure 23). Two of the sites had small samples sizes and were not used in the published literature. As such, we also used the data from 21 sites to evaluate each analytical unit with regard to annual survival, fecundity, and population growth rates.

Figure 23. Map for SHARP demographic plot locations. Note, the analytical unit polygons extend further inland than is accurate because tidal marshes are too narrow to view on a map of this spatial extent.

67 Growth rates - Of the 21 demographic sites monitored between 2010 and 2014, only five sites demonstrate stable to positive growth rates (≥ 0) while all other sites (76 percent) were declining (<0) (Figure 24). Three of the sites with growth rates that were stable or increasing were in the Northern Analytical Unit and two were in the Southern Analytical Unit along the south shore of Long Island, New York. All of the sites within the Central Analytical Unit exhibited negative growth rates. Although there is variability between sites and between years, the median growth rate was -0.14, indicating that the overall population trend across all sites is declining (Field et al. 2018, p. 977). Based on these demographic data as well as estimates presented previously of a negative growth rate of -9 percent per year between 1998 and 2012 (Fig. 25; Correll et al. 2017, p. 177), we have determined that the population overall has low resiliency for all analytical units.

Figure 24. Annual population growth rate at 21 demographic sites from throughout the range of saltmarsh sparrow from 2010 to 2014. Values above zero indicate an increasing population trend and those below zero a declining population. Vertical lines are the 95% credible intervals. White dots are sites within the hybrid zone (From Field et al. 2018, p. 977).

68 Figure 25. Saltmarsh sparrow relative abundance estimated between 1998 and 2012 indicating a 9 percent negative growth rate overall between 1998 and 2012 (From Correll et al. 2017, p.178).

Adult Survival - Based on data from the 21 demographic sites, surveyed from 2010-2014, the mean adult apparent survival rates were 0.44 (95 percent CI: 0.37–0.52) and 0.49 (95 percent CI: 0.42–0.56) for females and males, respectively (Field et al. 2018, p. 976). This study found site-level variation in adult survival, with Maine having the highest apparent survival rates and New Jersey having the lowest apparent adult survival (see Field et al. 2018 figure 4, p. 977). Variation between sites was not influenced by large-scale landscape variables such as latitude or marsh size. Although there were no differences among years, the timespan of the study was short, limiting the scope to assess temporal variation in survival. However, similar, but slightly lower, average adult survival was observed (0.40; range 0.266–0.658)) between 1996 and 1998 in Rhode Island (DiQuinzio et al., 2001, p. 891). Adult survival has declined when compared to historic rates recorded from banding studies conducted in New York from 1967 to 1972 (n=11 females and 73 males) and again from 1976 to 1977 (n=66 females and 106 males) at the same marsh. Based on re-sighting of marked birds through 1980, the annual survival rates averaged 0.63 and 0.53 for females, and 0.60 and 0.55 for males in the 1967-72 and 1976-77 cohorts respectively (Post and Greenlaw 1982, p. 105).

Annual survival has also been explored on the wintering range. Annual adult survival of birds captured at wintering sites was estimated at 0.52 (± 0.12) in North Carolina from 2006 to 2010 (Winder et al. 2012b, p. 427). Additionally, weekly survival estimates from sites on the breeding (Connecticut) and wintering (South Carolina) grounds from 2010 to 2013 were not different and were sufficiently high (>0.999 for both seasons, with no differences between sexes; Borowske 2015, p. 103). This indicates that adult survival is relatively consistent within the

69 breeding and wintering periods and suggests that most adult mortality currently occurs during migratory periods (Borowske 2015, p. 103). To assess the resiliency of adult survival across the range we utilized survival data from Field et al. (2017b, entire) and assigned a high resiliency score to sites with greater than 50 percent apparent adult survival, a moderate score to sites with 35 to 50 percent apparent adult survival and low resiliency to sites with less than 35 percent apparent adult survival. We selected the 50 percent upper limit because five of the six sites with that level of apparent adult survival also had an increasing population trend. We do not know how much of the observed adult survival is related to immigration or emigration, so there is some uncertainty as to whether this level of survival would allow persistence following a catastrophic event. In the Northern Analytical Unit, three of the seven sites had apparent adult survival above 50 percent and those three also had increasing population trend, although the mean apparent adult survival across sites was 0.49. None of the sites in the Central Analytical Unit had apparent adult survival above 50 percent (mean = 0.39) and all populations were declining. In the Southern Analytical Unit, three of the seven sites had apparent adult survival above 50 percent, but only two of the sites had an increasing population trend and the mean for all sites was 0.44 (Fig. 26) Based on this assessment, the Northern Analytical Unit is more resilient than the other units, and the Central Analytical Unit has the lowest resiliency. Since each analytical unit has a mean apparent adult survival between 0.35 and 0.49, we determined that overall resiliency for apparent adult survival is moderate.

70 Figure 26. Saltmarsh sparrow adult mean annual survival estimates at demographic sites throughout the range (Field et al. 2018, entire). Green = High Resilience (>50 percent); Yellow = Moderate Resilience (35–50 percent); Red = Low Resilience (<35 percent).

Juvenile Survival - Little information is available for first-year survival of saltmarsh sparrows. At a site in Rhode Island, researchers found the estimated mean apparent juvenile survival rate was found to be 0.14 (95 percent CI: 0.10–0.19), which was less than half the estimate for adults (DiQuinzio et al. 2001, p. 892). Only 11 percent of juveniles banded at this site were found to return in subsequent years. Juvenile recapture rates before September of the same year (fledgling survival) were 0.36 (Post and Greenlaw 1982, p. 105). Rates are identified as apparent since there were not comprehensive surveys in surrounding marshes to account for immigration and emigration. Recent publications do not report juvenile survival for demographic sites so we were not able to assign a resiliency score to any of the analytical units independently or for juvenile survival overall.

71 Nest Survival and Fecundity - Saltmarsh sparrow nest survival/success rates have been extensively studied with six early studies documenting relatively consistent 27 to 44 percent nest success between 1977 and 1982 in New York (Post and Greenlaw 1982, pp. 104–105) and Rhode Island (DeRagon 1988, 61-63) respectively, nest success rates of 27 percent during 1993 to 1998 in Rhode Island (DiQuinzio et al. 2002, p. 182) and 27 to 36 percent between 2002 and 2003 in Connecticut and Maryland (Gjerdrum et al. 2005, entire; Greenberg et al. 2006). This compares to a study of 191 nests in Connecticut between 2007 and 2009 documented 18 percent of nests successfully fledging young (Bayard and Elphick 2011, p. 397). On marshes in Maine during 2012-2013 daily nest survival rates (0.94 to 0.95) recorded are consistent with nest success between 27-36 percent (Ruskin et al. 2015, p.1643). This compares to 53 percent nest success recorded in Maine between 1998-2001 (Shriver et al. 2007, entire).

The most recent comprehensive summary of nest survival/success utilized data from 23 sites (837 nests), encompassing approximately 59 percent of the breeding range from 2011 through 2013 (Ruskin et al. 2017a, entire; Ruskin et al. 2017b, entire) (see Fig. 23). Seasonal fecundity is relatively consistent within a site, with some variability between years and considerable variability between sites as little as 1 km (0.6 mi) apart (Ruskin et al. 2017b, p. 298). Seasonal fecundity estimates across plots range from 0.09 (95 % confidence interval: 0.05–0.13) to 0.78 (0.61–0.94) successful broods per female each season. The mean seasonal fecundity (±sd) across all populations and years was 0.46 ± 0.17 successful broods per female each season (Ruskin et al. 2017b p. 297).

Given that the mean brood size (number of chick hatched per nest) is 2.73 chicks across the range, this equates to 0.92 to 2.04 fledged chicks per female per year in the Northern Analytical Unit, 0.44 to 1.88 in the Central Analytical Unit, and 1.31 to 1.64 in the Southern Analytical Unit (Ruskin et al. unpublished). These are all lower than historical data which estimated annual reproductive success to range from 2.65 to 5.25 fledged young per female (Post and Greenlaw 1982, p. 104).

To assess resiliency with regard to reproductive success we used the number of successful nests per female per year (mean annual fecundity). Based on our literature review, we established a mean annual fecundity greater than 45 percent at a site as a highly resilient site, fecundity estimates of less than 35 percent as low and intermediate values as moderate. At the 21 sites with detailed nesting data (as summarized by Ruskin et al. 2017b; Field et al. 2018), annual fecundity averaged 0.49 (0.34–0.75) in the Northern Analytical Unit, 0.39 (0.16– 0.69) in the Central Analytical Unit, and 0.44 (0.36–0.55) in the Southern Analytical Unit (Figure 27). Based on the average values, the Northern Analytical Unit is considered to have had high resiliency overall while the Central and Southern Analytical Units have moderate resiliency (Fig. 27). We base our conclusion on the available literature with regard to historic nest success data and the best available current information.

72 Figure 27. Estimates of mean annual fecundity (successful nests/year/female) from data collected at 21 marshes throughout the range of saltmarsh sparrow 2010–2014 (Field et al. 2018, entire). Green = High Resilience (>45 percent); Yellow = Moderate Resilience (35–45 percent); Red = Low Resilience (<35 percent).

Although we assigned an overall measure of resilience for each analytical unit based on the mean annual fecundity, we were also interested in assessing the relative impacts of the two primary factors that influence nest success: flooding and predation. We were interested in quantifying the variation in these two factors at sites across the breeding range of the species.

Marsh flooding and predation: Our team decided a priori upon marsh flooding and predation thresholds for high, medium, and low resiliency (Table 3). Although these values are related to fecundity, a variable already included in the resiliency table, we wanted to separately address causes of nest failure (nest flooding and predation) because we believe they are the main drivers of population decline based on the best available literature. We were unable to acquire data that fit our resiliency table’s definitions of high, medium, and low for marsh flooding. We do have values available to evaluate our predation assessment, but given the compensatory nature of nest

73 failure due to flooding or predation, we decided it would be inappropriate to assign separate resiliency values for each category. For example, sites that have high predation values often have low flooding values and vice versa. Further, flooding and predation values are highly variable among years, making it inappropriate to assign resiliency values for a dataset spanning four years. Rather, we present the data for both conditions within each analytical unit graphically without assigning high, medium, or low resiliency categories to each (Figure 28).

The data used represent the best available data that illustrate the impacts of marsh flooding and predation on saltmarsh sparrows in each analytical unit. Both sets of data were collected by the SHARP team. The data on percentages of nests failed from flooding derive from 22 sites across the species’ range from 2011 to 2015 (Fig. 28). The second type of data (daily probabilities of nest failure due to flooding) use the same data but from 2011 to 2013 at 22 sites and are taken from a published manuscript (Ruskin et al. 2017b, Appendix).

In the Northern Analytical Unit, the SHARP team regularly monitored nesting at eight sites. Of 1,298 nests monitored from 2011-2015, approximately 33 percent failed due to flooding and 19 percent failed due to predation (SHARP unpublished). In the Central Analytical Unit, across 7 regularly monitored sites (n = 356 nests), approximately 35 percent of nests failed from flooding and 9 percent failed due to predation (SHARP unpublished). In the Southern Analytical Unit, across 8 regularly monitored sites (n= 571 nests), approximately 21 percent of nests failed from flooding and 22 percent failed from predation (SHARP unpublished data) (Figure 28).

Figure 28. Proportion of nest failure by causes from 2011–2015 at 21 demographic sites.

74 4.3.3 Human Alterations to Habitat: In our review of the factors influencing the viability of the species (Section 3.1.2–3.1.3) we identified the loss and degradation of habitat due to human alterations to the marsh directly and to the surrounding landscape as one of the primary factors. We identified and quantitatively assessed four metrics to determine the resilience of tidal marsh habitat related to human alterations (Table 3). These include: (1) within-marsh alterations; (2) development within 1,000 m (3,281 ft) of each marsh patch; (3) tidal restrictions; and (4), the degree of shoreline hardening. For the first metric we discuss the available literature but were not able to complete our own qualitative assessment for each analytical unit. For the next two metrics we summarize the data that are available to us to come up with resiliency scores as identified a priori (Table 3). For the last metric, we ultimately were unable to find sufficient information to assess the current extent of shoreline hardening across the range of the species to determine what proportion of marshes are being impacted at this time and therefore we did not assign resiliency scores for this metric.

4.3.3.1 Degree of Human Alteration (within-marsh alterations)

Over 90 percent of tidal marshes in New England have experienced grid ditching (Bourn and Cottam 1950, p. 1) and at least 50 percent of the marshes in the mid-Atlantic were impounded or have had their hydrology altered since 1931 (Smith et al. 2017, p. 37). These activities have resulted in numerous adverse impacts to the health of tidal marsh habitats as described in Section 3.1 (above). As such it is reasonable to classify the majority of all marshes in the four analytical units as having low resiliency based on this metric.

4.3.3.2 Urban and Suburban Development

To quantify the degree of development in the landscape surrounding marsh patches, we followed methods developed by Weist et al. (2018, entire), which utilize the NatureServe land use cover categories and National Land Cover Database (NLCD) developed classes. We then summarized the amount of land classified as urban, suburban or agricultural within 1,000 m (3,280 ft) of each marsh patch to align with models designed by Wiest et al. (2019, p. 114). Based on our review of the literature, we considered marsh patches with less than 15 percent in any of these cover classes within 1,000 m (3,280 ft) of the marsh edge as highly resilient while those with 15– 40 percent coverage were classified as having moderate resilience and those with greater than 40 percent coverage in these three cover classes considered to have low resilience. Although we found limited research that quantified specific thresholds of development that would impact marsh health and persistence, Mitchell et al. (2017, p. 8-9) found significantly more marsh loss in areas with greater than 15% development within 1,500 m. of the marsh edge and generally there is a high degree of certainty that there is a direct inverse correlation between development and marsh persistence (Section 3.1). Marsh habitat not constrained by developed areas maintains its capability to withstand changes from natural ecological processes and allows marshes to migrate upslope as the downslope conditions become unfavorable or change. Of the three classes used, there is the least amount of certainty as to the relative detrimental or beneficial impacts of agricultural land in the surrounding landscape. Conversion from native vegetative cover to agricultural land can result in heightened nutrient concentrations in the ground-water and lowering of the ground water table from irrigation, while also potentially increasing the presence of predators on the landscape.

75 We also recognize that there could be the potential for positive influences such as absorption and filtration of rainwater during storm events and the ability of marsh transgression into these agricultural areas with sea-level rise.. However, due to these aspects being uncertain and dealing with future conditions, we decided that for our current condition assessment, we would define all three cover classes as development given that it is no longer native habitat.

The Northern Analytical Unit has the highest proportion of marsh patches ranked as having high resilience (37.2 percent) based on our assessment. However, the Southern Analytical Unit has the highest proportion of area ranked as having high resilience given the larger average marsh patches size in the unit (Figures 29 and 30).

Figure 29. Proportion of patches (regardless of patch size) that were scored as high, moderate or low resilience based on the amount of development within 1000 m. of the patch

76 Figure 30. The proportion of all marsh hectares within each analytical unit that are defined as having high, moderate or low resiliency. Rather than assessing the number of patches, this assessment recognizes the variation in patch size.

4.3.3.3 Tidal Restrictions

To assess the relative resilience of marshes across the range with regard to the presence of tidal restrictions we utilized data from researchers who evaluated tidal restrictions across the Northeast (McGarigal et al. 2017, pp. 2–3). For our assessment, we summarized the number of marsh patches in each analytical unit that had one or more tidal restrictions. We categorize those marsh patches as having low resilience. We categorized any marsh patches that were free from restrictions as having high resilience with regard to this metric. Overall the highest proportion of tidally restricted marshes are in the southern analytical unit, and since these are larger marshes on average, the greatest amount of marsh habitat overall is restricted in this portion of the saltmarsh sparrow range. The fewest patches with restrictions are in the Central Analytical Unit, but overall between 19 and 39 percent of the marshes across the breeding range of the saltmarsh sparrow have been degraded due to tidal restrictions of some kind (Figure 31). A further evaluation of the degree of restrictions within the affected marsh patches assigned scores of >0.6 on a scale of 0 (no restriction) to 1 (most severely restricted). Results showed that the Southern Analytical Unit has the highest proportion of patches considered severely impacted (16 percent

77 of all patches), followed by the Central and Northern Analytical Units which had 9 percent and 8 percent considered severely affected, respectively. Based on both of these assessments, we determined that the Northern and Central Analytical units have moderate resiliency while the Southern Analytical Unit has low resiliency.

Figure 31: The proportion of marshes within the three breeding analytical units that still have a tidal restriction. Only sites with zero restrictions are considered highly resilient, while sites with one or more restrictions are considered to have low resilience.

4.4 Summary of the 3 R’s - Current 4.4.1 Redundancy

Redundancy is defined as the ability for a species to withstand catastrophic events, which is directly correlated with their distribution on the landscape and measures of connectivity.

The breeding range for saltmarsh sparrow extends along the Atlantic coast from Maine to Maryland (579 km (360 mi) (Wiest et al. 2016, entire) with the species limited to the thin band of tidal marsh habitat that exists in disjunct portions of this range. Southern marshes in Virginia and parts of Maryland that historically supported nesting are no longer considered viable and do not support nesting. At the northern end of the range, saltmarsh sparrow are declining relative to Nelson’s sparrows at 66 percent of the sites monitored where the two species overlap (Walsh et al. 2017, p. 462). These both indicate a contraction at both ends of the species range, although it has not been quantified overall. Individuals are also not evenly distributed across the breeding range. Approximately one third of the entire population breeds in coastal New Jersey, and according to recent estimates approximately 78 percent of the breeding population is

78 concentrated in marshes of the mid-Atlantic states (New Jersey to Maryland). It is conceivable then that a single event in that geographic area would have the potential to impact a significant proportion of the population.

Suitable nesting habitat is distributed across the landscape roughly proportional to the distribution of nesting birds, with roughly 77 percent of the habitat in the Southern Analytical Unit. More northern marshes tend to be smaller overall and more fragmented but support higher densities of high marsh as well as nesting females per hectare. So, although the species still occupies a majority of its historical range, the number of individuals (especially in the northern and central portions of its breeding range) has been significantly reduced.

There is no evidence that the distribution of individuals within the wintering range has changed significantly through time, but the species is primarily restricted to the narrow band of salt marsh that occurs along the coast of the south-eastern United States between Virginia and Florida. These marshes are also becoming more fragmented as sea level rises and the amount of habitat available for roosting is expected to become more limited. Tropical storms and hurricanes during the nonbreeding season have the potential to impact a significant proportion of the population as they are concentrated during this time. As a result we conclude that redundancy for the saltmarsh sparrow is moderate.

4.4.2 Representation

Representation is defined as the ability for a species to adapt to change. A principal component of a species’ ability to adapt is how flexible its life strategy is.

The saltmarsh sparrow is considered the most specialized of tidal marsh specialists (Correll et al. 2016, p. 6), occupying a narrow band of tidal marsh habitat found along the east coast of the United States (Greenlaw et al. 2018, (Distribution, Migration, Habitat) unpaginated). Within that habitat they are further specialized in their association with high marsh vegetation for nesting. The species is more of a generalist when it comes to their diet (Post and Greenlaw 2006), and they can presumably tolerate a range of temperatures given the latitudinal range of nesting (Conway et al. unpublished).

Currently the most rapidly changing conditions within tidal marshes is the rate of sea-level rise and the increased frequency and duration of tidal flooding in marshes. This has resulted in direct loss of eggs and chicks, as well as a loss of nesting habitat. There is some evidence that saltmarsh sparrows will alter their nesting location in response to flooding such that nests are placed higher in the vegetation or closer to the upland (see Nest Success above). Alternately, when nests are lost to predation, females will renest at sites further from the upland or lower in the vegetation (see Nest Success above). In all of these instances, nesting is still restricted to some portion of what would be defined as the high marsh zone within the marsh. Saltmarsh sparrow are also known to continue to initiate nests later in the season than other species, although there has not been an assessment of the relative survival of chicks fledged later in the season (Greenlaw, pers. comm.).

There are a few isolated examples of nests being placed in patches of S. alterniflora within the high marsh - low marsh interface and successfully fledging chicks due to higher placement above the marsh surface (DeRagon 1988, p.54-55). There is also limited evidence of saltmarsh

79 sparrows nesting higher up in marsh elder (Iva frutescens) on the terrestrial edge of marshes (DeRagon 1988, p. 51-52 [1.5 percent of nests, RI]). These indicate some willingness by individuals to experiment with alternate nest placement with the nest height above the marsh surface varying, but no significant difference in nest elevation (i.e. above tide level) among plant communities (DeRagon 1988, p.54-55). However, despite significant monitoring across the range of the species these observations have been exceedingly rare despite the ongoing decline in high marsh vegetation. This implies that there is very limited adaptive capacity with regard to nest site selection. The data and research strongly suggests that sea-level rise is the impetus for the observed population declines, and that despite having many adaptations that allow them to use tidal marsh habitats, with suitable habitat being limited, saltmarsh sparrows have not been able to adapt at a pace that results in a stable population.

At the northern end of their range saltmarsh sparrow hybridize with Nelson’s sparrow, which is a species that occupies a broader range of habitats, including more brackish and inland marshes. Presumably this could impart some genetic representation to hybrid individuals that would allow them to adapt to utilizing habitats that are not being impacted by sea-level rise at the same rate as salt marsh. There is extensive literature that has evaluated these populations and determined that there has been a net increase in the relative abundance of Nelson’s sparrows as compared to saltmarsh sparrows over the past 15 years in areas where the parent populations overlap. However, there is reduced fitness of hybrid individuals indicating that hybrids are not likely to swamp the parent populations and that hybridization is not driving the declines in saltmarsh sparrows. Instead, the consensus is that saltmarsh sparrow populations are declining in response to sea-level rise driven flooding of nests and loss of high marsh and those declines are exceeding any change in Nelson’s sparrow populations. This information leads us to conclude that hybridization is not providing any additional adaptive capacity for saltmarsh sparrows.

As a result of the information discussed above, we conclude that the saltmarsh sparrow has little observed ability to adapt at the pace of habitat change they are currently experiencing and, therefore, have low representation.

4.4.3 Resiliency

Resiliency is defined as a species’ ability to withstand stochastic events which is influenced by the quantity and quality of habitat as well as population size, growth rate and demographic rates (Section 4.3).

Saltmarsh sparrow populations have been declining across the range at an average of 9 percent per year between 1998 and 2012 and at 14 percent per year at demographic sites between 2010–2014. By 2012, the population had declined by 72 percent from 1998 numbers to about 60,000 individuals. If the population continued to decline at 9 percent per year through 2018, current population estimates would be 34,072 individuals, representing an 84 percent decline over a 20 year window (1998 to 2018). At the higher rate of decline observed between 2010 and 2014 projected through 2018, current population estimates would be 24,274 individuals. We consider this a generous estimate of trends from 2012–2018 as there is evidence that loss of habitat has accelerated since 2012 as a direct result of increasing rates of sea-level rise (Bowman 2015, entire; Hill and Anisfeld 2015, p. 5; Smith 2015, pp. 132–134; Beckett et al. 2016, p. 2; Mitchell et al. 2017, entire; Raposa et al. 2017a, p. 393). Declines have coincided with declines

80 in fecundity related to both flooding and predation, as well as fragmentation and loss of habitat across the range.

The primary cause of nest failure (loss of eggs and chicks) across the range is flooding. Flooding has increased in frequency and duration primarily due to sea-level rise and secondarily to increased precipitation and storm intensity during the breeding season. This is impacting all parts of the breeding range in a relatively consistent manner. Predation impacts a higher percentage of nests in New Jersey, which supports approximately 33 percent of the saltmarsh sparrow breeding population. Predation impacts across the rest of the species range is unclear as flooding of nests far outpaces any losses otherwise associated with predation. There is less certainty as to the additive impacts of other threats associated with environmental contaminants, adult or juvenile mortality during migration, or mortality in the wintering range.

There has also been range-wide degradation and loss of salt marsh habitat that has resulted in smaller and more fragmented patches of marsh habitat overall as well as a disproportionate loss in suitable nesting habitat for saltmarsh sparrows. These reductions in both habitat quantity and quality are well documented in the literature and have been extensively caused by anthropogenic activities. In addition there is a high degree of certainty that sea-level rise is a significant driver of salt marsh habitat loss within the range of the species. Site specific variability in rates of loss exist and can be correlated with numerous localized factors, including nutrient input from agricultural and urban runoff, geomorphic setting, tidal range, the presence of invasive plants or animals, and shoreline hardening. There is consensus that individually or in combination, these factors influence how well marshes have been able to persist with current rates of sea- level rise.

A summary of the resiliency scores for each of the metrics in each analytical unit follows (Table 4). Based upon this analysis, resiliency for the Southern and Central breeding range analytical units is determined to be low, while the Northern breeding range and the wintering range analytical units are considered moderate.

81 Table 4. Summary of Resiliency Scores for each Analytical Unit

Factor North AU Central AU Southern AU Winter AU Habitat Quantity Habitat Quality -Patch Size Habitat Quality -Vegetation Proximity to Next Marsh Rate of shoreline hardening Population Size Growth rate Adult Survival Juvenile Survival Nest Success Degree of Human Alteration Development within 1000m Tidal Restrictions

Overall

High Resiliency Moderate Resiliency Low Resiliency Data Unavailable or Ambiguous Not applicable

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103 Appendix; Table 1: Saltmarsh Sparrow nesting population and habitat distribution by state (calculated as part of Wiest et al. 2016).

SALS Marsh Area Total High Proportion Number of State Abundance (ha) Marsh (ha) High Marsh Patches ME 1,263.06 5,020.80 2,398.27 0.48 389 NH 1,043.93 3,010.66 1,941.53 0.64 38 MA 7,387.16 19,961.68 12,211.30 0.61 648 RI 779.17 1,197.70 67.76 0.06 193 CT 1,546.48 4,819.82 887.65 0.18 272 NY 4,759.60 10,497.98 1,905.97 0.18 686 NJ 19,908.05 81,909.53 20,417.55 0.25 587 DE 4,121.89 31,379.46 2,464.67 0.08 181 MD 14,987.79 76,711.12 12,907.39 0.17 2771 VA 4,204.45 43,203.07 24,558.40 0.57 926

104 Appendix; Table 2: Nest success data summary.

Citation Location(s) Numbe Year(s) Number of Daily Ness Nest failure Nest failure Nest failure r of Nests Probability of Success due to due to due to known Sites (sample Nest Success Rate flooding depredation causes size) (percent of (percent total) (percent total) (percent total) total) Post & New York 1 1977 - 1980 238 Not Reported 27 Not Reported Not Reported Not Reported Greenlaw 1982 DeRagon 1988 Rhode Island 4 1981 - 1982 172 Not Reported 58 26 4 10 DiQuinzio et al. Rhode Island 1 1993 - 1998 136 0.86 - 0.99 52 17 25 5 2002 Shriver et al. Maine 1 1998 - 2001 69 0.97 46 Not Reported Not Reported Not Reported 2007 Marra & Maryland 1 2002 - 2003 18 Not Reported 44 16 40 Not Reported Shmeling unpublished (from Greenberg et al. 2006) Gjerdrum et al. Connecticut 7 2002 - 2003 125 0.94 41 38 20 Not Reported 2005 Ruskin et al. Maine - New 23 2011 - 2013 837 0.86 - 0.97 Not Not Reported Not Reported Not Reported 2017b Jersey Reported Roberts et al. New Jersey 3 2011 - 2015 314 0.91 - 0.95 39 16 29 16 2017 Ruskin et al. Maine - New 23 2011 - 2015 1,416 0.95 Not 29 15 13 unpublished Jersey (average) Reported (from Greenlaw et al. 2018) Benvenuti et al. Maine - 4 2011 - 2015 555 Not Reported 49 40 11 Not Reported 2018 Massachuset ts Maxwell 2018* Maine 2 2016 -2017 201 0.96 31 34 19 16 *Includes hybrid individuals

105 Appendix; Table 3: Annual survival rates calculated from breeding or wintering sites.

Authors Location(s) Year(s) Number of Birds Adult Annual Sex Survival Rate Post & Greenlaw New York 1967 - 1972 11 0.63 Female 1982 Post & Greenlaw New York 1967 - 1972 73 0.6 Male 1982 Post & Greenlaw New York 1967 - 1972 106 0.54 Male 1982 Post & Greenlaw New York 1976-1977 66 0.53 Female 1982 DiQuinzio et al. Rhode Island 1993 - 1998 446 0.26 - 0.65 Both 2001 Winder et al. North Carolina 2006 - 2010 197 0.60 - 0.71 Both 2012 Field et al. 2018 Maine – New 2011 - 2014 3,648 0.25 - 0.67 Both Jersey Field et al. 2018 Maine – New 2011 - 2014 0.49 Male Jersey Field et al. 2018 Maine – New 2011 - 2014 0.44 Female Jersey Roberts et al. New Jersey 2011 - 2015 1,029 0.39 Both 2017 (JFO) Post unpublished South Carolina 0.58 Shell Key Banding Florida 0.21 Report Shaw 2012 South Carolina 0.63

106