i

Prepared in cooperation with the U.S. Fish and Wildlife Service

Species Accounts for the Alamosa/Monte Vista/Baca 

National Wildlife Refuge Complex

 

Open-File Report 2011–1285

U.S. Department of the Interior

U.S. Geological Survey
ii

Cover photographs: Baca National Wildlife Refuge, courtesy of Laura Ellison.
i

Prepared in cooperation with the U.S. Fish and Wildlife Service

Species Accounts for the Alamosa/Monte Vista/Baca 

National Wildlife Refuge Complex

By Laura E. Ellison

Open-File Report 2011–1285

U.S. Department of the Interior

U.S. Geological Survey 
ii

U.S. Department of the Interior

KEN SALAZAR, Secretary

U.S. Geological Survey

Marcia K. McNutt, Director

U.S. Geological Survey, Reston, Virginia 2011

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its natural and living resources, natural hazards, and the environment:

World Wide Web: http://www.usgs.gov

Telephone: 1-888-ASK-USGS

Suggested citation:

Ellison, L.E., 2011, Species accounts for the Alamosa/Monte Vista/Baca National Wildlife Refuge Complex: 

U.S. Geological Survey Open-File Report 2011-1285, 195 p.

Any use of trade, product, or firm names is for descriptive purposes only and does not imply 

endorsement by the U.S. Government. 

Although this report is in the public domain, permission must be secured from the individual 

copyright owners to reproduce any copyrighted material contained within this report.
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Contents

Executive Summary.................................................................................................................................................... 1

Literature Review Methods ......................................................................................................................................... 1

Format of Report ........................................................................................................................................................ 1

Acknowledgments ...................................................................................................................................................... 2

I. American Avocet (Recurvirostra americana) ........................................................................................................... 3

Introduction ............................................................................................................................................................. 3

Natural History ........................................................................................................................................................ 4

Morphology and Systematics .............................................................................................................................. 4

Distribution .......................................................................................................................................................... 4

Migration ............................................................................................................................................................. 4

Habitat................................................................................................................................................................. 5

Foraging Habits................................................................................................................................................... 6

Breeding.............................................................................................................................................................. 6

Demography and Populations ............................................................................................................................. 8

Threats.................................................................................................................................................................... 9

Sources of Mortality............................................................................................................................................. 9

Pesticides and Other Contaminants .................................................................................................................... 9

Loss and Degradation of Wetlands ................................................................................................................... 11

Disturbance ....................................................................................................................................................... 12

Management......................................................................................................................................................... 12

Response to Management ................................................................................................................................ 12

Recommendations ............................................................................................................................................ 13

Research Priorities................................................................................................................................................ 15

Bibliography .......................................................................................................................................................... 15

II. Wilson’s Phalarope (Phalaropus tricolor) .............................................................................................................. 27

Introduction ........................................................................................................................................................... 27

Natural History ...................................................................................................................................................... 27

Morphology ....................................................................................................................................................... 27

Distribution ........................................................................................................................................................ 28

Migration ........................................................................................................................................................... 28

Habitat............................................................................................................................................................... 29

Foraging Habits................................................................................................................................................. 30

Breeding............................................................................................................................................................ 30

Demography & Population Status ..................................................................................................................... 31

Threats.................................................................................................................................................................. 32

Sources of Mortality........................................................................................................................................... 32

Contaminants .................................................................................................................................................... 32

Loss and Degradation of Wetlands ................................................................................................................... 33

Climate Change................................................................................................................................................. 33

Management......................................................................................................................................................... 33

Response to Management ................................................................................................................................ 33

Recommendations ............................................................................................................................................ 34

Research Priorities................................................................................................................................................ 35

Bibliography .......................................................................................................................................................... 35

III. Sora (Porzana carolina) ....................................................................................................................................... 42

Introduction ........................................................................................................................................................... 42
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Natural History ...................................................................................................................................................... 42

Morphology ....................................................................................................................................................... 42

Distribution ........................................................................................................................................................ 42

Migration ........................................................................................................................................................... 43

Habitat............................................................................................................................................................... 44

Foraging Habits................................................................................................................................................. 45

Breeding............................................................................................................................................................ 46

Demography & Population Status ..................................................................................................................... 46

Threats.................................................................................................................................................................. 47

Shooting and Trapping...................................................................................................................................... 47

Other Potential Threats ..................................................................................................................................... 47

Management......................................................................................................................................................... 48

Response to Management ................................................................................................................................ 48

Recommendations ............................................................................................................................................ 49

Research Priorities................................................................................................................................................ 50

Bibliography .......................................................................................................................................................... 50

IV. White-faced Ibis (Plegadis chihi) ......................................................................................................................... 56

Introduction ........................................................................................................................................................... 56

Natural History ...................................................................................................................................................... 56

Distribution ........................................................................................................................................................ 56

Migration ........................................................................................................................................................... 57

Habitat............................................................................................................................................................... 57

Food Habits....................................................................................................................................................... 59

Breeding............................................................................................................................................................ 59

Demography & Populations............................................................................................................................... 61

Threats.................................................................................................................................................................. 62

Management......................................................................................................................................................... 64

Response to Management ................................................................................................................................ 64

Recommendations ............................................................................................................................................ 64

Research Priorities................................................................................................................................................ 65

Bibliography .......................................................................................................................................................... 66

V. Black Tern (Chlidonias niger) ............................................................................................................................... 72

Introduction ........................................................................................................................................................... 72

Natural History ...................................................................................................................................................... 72

Systematics and Morphology ............................................................................................................................ 72

Distribution ........................................................................................................................................................ 73

Migration ........................................................................................................................................................... 73

Habitat............................................................................................................................................................... 74

Foraging Habits................................................................................................................................................. 76

Breeding............................................................................................................................................................ 76

Demography and Populations ........................................................................................................................... 77

Threats.................................................................................................................................................................. 77

Sources of Mortality........................................................................................................................................... 77

Loss and Degradation of Habitat....................................................................................................................... 78

Management......................................................................................................................................................... 78

Response to Management ................................................................................................................................ 78

Recommendations ............................................................................................................................................ 79

Research Priorities................................................................................................................................................ 81
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Bibliography .......................................................................................................................................................... 81

VI. Short-eared Owl (Asio flammeus) ....................................................................................................................... 89

Introduction ........................................................................................................................................................... 89

Natural History ...................................................................................................................................................... 90

Morphology & Systematics................................................................................................................................ 90

Migration ........................................................................................................................................................... 90

Habitat............................................................................................................................................................... 90

Foraging Habits................................................................................................................................................. 91

Breeding............................................................................................................................................................ 91

Demography & Populations............................................................................................................................... 92

Threats.................................................................................................................................................................. 92

Management......................................................................................................................................................... 93

Response to Management ................................................................................................................................ 93

Recommendations ............................................................................................................................................ 94

Research Priorities................................................................................................................................................ 95

Bibliography .......................................................................................................................................................... 96

VII. Brewer’s Sparrow (Spizella breweri) ................................................................................................................ 104

Introduction ......................................................................................................................................................... 104

Natural History .................................................................................................................................................... 104

Morphology & Systematics.............................................................................................................................. 104

Migration ......................................................................................................................................................... 104

Habitat............................................................................................................................................................. 105

Foraging Habits............................................................................................................................................... 105

Breeding.......................................................................................................................................................... 106

Demography & Populations............................................................................................................................. 106

Threats................................................................................................................................................................ 107

Sources of Mortality......................................................................................................................................... 107

Habitat Loss & Fragmentation......................................................................................................................... 108

Agriculture ....................................................................................................................................................... 109

Invasion of Exotic Annual Grasses.................................................................................................................. 109

Prescribed Fire................................................................................................................................................ 109

Mining and Oil/Gas Development.................................................................................................................... 110

Livestock Grazing............................................................................................................................................ 110

Management....................................................................................................................................................... 110

Response to Management .............................................................................................................................. 110

Recommendations .......................................................................................................................................... 111

Management Approaches ............................................................................................................................... 111

Research Priorities and Information Needs......................................................................................................... 112

Bibliography ........................................................................................................................................................ 114

VIII. Savannah Sparrow (Passerculus sandwichensis) ........................................................................................... 120

Introduction ......................................................................................................................................................... 120

Natural History .................................................................................................................................................... 121

Morphology & Systematics.............................................................................................................................. 121

Migration ......................................................................................................................................................... 121

Habitat............................................................................................................................................................. 121

Foraging Habits............................................................................................................................................... 122

Breeding.......................................................................................................................................................... 122

Demography & Populations............................................................................................................................. 123
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Threats................................................................................................................................................................ 123

Sources of Mortality......................................................................................................................................... 123

Effects of Human Activity ................................................................................................................................ 124

Management....................................................................................................................................................... 124

Mowing/Haying................................................................................................................................................ 124

Grazing............................................................................................................................................................ 126

Burning............................................................................................................................................................ 127

Other Management ......................................................................................................................................... 128

Research Priorities.............................................................................................................................................. 129

Bibliography ........................................................................................................................................................ 129

IX. Northern Leopard Frog [Lithobates (=Rana) pipiens] ........................................................................................ 137

Introduction ......................................................................................................................................................... 137

Natural History .................................................................................................................................................... 137

Morphological Characteristics ......................................................................................................................... 137

Distribution and Abundance ............................................................................................................................ 138

Habitat............................................................................................................................................................. 139

Activity and Movement Patterns...................................................................................................................... 139

Reproduction and Growth ............................................................................................................................... 140

Food Habits..................................................................................................................................................... 140

Predators......................................................................................................................................................... 140

Threats................................................................................................................................................................ 141

Extrinsic Threats.............................................................................................................................................. 141

Landscape-level Processes......................................................................................................................... 141

Nonindigenous Biotic Agents....................................................................................................................... 143

Water Quality............................................................................................................................................... 144

(1) Pesticides ........................................................................................................................................... 144

(2) Fertilizers ............................................................................................................................................ 146

(3) Mining & Heavy Metals ....................................................................................................................... 146

(4) pH/Acidification................................................................................................................................... 147

(5) Motorized Transportation.................................................................................................................... 148

(6) Cattle Grazing..................................................................................................................................... 148

(7) Sedimentation..................................................................................................................................... 149

(8) Rotenone............................................................................................................................................ 149

(9) Polychlorinated biphenyls (PCBs)....................................................................................................... 149

(10) Other Contamination and Limb Malformation ................................................................................... 149

Intrinsic Threats............................................................................................................................................... 150

Management....................................................................................................................................................... 151

Research Priorities.............................................................................................................................................. 154

Bibliography ........................................................................................................................................................ 155

X. Tadpole Shrimp (Triops longicaudatus) .............................................................................................................. 174

Introduction ......................................................................................................................................................... 174

Natural History .................................................................................................................................................... 174

Threats................................................................................................................................................................ 176

Management....................................................................................................................................................... 177

Research Priorities.............................................................................................................................................. 180

Bibliography ........................................................................................................................................................ 180

Summary ................................................................................................................................................................ 187
1

Species Accounts for the Alamosa/Monte Vista/Baca 

National Wildlife Refuge Complex

By Laura E. Ellison

Executive Summary

As part of an interagency agreement between the U.S. Geological Survey (USGS) and the U.S. 

Fish and Wildlife Service (USFWS), the Alamosa/Monte Vista/Baca National Wildlife Refuge 

Complex requested help with the synthesis of scientific information for 10 focal species and their 

habitat requirements in response to common Refuge management activities in the San Luis Valley, 

Colorado. This information will be instrumental in developing the Service’s Comprehensive 

Conservation Plan (CCP), which is required by law for each unit of the National Wildlife Refuge 

System. After consultation with Refuge managers and USGS staff, the 10 species chosen for detailed 

literature reviews and synthesis of information were the following: (1) American Avocet (Recurvirostra

americana); (2) Wilson’s Phalarope (Phalaropus tricolor); (3) Sora (Porzana carolina); (4) Whitefaced Ibis (Plegadis chihi); (5) Black Tern (Chlidonias niger); (6) Short-eared Owl (Asio flammeus); (7) 

Brewer’s Sparrow (Spizella breweri); (8) Savannah Sparrow (Passerculus sandwichensis); (9) Northern 

Leopard Frog [Lithobates (=Rana) pipiens]; and, (10) Tadpole Shrimp (Triops longicaudatus).

Literature Review Methods

A detailed literature review was conducted for each of the focal species using Internet resources 

and databases. The keywords chosen for these searches included the species common and scientific 

names, alone and in combination with the terms management, grazing, haying, mowing, threats, fire and 

prescribed fire, burning, water and water management, and Colorado. Databases and search engines 

used included Google (http://www.google.com), Google Scholar (http://scholar.google.com), SciVerse 

Scopus (http://www.scopus.com), ISI Web of Knowledge (http://apps.isiknowledge.com), and the U.S. 

Geological Survey (USGS) Library’s Digital Desktop (http://library.usgs.gov). The literature cited 

sections of manuscripts obtained from keyword searches were cross-referenced to identify additional 

citations or “gray literature” that was missed by the Internet search engines. 

Although Colorado was a keyword in the literature searches, information gathered to develop the 

different species accounts is based on research conducted throughout the United States. As a result, 

research conclusions based on coastal or plains studies may, or may not, be applicable to the high 

elevation, intermountain basin in which the Alamosa/Monte Vista/Baca complex is located, but do 

provide general context to each topical theme being discussed herein. 

Format of Report

This report is divided into 10 sections, with each section pertaining to one of the focal species. 

Each section is formatted on the basis of the following themes: an introduction to the species, a natural 

history subsection with pertinent facts about its unique biology and habitat requirements, a subsection 

describing threats to the species, a subsection on management activities of concern to the species, and a 
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bibliography. Not all bibliographic references are cited within the text of the document. The final 

Summary subsection provides a table of general habitat requirements for the 10 focal species. 

Acknowledgments

I thank Megan K. Eberhardt-Frank [U.S. Geological Survey (USGS), Fort Collins Science 

Center (FORT)] for her invaluable help with the detailed literature reviews for each species. David 

Hamilton (USGS, FORT) and Murray Laubhan (U.S. Fish and Wildlife Service, Quivira National 

Wildlife Refuge) provided helpful peer review comments. Eleanor Briccetti, photographer, provided the 

photographs of American avocets, savannah sparrow, short-eared owl, white-faced ibis, and Wilson’s 

phalarope. The USFWS provided photographs of the Brewer’s sparrow and sora. The northern leopard 

frog photograph is by Laurie Averill-Murray, Pacific Southwest Region, USFWS. The tadpole shrimp 

photograph is from Wikipedia (http://en.wikipedia.org/wiki/File:Triops_longicaudatus.jpg). The 

photograph of the black tern is by Jessi Miller. Jennifer Shoemaker (USGS, Science Publishing 

Network) provided assistance with formatting the final report. 
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I. American Avocet 

(Recurvirostra americana) 

Introduction

The American Avocet

(Recurvirostra americana) is among 

the most abundant avian species 

inhabiting inland, saline wetlands in 

North America (Wells, 2003). The 

American Avocet, hereafter “avocet,”

specializes in using temporally 

unpredictable wetlands of the arid 

western United States and breeds in 

large numbers at the marshes of Great 

Salt Lake, the Tulare Basin of California, and in shifting abundance across the northern Great Basin 

(Robinson, 1996; Robinson and others, 1997). In the arid west, wetland priorities and urban priorities 

compete directly for limited supplies of fresh water. The use of water to meet the needs of healthy 

wetlands is competing with other uses for water, including human population growth and increasing 

urbanization (Meyer and others, 2007). Some wetlands that were once important avocet breeding 

centers have declined in area by as much as 90 percent. Avocets have also been victims of the Kesterson 

effect—a set of embryo deformities associated with selenium contamination in irrigation drain water 

first noted at Kesterson National Wildlife Refuge, California (Presser, 1994). Avocets were extirpated 

from much of their eastern range in the 1900s, and their continued persistence in the western United 

States will depend on inland wetland conservation and restoration efforts. Avocets are protected from 

unregulated destruction under general wildlife laws such as the Federal Migratory Bird Treaty Act and 

various state codes (Robinson and others, 1997). 

Research on avocets has primarily focused on behavior and population ecology (Robinson and 

others, 1997). Hamilton (1975) described and compared the behaviors of avocets and black-necked stilts 

(Himantopus mexicanus) in California. Gibson (1971; 1978) recorded breeding-season time budgets 

from Oregon. The conspicuous antipredator behaviors of avocets were studied in depth by Sordahl 

(1980; 1981; 1982; 1984; 1986; 1988; 1990; 1994; 2004) in Utah. Winter behavior and habitat use was 

the focus of research conducted in South Carolina (Boettcher, 1994; Boettcher and others, 1994; 1995) 

and in California (Evans, 1988; Evans and Harris, 1994). Avocets and black-necked stilts also have been 

a primary focus of ecotoxicological studies on the effects of irrigation drainage water on breeding 

waterbirds (for example, studies by Ohlendorf, Skorupa). Robinson and Oring (Robinson, 1996; 

Robinson and Oring, 1996; 1997) conducted population studies of hundreds of marked individuals in 

California, Nevada, and Utah, providing data on migratory movements, natal and breeding dispersal, 

population regulation, and population spatial structure (Robinson and others, 1997). 

A simple Google search on the Internet for “Recurvirostra americana” revealed 115,000 results 

and a Google Scholar search provided 1,820 results (accessed 11 July 2011). “American avocet” 

revealed 474,000 results through Google and 2,040 results through Google Scholar (accessed 11 July 

2011). 

American Avocets, courtesy of Eleanor Briccetti.
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Natural History

Morphology and Systematics

Avocets are long-legged shorebirds approximately 43-47 cm (centimeters) long with a bill length 

of 8.2-9.5 cm and a wingspread up to 91 cm. They weigh between 275 and 350 g (grams). Sexes have 

similar plumage, but differ in size; females are smaller and have a shorter and more strongly recurved 

bill than males (Robinson and others, 1997). The plumage and morphology of this species are 

distinctive among other North American shorebirds. They are distinguished by strikingly contrasted 

black-and-white upperparts, white underparts, either rusty or gray head and neck, long upturned bill, 

large size, and long, grayish blue legs. Plumage changes seasonally; head and neck are rusty in breeding 

plumage and light gray in basic (winter) plumage. 

Avocets are members of the Recurvirostridae family, which comprises three genera: 

Recurvirostra (avocets), Cladorhynchus (banded stilt), and Himantopus (stilts). Chu (1995) reviewed 

the multiple phylogenetic data analyses for this clade; the clade is probably most closely related to 

plovers, lapwings, and oystercatchers. There is also some fossil evidence that Recurvirostridae and 

flamingoes (Phoenicopteridae) are related (Olson and Feduccia, 1980). A hybrid avocet and blacknecked stilt hatched in captivity and the morphological characteristics were intermediate between the 

two species (Principe, 1977). 

Distribution

Avocets breed commonly throughout the West in the Great Basin and Great Plains south to 

California and Texas (Robinson and others, 1997). The northern extent of their breeding range is central 

Alberta, southern Alberta, southern Saskatchewan, southwestern Manitoba, and western Minnesota. 

They winter on the Pacific and Gulf coasts from southern California into Mexico. 

Migration

Avocets are medium-distance continental migrants from southern Canada to Mexico. The 

patterns of migration are poorly known and variable among years (Robinson and others, 1997). In the 

western United States, migration has both a north-south component and an inland-coastal component. 

They are known to stop for prolonged periods of time at intermediate sites throughout the migratory 

period (Robinson and Oring, 1996). The greatest fall and spring migration concentrations are through 

southeastern Oregon, central California, San Francisco Bay, Lahontan Valley, Nevada, and the Great 

Salt Lake, Utah. In central United States, the migration pattern is more dispersed and opportunistic 

(Skagen and Knopf, 1993). 

In the western Great Basin, avocets exhibited movement prior to fall migration, possibly due to 

advantages of flocking during molt, location of an abundant food source, or assessing future potential 

breeding or staging sites (Plissner and others, 2000). For 25 avocets known to have remained in the 

survey region from early June through late September, the average number of lakes used was 2.1. Of 

161 banded and radio-tagged birds, the mean distance traveled between lakes within the region prior to 

migration was 145 km. No difference was detected between sexes in number of movements, number of 

lakes visited, or persistence in the region. Contrary to fall migratory movements, avocets during the 

premigratory (both spring and fall) period tended to move in a northerly rather than southerly direction, 

possibly to areas of abundant food resources. Resightings of color-banded avocets on breeding grounds 

in California and Utah indicated that avocets used several migratory routes, migrating individuals spent 

as long as 48 days at stopover sites, and traveled relatively short (about 250-500 km) distances between 
5

wetland sites, and siblings may travel together (Robinson and Oring, 1996). Compared to adults, firstyear avocets were more likely to be at coastal sites rather than inland sites.

Habitat

The Colorado Breeding Bird Atlas reported 74 percent of their avocet sightings occurred in 

emergent wetland marshes and lakes (Winternitz, 1998). Little is known about the area requirements of 

avocets. However, in a study in Manitoba examining the effects of wetland size on the occurrence of 

bird species, Daub (1993) found that avocets occurred only on wetlands 6-19.3 ha in size. Daub (1993) 

sampled wetlands ranging from <1 to 19.3 ha, and concluded that the avocet is an area-dependent 

species. 

Avocets use exposed, sparsely vegetated salt flats, sandbars, peninsulas, mudflats, or islands 

adjacent to shallow (<1 m deep) water, conditions that occur in wetlands, lakes, fallow and flooded 

fields, or impoundments (Mitchell, 1917; Bent, 1962; Vermeer, 1971; Maher, 1974; Kondla, 1977; 

Kondla and Pinel, 1978; Grover, 1979; Grover and Grover, 1981; Grover and Knopf, 1982; Sidle and 

Arnold, 1982; Weber and others, 1982; Hirsch and Fouchi, 1992; Gunderson and others, 1992; See and 

others, 1992; Faanes and Lingle, 1995; Robinson and others, 1997; Beaver, 1998). Breeding habitat can 

be found in tilled, alkali, ephemeral, temporary, seasonal, semipermanent, and permanent wetlands, and 

in impoundments, sewage lagoons, and evaporation ponds (Stewart and Kantrud, 1965; Weber, 1978; 

Faanes, 1982; Weber and others, 1982; Kantrud and Steward, 1984; Giroux, 1985; Colwell, 1986; Dole, 

1986; Prescott and others, 1995; Baylor, 1996; Robinson and others, 1997). In North Dakota, density of 

avocets was greatest in tilled wetlands, followed by alkali, seasonal, and semipermanent wetlands 

(Kantrud and Stewart, 1984). In South Dakota, avocets most frequently occurred in ephemeral wetlands, 

followed by tilled wetlands and wetlands located in pastures (Weber, 1978; Weber and others, 1982). In 

Alberta, avocets were most abundant in large (>8 ha) saline wetlands, followed by medium-sized (1-8 

ha) saline wetlands, large (>8 ha) fresh wetlands, small (<1 ha) saline wetlands, medium-sized fresh 

wetlands, and small fresh wetlands (Prescott and others, 1995). Avocets may exhibit annual variation in 

nest site selection, moving to deeper, more permanent wetlands in dry years (Colwell, 1986; 1991). In 

Colorado, avocets preferred seasonal wetlands and habitats dominated by Baltic rush (Juncus balticus), 

sedges (Carex spp.), and grasses <40 cm tall over semipermanent wetlands, habitats dominated by 

cattail (Typha spp.) and softstem bulrush (Scirpus validus) >40 cm tall, saltgrass habitats, or upland 

shrub habitats (Laubhan and Gammonley, 2000). 

Avocets generally nest on unvegetated ground or in areas with short, sparse vegetation that 

provide an unobstructed view from the nest (Sidle and Arnold, 1982; Giroux, 1985; Dole, 1986; 

Robinson and others, 1997; Conway and others, 2005b). Of 199 avocet pairs in Montana, 65 percent

nested on bare ground, 13 percent nested in short (<12 cm tall) grasses, 15 percent nested in grasses of 

moderate height (13-37 cm tall), 7 percent nested in tall (38-61 cm) grasses, and no avocets nested in 

very tall (>61 cm) grasses (Dole, 1986). Nests are often slightly elevated (2-15 cm) above the 

surrounding substrate (Sidle and Arnold, 1982; Dole, 1986). In North Dakota and Montana, nests that 

were closest to the water’s edge, on very wet sites, or in frequently flooded areas usually were elevated 

from the ground with vegetation (Sidle and Arnold, 1982; Dole, 1986). 

Avocet nests are often placed near water and near clumps of vegetation or debris (for example, 

driftwood, fence posts) (Grover, 1979; Winton and Leslie, 1997). Distance of avocet nests to water 

range from 0 to 300 m, although most nests are located within 60 m of water (Grover, 1979; Grover and 

Knopf, 1982; Sidle and Arnold, 1982; Giroux, 1985; Gunderson and others, 1992; Kuyt and Johns, 

1992). Nests in Oklahoma were significantly closer to water than random points on a study area 

(Grover, 1979; Grover and Knopf, 1982). Of 23 nests in Oklahoma, 57 percent were located within 5 
6

cm of debris (Grover, 1979). In another Oklahoma study, 52 percent of 41 avocet nests were found 

within 5 cm of driftwood clumps or dead vegetation, 29 percent on human-made elevated habitat 

structures (for example, on plowed ridges and gravel mounds), 12 percent on open soil or sand, and 7 

percent within 5 cm of live vegetation. Nests on open soil/sand experienced the lowest apparent nest 

success. 

In the prairie regions of Alberta, Manitoba, Saskatchewan, and North Dakota, breeding densities 

of avocets were higher on islands than along wetland shorelines (Kondla and Pinel, 1978; Sidle and 

Arnold, 1982). Nest densities of avocets on islands in Montana, North Dakota, and South Dakota were 

significantly and positively related to percent bare ground and presence of saline wetlands (Lokemoen 

and Woodward, 1992). [However, the frequency of intraspecific brood parasitism is greater on islands 

and in the more northerly portions of the breeding range (Kondla and Pinel, 1978; Robinson and others, 

1997) so drawing conclusions from nest “densities” as opposed to nest “success” can be misleading.] On 

constructed islands in wetlands of the Prairie Pothole Region of North Dakota, the density of avocet 

nests was higher on islands with beaches than than on islands without beaches (Dahl and others, 2003). 

In Alberta, the density of nesting avocets on islands in wetlands created to enhance waterfowl 

production was positively related to percentage of the island surface covered by sparse vegetation and 

with maximum water depth between the island and the nearest shoreline (Giroux, 1985). Newly 

constructed islands were preferred over older, more vegetated islands. Amount of sparse vegetation was 

inversely related to years since island construction. In Colorado, a series of newly constructed shallow 

wetlands and resting spots created for waterfowl contained nonvegetated areas suitable for nesting 

avocets (Beaver, 1998). 

Foraging Habits

Avocets forage at various water depths, depending on age and bill length (Dole, 1986). Young 

(0-3 week-old) chicks forage at water depths of 0-90 mm (millimeters), but mainly at about 8 mm. 

Older (3-6 week-old) chicks forage at depths from 0 to 100 mm, but mostly at about 53 mm. Adult 

females forage at water depths >80 mm, and adult males forage at depths of about 100 mm. Osmundson 

(1990) also reported that adult avocets forage at depths ranging from 90 to 160 mm. Avocets partition 

food resources depending on water depth and method of prey capture. Food is obtained by scythe-like 

movements in shallow water or in pools of water standing on the mud or by pecking (Burger and 

Gochfeld, 1986). Foraging avocets use habitats with fine sediments (Quammen, 1982; Danufsky and 

Colwell, 2003). The main foods taken by avocets are aquatic invertebrates in the water column and in 

sediment. They also eat terrestrial invertebrates, small fish, and seeds (Robinson and others, 1997). 

Ceratopogonidae (biting midges) and Chironomidae (nonbiting midges) were the dominant families of 

invertebrates eaten by avocets in migratory stopover sites in the Southern Great Plains (Andrei and 

others, 2009). Chironomid larvae were the major dietary component (44-77 percent) for migrating 

avocets on the Texas high plains (Baldassare and Fischer, 1984). Avocets also preferred intermediate 

and larger prey than conspecific foragers such as least and western sandpipers (Davis and Smith, 2001). 

In hypersaline foraging conditions, through a combination of behavioral and anatomical adaptations, 

avocets are able to rid their prey of most of the adherent lake water and thereby largely avoid problems 

of salt-loading and the ingestion of harmful ions (Mahoney and Jehl, 1985). 

Breeding 

Avocets arrive on the breeding grounds from early April to mid-May and depart from midAugust to mid-October (Bent, 1962; Salt and Salt, 1976; Knapton, 1979; Johnsgard, 1980; Sidle and 

Arnold, 1982; Hill, 1985; Dole, 1986; Janssen, 1987; Colwell and others, 1988; See and others, 1992; 
7

Robinson and others, 1997; Svingen, 1998). In Montana, peak egg-laying occurred in May (Dole, 

1986). The species is single-brooded but commonly will renest following failure of the initial nest 

(Gibson, 1971; Sordahl, 1984; Dole, 1986; Robinson and others, 1997). Colorado Breeding Bird 

Atlasers found the earliest nest in April in Colorado, but the remainder were found from mid-May to 30 

June. Precocial young were observed only from mid-June to mid-July. Atlas workers also found most 

nests and young in marshes (42 percent), around lakeshores (30 percent), and in salt meadows (8 

percent). In 1875, the first avocet nest discovered in Colorado was found by Henshaw of the Wheeler 

Survey in Costilla County in the San Luis Valley (SLV) (Winternitz, 1998). Large numbers of avocets 

nest around the Great Salt Lake in Utah and a gap in their distribution used to exist between there and 

breeding avocets in the Colorado mountain parks. Since 1986, however, avocets have extended their 

breeding range to the western slope of Colorado. The Colorado Breeding Bird Atlas map showed 

substantial breeding in northeastern Colorado, both along the South Platte corridor and the edge of the 

Front Range from Denver north. A cluster of avocets in east-central Colorado consisted of birds 

scattered at reservoirs and at permanent and ephemeral ponds from the Arkansas River north to the 

Smoky Hill River in Cheyenne County. The SLV and North Park had concentrations of breeding 

avocets, but Atlasers also found more in South Park than expected and a surprising cluster in four 

blocks surveyed east of Interstate 25 in Huerfano and Las Animas Counties. The Atlas map suggested a 

gradual expansion of breeding populations of avocets in Colorado (Winternitz, 1998). 

Avocets are semicolonial nesters (Robinson and others, 1997). Nesting most often occurs in 

loose colonies, although solitary nests also occur (Gibson, 1971; Grover, 1979; Grover and Knopf, 

1982; Sidle and Arnold, 1982; Giroux, 1985; Dole, 1986; Robinson and others, 1997). Avocets also 

may nest in association with other species. In Oklahoma, avocets nested near snowy plovers (Chardrius

alexandrines) and least terns (Sterna antillarum) (Hill, 1985; Winton and Leslie, 1997). Of 41 avocet 

nests, 64 percent nested <180 m and 37 percent nested >180 m from another active plover, tern, or 

avocet nest (Hill, 1985). Mortality rate was lowest for nests in which the nearest nest belonged to a least 

tern, suggesting that avocets benefitted from nest defense behaviors of least terns. Mortality rates of 

nests of colonial and solitary avocets, however, were not significantly different. In North Dakota and 

Montana, mean distance between nests varied according to habitat (Sidle and Arnold, 1982; Dole, 

1986). Mean distance between nests on islands or isolated sandbars ranged from 0 to 19 m, whereas 

mean distance between nests on open salt flats, peninsulas, or shorelines ranged from 10 to 141 m 

(Vermeer, 1971; Grover, 1979; Grover and Knopf, 1982; Sidle and Arnold, 1982, Giroux, 1985; Dole, 

1986). In North Dakota, islands occupied by American white pelicans (Pelecanus erythrorhynchos), 

double-crested cormorants (Phalacrocorax auritus), and gulls (Larus spp.) were avoided by nesting 

avocets (Sidle and Arnold, 1982). 

Intraspecific brood parasitism commonly occurs within avocet colonies (Bent, 1962; Vermeer, 

1971; Kondla, 1977; Kondla and Pinel, 1978; Giroux, 1985; Koonz, 1985; Dole, 1986; Kuyt and Johns, 

1992; Robinson and others, 1997). The normal clutch size for avocets is four eggs, but when two or 

more females lay eggs in the same nest, superclutches of >6 eggs can occur (Gibson, 1971; Koonz, 

1985). The frequency of intraspecific brood parasitism is greater on islands and in more northerly 

portions of the breeding range (Kondla and Pinel, 1978; Robinson and others, 1997). Reported 

frequencies of intraspecific brood parasitism ranged from 8 percent of 13 nests to 50 percent of 30 nests 

(Vermeer, 1971; Kondla and Pinel, 1978; Giroux, 1985; Koonz, 1985). Dole (1986) found that nest 

success was greater for four-egg clutches than clutches with either less than or greater than 4 eggs. Dole 

suggested that clutches of <4 eggs do not stimulate adult avocets to incubate, whereas it may be 

physically difficult for avocets to effectively incubate clutches with >4 eggs.
8

Preference for islands by breeding avocets may be due to greater protection from nest predators 

(Sidle and Arnold, 1982; Giroux, 1985). Robinson and others (1997) suggested that breeding densities 

of avocets were most influenced by the availability of nest sites, especially on islands, rather than 

overall wetland area. In Montana, hatching success was associated with island size and the distance of 

the island from shore (Dole, 1986). In contrast to findings in North Dakota, hatching success in

Montana was lower on islands than along shorelines, mudflats, or a dry lake bed. Mammalian predators 

may have been able to swim or wade to islands that were near shore. Small islands with high nest 

densities were at risk of greater depredation because of the ease with which predators located nests 

(actual island sizes and distance of islands to shore were not given in the publication). 

Although intraspecific brood parasitism is considered common, interspecific brood parasitism 

rarely occurs. An incident of interspecific brood parasitism by avocets was noted by Kuyt and Johns 

(1992) on breeding grounds in the Northwest Territories. Avocet eggs were found in two Mew gull 

(Larus canus) nests that were located near a small group of avocet nests. No known records of brood 

parasitism by brown-headed cowbirds (Molothrus ater) exist. Avocets are not suitable hosts for 

cowbirds because avocet young are precocial, whereas cowbird young are altricial.

Adult avocets occasionally return to the same breeding sites in subsequent years and may retain 

their mates of the previous year or years (Gibson, 1971; Sordahl, 1984; Colwell and Oring, 1988b; 

Robinson and Oring, 1997; Robinson and others, 1997). In California and Nevada, 116 of 478 adults 

(24 percent) bred one to two years after banding within 20 km of the banding site, 140 (29.3 percent) 

were seen in subsequent years as nonbreeders or as migrants, and 222 (46.4 percent) were never seen 

again (Robinson and Oring, 1997). No evidence of sex difference in philopatry was detected. Of 25 

females and 35 males, 13 (52 percent) and 14 (40 percent), respectively, retained their mates from 

previous years (Robinson and Oring, 1997). In Oregon, one pair was observed with young the first year 

after banding, was absent the second year after banding, but was seen together again the third year after 

banding (Gibson, 1971). One pair remained paired for at least 3 years. In Utah, six of 21 (29 percent) 

adults returned to the study area the first year after banding (Sordahl, 1984). One pair was seen in two 

consecutive years, and one adult male was seen during three years. In Saskatchewan, two of two males 

and one of two females returned to the study area one year after marking (Colwell and Oring, 1988b). 

After-hatch-year avocets occasionally return to breed at natal sites. In California, of 811 nestling 

avocets banded at one site, 34 (4.2 percent) returned when they were one or two years old (Robinson 

and Oring, 1997). Most avocets do not breed until they are two years old; of the 811 nestlings, 13 (1.6 

percent) returned as breeding adults (Robinson and Oring, 1997). One example of long-distance 

dispersal was documented: a male banded as a chick was captured on a nest 480 km from the banding 

location. In Saskatchewan, of 13 marked chicks, none returned to the breeding site (Colwell and Oring, 

1988b). In Utah, of 49 banded young, none returned in subsequent years (Sordahl, 1984). 

Demography and Populations 

An adult avocet banded in 1977 was later recovered in 1985, so the bird was at least 9 years old 

(Robinson and others, 1997). There are no data available on annual survival of avocets. Population 

status of breeders is not known (Robinson and others, 1997). At least 100,000 individuals winter in 

North America west of the Rocky Mountains (Page and Gill, 1994). These numbers are probably 

reduced compared to those of presettlement conditions because of wetland losses during the last 200 

years (Page and Gill, 1994). Peak numbers of migrants are approximately 14,000 in southeastern 

Oregon, 10,000 in the Central Valley, California (Shuford and others, 1994a); 100,000 at the Salton Sea, 

California (Setmire and others, 1993), 24,000 in the Lahontan Valley, Nevada, and 252,000 at the Great 

Salt Lake, Utah (Shuford and others, 1994b). 
9

Breeding Bird Survey (BBS) data showed no significant trends for the period 1966-1994 in the 

conterminous United States, its regions, or the entire survey area (Sauer and others, 1996). During 1965-

1979, statistically significant trends in avocet numbers were detected, with increases in the central 

region, in the conterminous United States, and continentwide (Robbins and others, 1986). There were no 

significant trends for these regions, however, during 1980-1994 (Sauer and others, 1996). The BBS 

regional trend analysis for avocets along 30 Colorado BBS routes from 1966 to 1999 showed a negative 

trend estimate of -4.8 (-9.8, 0.2; 95 percent CI) (Sauer and others, 2011). From 1999 to 2009, the trend 

was also negative and estimated to be -5.0 (-12.4, 2.2; 95 percent CI). However, the regional credibility 

measure for this species for Colorado was category “yellow,” meaning that data were deficient because 

of overall low abundances of avocets per route and/or the results were too imprecise to detect long-term 

trends. 

Threats

Sources of Mortality

Avocets and black-necked stilts are the shorebird species most frequently stricken with botulism, 

a disease caused by the bacterium Clostridium botulinum type C (Rosen, 1971; Adams and others, 

2003). The diets, ranges, and habitat preferences of avocets all correspond to risk factors for occurrence 

of avian botulism epidemics. Large numbers of individuals are most likely to be affected in the western 

United States, especially California, Utah, and Nevada, in late summer and early fall, and in years when 

higher-than-normal precipitation produces abundance of shallow, stagnant wetlands (Robinson and 

others, 1997). The botulism toxin causes flaccid paralysis, eliminating the ability to use legs (Rosen, 

1971), and in avocets, causes twisting of the neck (Alcorn, 1942). 

Major predators of adult avocets are peregrine falcons (Falco peregrinus; Evans, 1988; Porter 

and White, 1973), prairie falcon (F. mexicanus; Porter and White, 1973), great-horned owl (Bubo

virginianus; Robinson and others, 1997), northern harrier (Circus cyaneus; Evans, 1988), and red fox 

(Vulpes vulpes; Robinson and others, 1997). On young avocets, predators include peregrine falcon, 

northern harrier, and mink (Mustela vison; Robinson and others, 1997). Egg predators include California 

(Larus californicus) and ring-billed (L. delawarensis) gulls, common raven (Corvus corax), badger 

(Taxidea taxus), striped skunk (Mephitis mephitis), long-tailed weasel (Mustela frenata), coyote (Canis

latrans), raccoon (Procyon lotor), red fox, and gopher snake (Pituophis melanoleucus) (Gibson, 1971; 

Grover and Knopf, 1982; Sidle and Arnold, 1982; Alberico, 1993; Robinson and others, 1997). 

In addition to disease and predation, mortality of flightless young was observed following 

unseasonably cold storms, and chick mortality due to the effects of contaminants was observed 

(Robinson and others, 1997 – see next section for more information on contaminants). Shooting and 

trapping led to population declines, particularly on the Atlantic Coast, before the 1900s (Robinson and 

others, 1997). Avocets were formerly an important game bird in California, and large numbers were 

sold in markets in Stockton and San Francisco (Grinnell and others, 1918). Avocets are now protected 

under the Migratory Bird Treaty Act of 1918 and illegal shooting and trapping is now probably a minor 

source of mortality. 

Pesticides and Other Contaminants

Eight years after DDT was banned in the United States, DDE was still accumulating in wintering 

avocets along the south Texas coast (Robinson and others, 1997). The levels of these organic 

compounds detected in birds increased by 40 percent from October to December, then remained stable 

until February (White and others, 1983). Elevated, but most likely nontoxic, levels of the residues of 
10

DDT metabolites (particularly DDE) in eggs were detected in California, Idaho, Oregon, Montana, and 

Wyoming from 1986-1991 (Knapton and others, 1988; Lambing and others, 1988; Peterson and others, 

1988; Schroeder and others, 1988; Rinella and Schuler, 1992; Rinella and others, 1994; Henny and 

others, 2008). The maximum recorded DDE concentration of 5.7 ppm (parts per million; wet weight) 

was found in an egg from Montana (1986-1987). The presumptive adverse effects threshold for DDE is 

from 3-8 ppm (wet weight). 

Many wetlands used by avocets in the western United States have been contaminated as a result 

of irrigation and other human activities (U.S. Fish and Wildlife Service, 1992). The first indication that 

irrigation practices were contaminating wetlands on a dramatic scale was at Kesterson National Wildlife 

Refuge, San Joaquin Valley, California (Ohlendorf and others, 1986b; 1989; 1990). Irrigation drainage 

water was diverted to the Refuge for disposal and the project was initially viewed as a habitat 

replacement success. Like other waterbirds, however, avocets breeding at Kesterson failed to reproduce 

successfully (Williams and others, 1989) because drain water supplying the wetland was contaminated 

with selenium concentrated through soil-leaching and water-recycling (Ohlendorf and others, 1986a). 

Bioaccumulation of these toxicants resulted in low hatchability of eggs, dramatic developmental defects 

in embryos that survived to later stages of development (teratogenesis; Hoffman and others, 1988), and 

subsequent death of chicks that did not hatch (Ohlendorf and others, 1986b; 1989; 1990; Williams and 

others, 1989). Selenium also impairs growth and immune function of avocet chicks (Fairbrother and 

others, 1994). Kesterson has since been filled with uncontaminated soil in order to displace breeding 

birds to wetlands of better habitat quality, and habitat managers are monitoring potential 

bioaccumulation through the new terrestrial food chain (Wu and others, 1995). Selenium pollution has 

since been documented to be a widespread problem associated with irrigated agriculture in the western 

United States (Seiler and Skorupa, 1995). Normally, sets of eggs average <3 ppm selenium (dry 

weight), but at contaminated sites, they can average up to 86 ppm (Skorupa and Ohlendorf, 1991; See 

and others, 1992). 

Hoffman (2002) reviewed the role of selenium toxicity and oxidative stress in aquatic birds, 

including avocets. Below are a few examples of studies examining selenium and its adverse effects on 

avocets:

• In California, an exposure-response relationship existed between egg selenium 

concentration and incidence of embryo deformity (Robinson and others, 1997; Skorupa, 

1998). Eggs with selenium concentrations > 40 ppm dry weight had 0 percent incidence 

(based on 486 eggs) of deformed embryos; eggs with 41-60 ppm selenium had 5 percent

(based on 26 eggs) deformity, eggs with 61-80 ppm selenium had 8.3 percent (based on 

14 eggs) deformity, and eggs with 81-110 ppm had 25 percent (based on 15 eggs) 

deformity. Percentages were calculated by weighting results for each 5-ppm interval 

equally. Skorupa (1998) conducted a logistic regression analysis of the embryonic 

response for avocet eggs exposed to selenium and reported effect concentrations of 41, 

74, and 105 ppm selenium dry weight for 1 percent, 10 percent, and 50 percent rates of 

embryo deformities, respectively (for example, in toxicological nomenclature the EC50 

=105 ppm). These effect concentrations indicated that avocets were significantly less 

vulnerable to selenium poisoning than were black-necked stilts (Himantopus mexicanus), 

a closely related and ecologically similar species of shorebird.

• Selenium-contaminated evaporation ponds in Tulare Basin, California, are used by 

avocets migrating from uncontaminated sites in the western United States (Robinson and 

Oring, 1996). Some birds occupy these ponds for as long as 48 days, long enough for 

significant amounts of selenium to accumulate in their bodies (Heinz and others, 1990; 
11

Robinson and Oring, 1996). A potential for significant population-level impacts exists at 

the selenium-contaminated Salton Sea, California, where an estimated 100,000 avocets 

winter (Setmire and others, 1993). 

• In Wyoming, selenium levels of 38 to 54 ppb (parts per billion) in lakes were associated 

with high selenium levels in eggs (>8 ppm) and livers (>30 ppm) of adult avocets (both 

live and dead birds) (Skorupa, 1998; See and others, 1992). Of 102 avocet eggs, 

selenium concentrations ranged from 24.2 to 135 ppm dry weight and averaged 82.7 ppm 

dry weight. Mean egg concentrations of >13-24 ppm dry weight were associated with 

embryo deformities. No malformed embryos were found in the first year of the study, but 

in the second year 24 percent of 37 embryos had malformations. Crossed bills and 

malformed legs and feet were the most common types of deformities. Mean 

concentrations of >8 ppm dry weight were associated with impaired egg hatchability. 

Cause of selenium embryotoxicity was irrigation over soils derived from Cody shale, 

which tend to have a high selenium content. Selenium discharge from sub-basins was 

related to irrigation intensity (measured by the area of irrigated land) and the 

concentrations of selenium in groundwater. 

• At Freezout Lake Wildlife Management Area and Benton Lake National Wildlife Refuge 

in westcentral Montana, elevated levels of selenium were found in avocet embryos and 

livers (Nimick and others, 1996). Water sources for both lakes were irrigation return 

flow, surface runoff, and direct precipitation. At Freezout Lake, based on 73 embryos 

collected one year and 17 embryos collected the following year, geometric mean 

selenium concentrations ranged from 6 to 17 ppm. Livers from 16 avocets contained 

selenium concentrations that ranged from 13 to 43 ppm and averaged 22 ppm. At Benton 

Lake, based on 36 embryos collected one year and 10 embryos collected the following 

year, geometric mean selenium concentrations ranged from 3.4 to 3.7 ppm. Livers from 

28 avocets contained selenium concentrations that ranged from 9.3 to 45 ppm and 

averaged 20 ppm. 

Loss and Degradation of Wetlands

Wetland losses and conversion in North America (Dahl, 1990) have led to population declines of 

avocets (for example, Bent, 1927; 1962; Page and Gill, 1994). Many traditional breeding areas depend 

on vernal flooding, which is now rare because of agricultural diversion of in-stream flows and urban 

flood control projects (Robinson and others, 1997). Rangewide loss of historical perennial wetlands and 

coastal intertidal wetlands also has been severe (Helmers, 1992). Use of salt ponds, sewage ponds, 

agricultural evaporation ponds, and wet agricultural areas such as rice fields and flood-irrigated fields 

may counterbalance some historical wetland losses. These surrogate habitats, however, are not as likely 

as natural habitats to include predator-safe islands for nesting (Robinson and others, 1997). Avocets are 

also threatened by widespread agricultural and industrial degradation of water quality. Major breeding, 

staging, and wintering areas in the Pacific Flyway, such as the San Francisco estuary, the San Joaquin 

Valley, and the Salton Sea are already seriously polluted (Moore and others, 1990; Harvey and others, 

1992). 

Increased salinization of water from agricultural drain water, surface flow, and subsurface flow 

is a problem in at least some inland wetlands in every western state in the United States (U.S. Fish and 

Wildlife Service, 1992). Because of the multitude of ecosystem-level changes that salinization produces, 

salinization is likely to have significant impacts on populations of temperate-breeding shorebirds, 
12

including avocets (Rubega and Robinson, 1997; Hill, 1989). At highly salinized wetlands, avocet chick 

survival is likely reduced (Hannam and others, 2003; Robinson and others, 1997). 

The ability of avocets to colonize newly available habitats provides some protection from the 

effects of locally changing patterns of habitat availability, but it will not help the species cope with the 

effects of permanent habitat loss. Avocets show fidelity to their breeding areas and they show little 

tendency to disperse in response to nest failure. Because avocets use ephemeral wetlands and wet playas 

extensively, they are likely to be affected by changing precipitation patterns associated with climate 

change. 

Disturbance

At agricultural evaporation ponds, sewage ponds, and probably commercial salt ponds, routine 

grading of levees destroys many avocet eggs (Robinson and others, 1997). At one evaporation pond, 

160 of 769 eggs were lost in one season (21 percent) (Robinson and others, 1997). At another 

evaporation pond, 100 chicks died because they were trapped inside discarded motor vehicle tires 

placed at shorelines to stabilize earthen levees against erosion (Robinson and others, 1997). Wintering 

flocks of avocets are easily disturbed, taking flight in response to cars, trucks, motorcycles, mopeds, 

airplanes, airboats, motorboats, shooting, human activity, and dogs. At Humboldt Bay, California, levels 

of human disturbance were probably not as detrimental to avocets relative to natural disturbances by 

raptors (Evans, 1988). 

Research activities have been shown to negatively affect avocets. Nests visited early, when only 

one egg is present, have moderately high probability of being abandoned (Robinson and others, 1997). 

In Lassen County, California, 36 percent of 201 nests found with one egg were abandoned, while 9.5 

percent of 1,056 nests found with >1 egg were abandoned (Robinson and others, 1997). Abandonment 

rates were similar for nests in Tulare Basin, California (35 percent if found with one egg, 0 percent

otherwise). Nest-trapping and banding can also influence movement behaviors of individuals. For 

example, 16 percent (4 of 25) of incubating birds disappeared from a study area immediately after being 

captured and were resighted at migration sites (some within as few as 3 days; Robinson and Oring, 

1996). Nest-trapping and banding also affected dispersal distances in the year immediately after banding 

(Robinson and Oring, 1997). 

Management

Response to Management

Little is known concerning the direct effects of burning, mowing, or grazing on avocets (Dechant 

and others, 2002). In North Dakota, avocets nested along a burned shoreline (Sidle and Arnold, 1982). 

However, nests on this shoreline had lower hatching success (number of eggs hatched per nest) than 

nests on an island, a sandbar, and four peninsulas. In this location, preference for islands by breeding 

avocets may have been due to greater protection from nest predators (Sidle and Arnold, 1982; Giroux, 

1985). Because avocets prefer islands for nesting, prescribed burning on islands during the breeding 

season is usually not recommended because fire eliminates all suitable nesting cover such as tall weeds, 

grasses, or low shrubs (Lokemoen, 1993). 

Gradual drawdowns of water levels in wetlands, mowing, burning, and shallow disking have 

been suggested to be beneficial to migrating shorebirds, including avocets. Davis and Smith (1998) 

studied the ecology and management of migrant shorebirds in the Playa Lakes Region of Texas. They 

suggested that for migrant shorebirds, the focus should be on creating and maintaining habitats 

characterized by sparse vegetation cover, and an adequate area of mudflat (at least 10-15 percent) and 
13

shallow water (at least 10-20 percent). Because invertebrates are important foods for migrant shorebirds, 

management should focus on enhancing invertebrate populations in playas (Eldridge, 1992). Gradual 

drawdowns of playas with deep water and flooding of dry playas were suggested to provide available 

habitat for shorebirds throughout the migration period. If vegetation cover in playas is dense (>50 

percent), mowing, shallow disking (for example, partially bury plant biomass), or controlled burning 

could be used to reduce vegetation cover and provide a detrital food base to enhance invertebrate 

populations (Davis and Smith, 1998). Aquatic invertebrates increase when wetlands are fertilized by 

mowing and grazing, but water control impoundments are easier management tools (Eldridge, 1992). 

The proper regime of drawdown and flooding can stimulate plant growth and decomposition and create 

a detrital food source for invertebrates. When water is drawn down slowly (2 to 4 cm per week) during 

the appropriate times of the year, migrating shorebirds are attracted to the available invertebrates. 

Complete drawdowns during migration, however, especially of semi-permanent wetlands, are 

discouraged because invertebrates are reduced or eliminated (Swanson and Meyer, 1977). 

Agricultural drainage can be successfully used to provide highly productive managed wetlands 

for avocets (and black-necked stilts)(Davis and others, 2008). In the Tulare Lake Drainage District, 

California, operators of selenium-contaminated evaporation ponds are required via Waste Discharge 

Permits to provide mitigation habitat targeted for breeding recurvirostrids (California State Water 

Resources Control Board, 1995). In a 10-year study of this area, a 117-ha (hectare) wetland was 

designed, constructed, and operated to provide foraging and nesting habitat for avocets (Davis and 

others, 2008). The design for this constructed wetland incorporated a 2:1 ratio of foraging habitat to 

nesting habitat and water depths of 10-15 cm. The constructed wetland included low-profile lanes 

(shoreline) immediately adjacent to extensive shallow-water areas to support macroinvertebrate 

production and provide extensive foraging habitat. There were 34 lanes, each approximately 1.6 km 

long, providing more than 59 km (kilometers) of nesting habitat approximately 10 m wide and wetted 

foraging areas approximately 20 m wide, with shoreline slope of 10-12:1 (horizontal: vertical). Water 

supply for the wetland was from low-selenium saline agricultural drainage water, or a blend of 

freshwater and saline supplies. Selenium concentrations and electrical conductivity were monitored. 

Invertebrate breeding ponds were created throughout the interior of the wetland habitat. The wetland 

was encircled by an electrified perimeter fence to inhibit large mammalian predators (for example, 

coyotes). Coincident with the wetland construction, 2 evaporation basins totaling 1,174 ha were 

physically modified, and hazing was implemented to discourage use by shorebirds. In the first year of 

wetland operation (1995), avocet and black-necked stilt nest construction density was 17.6 nests/ha. 

Before the mitigated wetland went into operation (1994), there was a combined density of only 1.9 

nests/ha at the evaporation basins. From 1995 to 2004, nesting attempts by avocets and stilts at the 

wetland averaged 2,896 per year (24.8 nests/ha). Nest success at the constructed wetland averaged 82 

percent for avocets and nest predation was estimated to be <1 percent, whereas predation at the 

evaporation ponds exceeded 30 percent. 

Recommendations

Fluctuations in water levels can place avocet nests at risk during the breeding season (Plissner 

and others, 2000); careful consideration of human-caused changes in water levels, or construction of 

nest ridges to elevate nests, may be necessary to prevent or at least minimize nest losses. For breeding 

avocets, elevated nest structures, such as ridges, and exclosures made of electrical fencing may increase 

nesting success of avocets by reducing loss of nests due to flooding and nest predation, respectively 

(Koenen, 1995; Winton and Leslie, 1997). Nest losses due to flooding and predation may be 

considerable. Of 193 avocet nests in California, 84 (44 percent) were lost to predation and 11 (6 
14

percent) to changes in water level (Ohlendorf and others, 1989). In Oklahoma, exclosures and nest 

ridges were used to reduce nest predation and flooding, respectively, on snowy plovers and least terns 

(Koenen, 1995). Although the effects of exclosures and ridges on avocet nest success were not 

examined, avocets did nest in proximity to the plovers and terns. Fluctuations in water levels may cause 

the destruction of nests because of the proximity (usually <60 m) with which avocets place their nests to 

water (Plissner and others, 2000). Nest ridges were 10 m long by 1 m wide by 0.5 m high and were 

spaced 20-30 m apart. Electric fences (1,000-6,000 volt) were used to form predator exclosures ranging 

in size from 3.75 to 24 ha. Five strands of wire were placed from 14 to 86 cm from the ground. 

To reduce risk of avian contamination and toxicity due to drainage and irrigation water, either 

the amount of contaminants in drainage water should be reduced, or avian use of contaminated ponds 

should be reduced (Skorupa and Ohlendorf, 1991; Skorupa, 1998). To prevent avian toxicity but not 

avian contamination, drainage water should be purified to <10 ppb waterborne selenium. However, as 

these standards cannot be met with current technology, deterring avian use of contaminated drainage 

water may be the only option. Drainage water containing 3-20 ppb selenium should be considered 

potentially hazardous to waterbirds, whereas concentrations >20 ppb should be considered hazardous. 

Power lines should not be constructed through or within 1 km of known historical high-water 

marks of wetlands, through dry basins known to hold water intermittently, or through heavily used 

waterbird migration routes (Malcolm, 1982). In cases where power lines must cross wetlands, an 

attempt should be made to hide the lines with structures such as bridges (McKenna, 1976). Power lines 

should be buried where possible and/or corridors established where power lines can be clumped to 

reduce their negative effects on shorebirds (McKenna, 1976). 

The Central Plains/Playa Lakes Regional Shorebird Conservation Plan (Fellows and others, 

2001) suggested that management for shorebirds should focus on enhancing invertebrate populations 

and maintaining sparse vegetative cover and adequate mudflat (at least 10-15 percent) and shallow 

water (at least 10-20 percent) habitats throughout the period used by the birds. Habitat can be managed 

through gradual drawdowns of playas with deep water and flooding of dry playas where water 

movement capabilities exist. Mowing and shallow discing can be used to create preferred habitat 

conditions and provide a detrital food base to enhance invertebrate populations. Variations of current 

moist-soil management practices used by waterfowl managers (Frederickson and Taylor, 1982; Haukos 

and Smith, 1993) may also benefit shorebirds, especially in the spring (Fellows and others, 2001). 

The Nevada Conservation Strategy for the American avocet 

(http://www.gbbo.org/pdf/bcp/32_American%20Avocet.pdf, Accessed 1 July 2011) provided the 

following habitat strategies: 

• Promote seasonal fresh-water runoff into ephemeral wetlands and playas, as well as into 

sparsely-vegetated permanent marshes, sufficient to create mudflats and maintain a 

shallow-water shoreline for the longest possible period. 

• Manage and restrict playa activities to protect the integrity of the clay soil pan and 

maximize water retention. 

• Promote wetlands with uneven bottoms and shallow islands, which are especially 

important for nesting. 

• Artifically mitigate wetlands to provide productive breeding habitat. Ideal configuration 

is >130 ha wetland with a 2:1 ratio of shallow water (<15 cm) feeding areas that 

transition to elevated nesting areas. 
15

Research Priorities

No species-specific research priorities for avocets were found during this literature search; 

however, the U.S. Shorebird Conservation Plan (Oring and others, 2000) provided the following priority 

research topics: 

• Identification of population limiting factors such as those influencing breeding success, juvenile 

and adult mortality, and disturbances that affect breeding success and mortality; 

• Understanding patterns of distribution and abundance, and the factors controlling them; 

• Space use and dispersal, including movements within and among years; 

• Increasing understanding of the dynamics of migration patterns, including how and why 

populations move among sites; 

• Understanding the timing of landscape level habitat use, which is critical for monitoring studies, 

and understanding factors affecting turnover rates; 

• Analysis of dietary requirements, elucidation of dietary preferences, nutritional requirements, 

and metabolic needs; 

• Identifying geographic population subdivision, determining conservation issues below the 

species level, and identification of the role of subspecies in the overall population dynamics of 

the species;

• Modeling potential impacts of an development of protocols to mitigate effects of global climate 

change;

• Development of techniques to deter predators, for example, including aversive conditioning to 

reduce losses from both avian and mammalian predators; 

• Disturbance effects on foraging and breeding, including measuring disturbance impact and/or 

studying ways of reducing disturbance;

• Techniques for increasing productivity, including captive breeding reintroduction and associated 

techniques; 

• Development of models to predict population trends; 

• Exploration of population fluctuations, and their impact on estimates of numbers using sites; and

• Evaluation of the validity of sampling techniques. 

 

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27

II. Wilson’s Phalarope 

(Phalaropus tricolor) 

Introduction

Wilson’s Phalarope (Phalaropus

tricolor) is the largest and most terrestrial of 

the three phalarope species (Colwell and Jehl, 

1994). All three species of phalaropes are 

known for their reversed sex-role mating 

system. The larger and more brightlyplumaged females compete for mates and are 

sometimes polyandrous, whereas males 

provide all the parental care (Johns, 1969). 

Unlike the other two phalarope species [Rednecked Phalarope (P. lobatus) and Red Phalarope (P. fulicarius)], which have Holarctic breeding ranges 

and pelagic nonbreeding distributions, Wilson’s phalaropes breed exclusively within the Nearctic, and 

its nonbreeding distribution is entirely continental. Wilson’s Phalaropes breed at shallow wetlands of 

interior western North America, but for most of the year the species is a salt-lake specialist (Colwell and 

Jehl, 1994). After the breeding season, virtually all adults undertake a molt migration and stage, often in 

huge flocks, at hypersaline/alkaline lakes of western North America, before migrating to similar 

wintering habitats mainly in the altiplano of Bolivia and Argentina. Small numbers winter at wetlands in 

the pampas and as far south as Tierra del Fuego. Relatively little is known about habitats in wintering 

areas for this species. Loss of prairie wetlands in the early to mid-twentieth century reduced overall 

habitat for this phalarope, and likely its numbers as well. However, its range has expanded dramatically 

in recent decades, with breeding occurring from regions as far east as Massachusetts, as far south (and 

west) as New Mexico, and as far north as Alaska and the Yukon (Colwell and Jehl, 1994). 

A simple Google search on the Internet for “Phalaropus tricolor” revealed 167,000 results and a 

Google Scholar search provided 1,180 results (accessed 11 July 2011). “Wilson’s phalarope” revealed 

150,000 results through Google and 2,020 results through Google Scholar (accessed 11 July 2011). 

Natural History

Morphology

Wilson’s phalaropes are small (22 to 24 cm in length), aquatic sandpipers with toes bordered by 

a narrow fleshy fringe, needle-like bills, and white underparts and rumps (Colwell and Jehl, 1994). 

During the breeding season, the larger females are also more brightly plumaged than the males. 

Breeding females possess pale blue-gray foreheads and crowns; a white streak trails from the back of 

the head, down the nape to the upper back; the throat is cinnamon buff-colored; scapulars and mantle 

are chestnut; the wings are gray brown; underparts and rump are white; the tail is pale grayish; bill is 

black; and legs are grayish to black (Colwell and Jehl, 1994). Breeding males are smaller in overall size 

and less brightly plumaged than females, with considerable individual variability. The sexes are similar 

in basic (nonbreeding) plumage (Colwell and Jehl, 1994). 

Wilson’s Phalarope, courtesy of Eleanor Briccetti.
28

Distribution

Wilson’s phalaropes breed from the southern Yukon Territories through British Columbia, 

southcentral Alberta, and southern Manitoba, south to central California, southern Nevada, southern 

Colorado, northern New Mexico, and northern Texas, and east to central Kansas, northwestern Iowa, 

and northwestern Minnesota (Dechant and others, 2002). They also breed from eastern Wisconsin and 

northeastern Illinois, east to Michigan, northern Ohio, eastern Ontario, and northern New York 

(National Geographic Society, 1987). An expanding population breeds around the Great Lakes and 

recently reached southwest Quebec (first recorded breeding in 1974). Outside the usual breeding range, 

isolated nests have been recorded from many areas, including southeastern Alaska east to James Bay, 

Ontario, and New Brunswick, south to central California, Arizona, New Mexico, Texas, Ohio, and 

Massachusetts. Recently, phalaropes have been suspected of nesting in Delaware, New Jersey, and New 

York (Veit and Peterson, 1993). 

In Colorado, phalaropes summer across much of the state, although their breeding status is not 

always easy to determine (Nelson, 1998). They prefer to nest in moist sedge and rush meadows 

characterized by low plant height (Nelson, 1998). Low sedges may grow in highly alkaline soil. 

Adjacent open water is also a habitat requirement (Bent, 1927). In Colorado, the known range in the late 

1800s restricted phalaropes to the northeast quarter of the state (Cooke, 1897), but by 1965 they were 

regarded as common nesters not only in prairie marshes but also mountain valleys. That situation has 

changed; the current Colorado Breeding Bird Atlas map indicates they nest in scattered places across 

much of the state, but nesting has subsided somewhat in the plains where many marshes have been 

converted to wheat fields, corn patches, or subdivisions. Colorado nesting habitat currently centers on 

the intermountain valleys at 7,000 – 10,000 ft (2135-2048 m). Nesting strongholds during the Atlas 

surveys included North Park, San Luis Valley (SLV), Gunnison Valley, and the Yampa watershed. 

Nest-finding for this species was considered exasperating by Bent (1927). Males sit tight when 

incubating, and nests avoid human detection unless the male flushes at the last possible moment. Young 

stay in heavy vegetation until old enough to fly, and shortly after fledging depart the breeding grounds. 

Six out of ten breeding confirmations from the Colorado Breeding Bird Atlas were from Alamosa, 

Monte Vista, and Arapahoe NWRs. 

Migration

Wilson’s phalaropes are transequatorial migrants (Colwell and Jehl, 1994). They are one of two 

shorebird species to undergo molt migration and they are one of the earliest fall migrants (Jehl, 1987; 

1990). Their post-breeding migratory movements are staggered by sex: females depart the breeding area 

and arrive at staging areas by mid-June, followed by adult males, and finally, the juveniles (Colwell and 

Oring, 1988a; Jehl, 1988). The southbound migration is characterized by large numbers of adults 

staging at hypersaline lakes (for example, the Great Salt Lake, Utah; Mono Lake, California) of western 

North America, where abundant food enables them to rapidly molt and store fat for migration. Wilson’s 

phalaropes amass greater fat loads (up to 54 percent) than the maximal 45 percent fat load necessary for 

most shorebird migrants (Jehl, 1997). For a brief period of time, they can become too heavy to fly. 

Southward migration of adults is characterized by a rapid and direct nonstop flight from the staging 

areas in the United States to coastal, western South America. The adults evidently use a “Great Circle 

Route” across the Pacific as indicated by a virtual absence of fall records of phalaropes from Central 

America and northern South America. Juvenile movements are slower and cross a broader overland 

front in western and central United States, through northern Mexico, and Central America to northern 

South America. Adults arrive in South America in early August after a flight of approximately 54 hours. 

Juveniles arrive in late August. The spring migration northward is largely overland and through the 
29

highlands of South America in mid-March, through Central America or across the Gulf of Mexico, 

through the intermountain west and southern plains (Oring and Davis, 1966; Skagen and Knopf, 1993), 

where 90 percent of the population is thought to pass through Cheyenne Bottoms in Kansas (Morrison 

and Myers, 1987) in some years. Breeding areas are first occupied in late April to early May. 

Habitat

During the breeding season, Wilson’s phalaropes use both fresh and alkali wetlands with three 

characteristics: open water, emergent vegetation, and open shoreline (Saunders, 1914; Stewart and 

Kantrud, 1965; Hohn, 1967; Stewart, 1975; Prescott and others, 1995; Naugle, 1997). Nesting habitat 

varies widely, including wetlands, wet meadows, upland grasslands, and road rights-of-way (Roberts, 

1932; Bent, 1962; Hohn, 1967; Stewart, 1975; Murray, 1983; Bomberger, 1984; Colwell, 1987; Colwell 

and Oring, 1990; Einemann, 1991; Faanes and Lingle, 1995; Dinsmore and Schuster, 1997). Wilson’s 

phalaropes occasionally occur in Conservation Reserve Program (CRP) fields and dense nesting cover 

(Johnson and Schwartz, 1993; Prescott and others, 1993). In North Dakota, Wilson phalarope densities 

were highest in undifferentiated tillage wetlands (wetlands with frequently tilled soils), followed by 

temporary, seasonal, semipermanent, fen, alkali, and permanent wetlands (Kantrud and Stewart, 1984). 

Wilson’s phalaropes often occupied the peripheral low-prairie and wet-meadow areas of most classes of 

wetlands in North Dakota. In South Dakota, the occurrence of Wilson’s phalaropes was associated 

positively with the presence of seasonal and semipermanent wetlands, stock ponds, and intermittent 

streams; area of alfalfa (Medicago sativa) hayland; area of surface water; and the percentage of grazed 

shoreline (Weber, 1978; Weber and others, 1982). In eastern South Dakota, the probability of 

occurrence of Wilson’s phalaropes in semipermanent wetlands was related positively to the proportion 

of untilled uplands and the number of emergent hydrophyte species (for example, willow [Salix spp.]) 

composing >10 percent of the vegetated wetland area; Wilson’s phalaropes were associated negatively 

with wetlands dominated by thick-stemmed plants (for example, cattail [Typha spp.] and river bulrush 

[Schoenoplectus fluviatilis]) (Naugle, 1997). Within seasonal wetlands, the probability of occurrence of 

Wilson’s Phalaropes was related negatively to wetlands dominated by thick-stemmed plants (Naugle, 

1997). In Colorado, Wilson’s phalaropes preferred seasonal wetlands and habitats dominated by baltic 

rush (Juncus balticus), sedges (Carex spp.), and grasses <40 cm tall over semipermanent wetlands, 

habitats dominated by cattail and softstem bulrush (Schoenoplectus tabernaemontani) >40 cm tall, 

saltgrass habitats, or upland shrub habitats (Laubhan and Gammonley 2000). 

Nest site selection varies seasonally for this species. Wilson’s phalaropes nest in upland 

vegetation early in the breeding season and wet-meadow vegetation later in the season (Colwell and 

Oring, 1990). They usually nest <100 m from shoreline (Hohn, 1967; Hatch, 1971; Colwell and Oring, 

1990; Eldridge, 1992). They also exhibit annual variation in nest site selection, moving to deeper, more 

permanent wetlands in dry years (Hohn, 1967; Colwell, 1991). Nest sites in Nebraska were in wet sedge 

meadows (Faanes and Lingle, 1995). In North Dakota and Iowa, Wilson’s phalaropes nested in wetlands 

associated with river floodplains (Murray, 1983; Koenig, 1984). In Alberta, Saskatchewan, and North 

Dakota, the birds nested in grasses of various heights on islands or in wet-meadow zones around lakes 

and wetlands; in Saskatchewan, brood rearing occurred in patches of foxtail barley (Hordeum jubatum) 

(Bent, 1962; Hohn, 1967; Kagarise, 1979; Colwell 1987). In Saskatchewan, Colwell and Oring (1990) 

found that nest sites of Wilson’s phalaropes had taller, denser, and more homogeneous vegetation and 

less bare ground than randomly selected sites. In the Nebraska sandhills, however, nest sites had shorter 

vegetation than random sites (Bomberger, 1984). A table near the end of this account lists the specific 

habitat characteristics for Wilson’s phalaropes by study. 
30

Naugle and others (1999) conducted bird surveys throughout South Dakota in 1995 and 1996 to 

assess the effects of woody vegetation encroachment on nongame wetland birds. They found that 

wetland bird species richness decreased as the extent of woody vegetation encompassing wetland 

perimeters increased. Logistic regression analyses indicated that Wilson’s phalaropes were less likely to 

occur in wetlands surrounded by trees. 

Foraging Habits

The main foods eaten by Wilson’s phalaropes are small aquatic invertebrates (dipterans, 

crustaceans, particularly brine flies and brine shrimp) in freshwater or hypersaline environments 

(Colwell and Jehl, 1994). They also feed on some terrestrial invertebrates and occasionally on seeds of 

aquatic plants (Colwell and Jehl, 1994; O’Brien and others, 2006). Wilson’s phalaropes are the most 

terrestrial of the 3 phalarope species, but are still highly aquatic (Baldassare and Fischer, 1984; Colwell 

and Oring, 1988b). On the breeding grounds they forage in open-water, flooded meadows, and less 

frequently in upland habitats and long beaches (Colwell, 1987; Colwell and Oring, 1988b). Males use a 

wide array of aquatic and terrestrial habitats whereas females are more aquatic (Colwell, 1987). 

Wilson’s phalaropes often forage in close proximity (commensally) with other aquatic species, 

including the northern shoveler (Anas clypeata; Siegfried and Batt, 1972), American avocet (Williams, 

1953; Pinkowski, 1981), Chilean flamingoes (Phenicopterus chilensis; Hurlbert and others, 1984), and 

blue-winged teal (Anas discors).

Unlike most shorebirds, Wilson’s phalaropes forage principally while swimming. The birds 

forage visually, except when probing soft substrates, and use the surface tension of water between their 

mandibles to transport prey from the beak tip to their mouths (Rubega and Obst, 1993; Rubega, 1997; 

Estrella and others, 2007). Foraging methods include typical phalarope “spinning,” chasing and pecking 

prey from water or mud surfaces (Wetmore, 1925), standing still and lunging/stabbing passing flies, 

probing soft substrates, and rarely, “hawking” insects (Benning, 1974) and scything in the manner of 

avocets. Spinning behavior is rarely used at Mono Lake and other major staging areas where food is 

abundant. Foraging behavior varies with location, season, and environmental conditions, which 

influence the available prey (Colwell and Jehl, 1994). 

Breeding

In the central and northern Great Plains (Minnesota, Nebraska, and North Dakota), Wilson’s 

phalaropes arrive on the breeding grounds from mid-April to early May and depart from mid-August to 

early September (Roberts, 1932; Howe, 1972; Johnsgard, 1980; Murray, 1983). In Alberta, Manitoba, 

and Saskatchewan, Wilson’s phalaropes arrive on the breeding grounds from late April to early May and 

are observed until early September (Hohn, 1967; Maher, 1974; Reynolds and others, 1986; Colwell, 

1987; Colwell and Oring, 1988a; 1988b). Females arrive on the breeding grounds earlier than males

(Reynolds et al. 1986, Colwell 1987), and commonly depart from breeding areas earlier than males, 

usually from early June to early July (Hohn, 1967; Howe, 1972; Colwell, 1987; Colwell and Oring, 

1988a; 1988b). 

Most of the information on breeding biology is based on a 6-year banding study in 

Saskatchewan near the northern extent of the breeding range (Reynolds and others, 1986; Colwell and 

Oring, 1988a; 1988b; 1988c; 1988d; 1988e; 1988f; Colwell and Oring, 1990; Colwell, 1992). Early 

season female-biased sex ratios and intense competition among females for mates suggest that birds 

arrive unpaired. The duration and types of interactions between pair members vary greatly. The average 

interval between pair formation and clutch completion is 8 days (range 7-10). About half (16/29) of the 

females desert mates upon clutch completion. 
31

Wilson’s phalaropes may renest after nest failure, and females are capable of laying multiple 

clutches (Colwell and Jehl, 1994). Incubation and brood-rearing is performed exclusively by males 

(Delehanty and others, 1998). Polyandry was first documented in the species in Saskatchewan, where a 

color-banded female laid two clutches with two individual males (Colwell, 1986a; Colwell, 1987). 

Philopatry is uncommon in Wilson’s phalaropes, although males return to breeding areas in successive 

years more often than females (Colwell, 1987; Colwell and Oring, 1988b). Of 154 adult male 

phalaropes banded over 4 yr in Saskatchewan, 16 percent returned to their previous breeding area in 

successive years, whereas only 2 percent of 69 banded adult females returned (Colwell 1987). 

The Wilson’s phalarope is an accidental and unsuitable host of the brown-headed cowbird 

(Molothrus ater), an obligate brood parasite (Friedmann, 1963; Hatch, 1971). In North Dakota, 0 of 20 

nests were parasitized (M. Winter and D. H. Johnson, unpublished data, cited in Colwell and Jehl, 

1994). In Saskatchewan, 1 percent of 386 nests were parasitized (Colwell and Jehl, 1994). One record of 

multiple parasitism has been reported (Friedmann, 1963). One case of intraspecific brood parasitism 

was reported by Colwell (1986a) in Saskatchewan. 

Demography & Population Status

Data on longevity of Wilson’s phalaropes is restricted to individuals returning to a Saskatchewan 

site where they were banded (Colwell and Oring, 1988f). Most males (87 percent, n = 55) and females 

(88 percent, n = 8) were observed breeding for one year after banding as adults, 7 percent of males and

12 percent of females were present in two consecutive years, and 6 percent of males and no females 

were present for three years. These estimates were probably low because of dispersal to new, 

unmonitored breeding sites (Colwell and Jehl, 1994). Many other sandpiper species live 8-10 years in 

the wild (Colwell and Jehl, 1994). No estimates of annual survival for Wilson’s phalaropes exist, 

according to the extensive literature search conducted for this project.

There is considerable temporal and spatial variation in breeding densities of Wilson’s phalaropes 

owing to use of ephemeral wetlands and a nonterritorial mating system. Estimated nest densities vary 

between 0 and 45 at a 40 ha ephemeral wetland, and 55-110 at a 100-ha permanent wetland in 

Saskatchewan (Colwell, 1986b; 1991; Colwell and Oring, 1988a). Breeding densities may be lower than 

this because of renesting by some males. 

Several population estimates for Wilson’s phalaropes have been published. Jehl (1988) estimated 

North American populations at 1.5 million birds in fall, based upon counts from major staging areas. 

Jehl’s estimate was based on a total count of 741,000 individuals (primarily adults) reported from all 

major known staging sites in North American in July 1986. Stewart and Kantrud (1972) estimated 

172,000 breeding pairs in North Dakota. The great majority of adult Wilson’s phalaropes is thought to 

stage at the Great Salt Lake, Utah, Mono Lake, California, and/or Lake Albert, Oregon, before making a 

nonstop flight to South America. Their concentrations at these few localities provides an exceptional 

opportunity for studying trends species-wide (Jehl, 1999). Observations from 1980 through 1997 

indicated that numbers of staging adults approximated 500,000-600,000 in the 1980s, then declined by 

about half in the 1990s. 

Currently, it is unclear whether Wilson’s phalarope populations are increasing, declining, or 

stable (Lesterhuis and Clay, 2010). While its breeding range has considerably expanded in recent 

decades, the species no longer breeds at a number of former sites, and the population has not shown a 

marked increase. At the national level, Wilson’s phalaropes have been considered a “Species of High 

Concern” in the United States and Canadian Shorebird Conservation Plans (Donaldson and others, 

2000; Brown and others, 2001; USFWS, 2004). The Breeding Bird Survey (BBS) regional trend 

analysis for Wilson’s phalaropes along 35 Colorado routes from 1966 to 1999 showed a negative trend 
32

estimate of -1.2 (-5.0, 2.7; 95 percent CI) (Sauer and others, 2011). From 1999 to 2009, the trend was 

also negative and estimated to be -0.2 (-6.9, 9.0; 95 percent CI). However, the regional credibility 

measure for this species was in the “yellow” category meaning that data have deficiencies such as 

overall low abundances of phalaropes per route and/or the results were too imprecise to detect long-term 

trends. 

Threats

Sources of Mortality

The potential predators of eggs and young of Wilson’ phalaropes include garter snake 

(Thamnophis spp.), various gulls (Larus spp.), common crow (Corvus brachyrhynchos), red-winged 

blackbird (Agelaius phoeniceus), yellow-headed blackbird (Xanthocephalus xanthocephalus), striped 

skunk (Mephitis mephitis), Franklin’s ground squirrel (Spermophilus franklini), Richardson’s ground 

squirrel (S. richardsoni), and raccoon (Procyon lotor) (Kagarise, 1979; Colwell and Oring, 1988). 

Clutch mortality is predominately due to predation (average 59 percent) (Colwell and Jehl, 

1994). The next highest losses for clutches (average 17 percent) are due to abandonment, which often 

follows partial clutch loss to predators (Colwell and Oring, 1988a; Delehanty and Oring, 1993). 

Predation rates for nests are highest during the years of greatest human activity around nests (Colwell, 

1992). In Saskatchewan, there was no association between nest success and degree of concealment from 

predators provided by vegetation in the immediate vicinity of the nest cup. 

The responses of Wilson’s phalaropes to predators vary seasonally. They are fairly tolerant of 

humans on nesting grounds and shortly after arriving at staging lakes prior to weight gain for migration. 

However, they are very wary of humans and gulls shortly before departure to South America, when they 

are obese and less mobile (Colwell and Jehl, 1994). The major diurnal predators of phalaropes at salt 

lake staging areas include peregrine falcons, prairie falcons, and northern harriers.

Contaminants

Wilson’s phalaropes are found in close proximity to agricultural areas, especially during the 

breeding season, where the widespread use of agrochemicals and irrigation water may be causing 

unrecognized mortality or other problems (Lesterhuis and Clay, 2010). Agricultural practices can result 

in elevated sedimentation rates (Martin and Hartman, 1986; Gleason and Euliss, 1996), and drift of 

agricultural chemicals into wetlands (Grue and others, 1989), large inputs of nutrients (Neely and Baker, 

1989), unnatural variance in water-level fluctuation (Euliss and Mushet, 1996), and altered vegetative 

communities (Kantrud and Newton, 1996). 

In Wyoming, high levels of selenium in lakes appeared to lead to high levels in the eggs (>13 

μg/g) and livers (>30 μg/g) of adult Wilson’s phalaropes (See and others, 1992). One adult, dead bird 

had a liver selenium content of >30 μg/g, a level associated with biological risk. Mean concentrations of 

selenium >13 μg/g dry weight were associated with embryo deformities. Of six eggs collected over 2 

years, selenium concentrations ranged from 5 to 19.9 μg/g dry weight and averaged 11.7 μg/g dry 

weight. Irrigation of soils with a high selenium content caused leaching of selenium from the soil to the 

ground water. Selenium discharge from basins was related to intensity of irrigation (measured by the 

area of irrigated land) and the concentration of selenium in the ground water. Also in Wyoming, 

mortality of Wilson’s phalarope was observed after fenthion, a chemical used to control mosquitoes 

(Culicidae), was aerially applied at a rate of 47 g/ha to an irrigated meadow (DeWeese and others, 

1983). Fenthion is a cholinesterase inhibitor, and activity of brain cholinesterase was significantly lower 

for 15 days post spray in phalaropes collected from treated areas than in phalaropes from control areas. 
33

Loss and Degradation of Wetlands

The contraction of breeding range for the Wilson’s Phalarope in the early 20th century is likely 

linked to the loss of prairie wetlands (Dahl and Johnson, 1991). For example, in Iowa there has been an 

89 percent loss in wetlands since the 1870s. The recent range shifts for the Wilson’s phalarope suggests 

this species is sufficiently vagile in discovering new habitats (Skagen and Knopf, 1993). In some areas 

of western North America, water diversion or reclamation projects are changing the ecology of 

hypersaline lake staging areas, which could have a potentially negative affect on phalaropes (Colwell 

and Jehl, 1994; Jehl, 1994). The main reason for the loss of habitat on breeding grounds has been 

hypothesized to be the conversion of most of the North American prairie grasslands and associated 

wetlands to agriculture (tillage)(Lesterhuis and Clay, 2010). Habitat loss and/or degradation of habitat 

during migration has also been implicated in declines (Lesterhuis and Clay, 2010). Wilson’s phalaropes 

are particularly susceptible to habitat loss and degradation at the few sites that hold significant molt 

congregations prior to the main migratory movement to wintering grounds (for example, up to 50 

percent of the population is believed to congregate at the Great Salt Lake, Utah). Extraction of water for 

agricultural use and the degradation of water quality at these sites are of particular concern. The 

majority of the population is believed to winter in the saline lakes of the altiplano in Argentina, Bolivia, 

and Peru, where mining activities are affecting the quality of water and the extent of saline lagoons. 

Large concentrations of phalaropes also are known to winter in saline lakes on the Argentinean Pampas 

and the Central Chaco (Argentina and Paraguay), where ever-expanding agriculture and increasing 

severity of drought are significant threats. 

In Minnesota, a possible reason for declines in Wilson’s phalaropes was postulated to be fire 

suppression (Hanowski and others, 1999). Fire suppression increased the shrub vegetation and 

subsequently decreased open grass/sedge areas in wetland ecosystems that phalaropes require for 

nesting. 

Climate Change

The species could be particularly vulnerable to the effects of climate change on its breeding 

habitat (Lesterhuis and Clay, 2010). Parts of the prairies are predicted to become drier, and drought has 

already been documented as main reasons for abandonment of breeding areas. On the other hand, other 

areas of prairie are predicted to become wetter; however, it is uncertain whether these areas will provide 

suitable nesting habitat for the species. The future availability of water in saline lakes and Southern 

Cone lowlands in the nonbreeding areas is also a cause for concern. For example, it has been shown that 

drought at Mar Chiquita, a large saline lagoon in Argentina, negatively affects the number of Wilson’s 

phalaropes.

Management

Response to Management

Wilson’s phalaropes nest in idle, hayed, and grazed grasslands adjacent to wetlands (Hohn, 

1967, Kantrud and Higgins, 1992). Management of grassland can create essential upland habitat for 

breeding shorebirds, including Wilson’s phalaropes, through grazing, mowing, or prescribed burning. 

Before European settlement, breeding shorebirds specialized in exploiting the grassland mosaics created 

by bison herbivory and fires (Eldridge, 1992). Prescribed burning can improve nesting habitat for all 

shorebirds (Eldridge, 1992). 
34

In North Dakota, Wilson’s phalaropes nested at higher densities in hayland mowed the previous 

year than in grazed areas (Kantrud, 1981). Idle grasslands and previously grazed areas provided habitat 

for nesting, but areas with cattle present during the breeding season were less suitable (Renken, 1983; 

Renken and Dinsmore, 1987; Kantrud and Higgins, 1992). In Alberta, Wilson’s phalaropes were present 

in deferred-grazed (grazed after 15 July) native pasture (Prescott and others, 1993). Nesting occurred in 

areas that were moderately grazed in Nebraska (Faanes and Lingle, 1995) and heavily grazed in 

Saskatchewan (Colwell, 1987). Although Wilson’s phalaropes occasionally nested in cropland (smallgrain stubble) in North Dakota (Higgins, 1975), native grassland was preferred over cropland and tame 

grassland in southern Canada and the northern United States (Owens and Myres, 1973; Eldridge 1992; 

Kantrud and Higgins, 1992). In the northern Great Plains, Wilson’s phalaropes favor Conservation 

Reserve Program (CRP) grassland blocks >100 ha in size (D. H. Johnson, unpublished data). Johnson 

and Schwartz (1993) reported that Wilson’s Phalaropes were present in low numbers in CRP fields in 

the northern Great Plains (North Dakota, South Dakota, and eastern Montana). In Saskatchewan aspen 

parkland, Wilson’s phalaropes were observed in dense nesting cover that contained wetlands (Prescott 

and others, 1993; 1995). 

Tillage agriculture is expanding into western prairie landscapes and the effects of this land-use 

change on habitats used by nongame wetland birds are unknown (May and others, 2002). From 1999-

2000, May and others (2002) surveyed 196 stock ponds within grassland (>95 percent grass) and 

cropland (>75 percent tillage) landscapes to evaluate effects of land use on nongame wetland bird 

densities in western South Dakota. Abundance of Wilson’s phalaropes was greater in grassland study 

areas where cattle grazing limited growth of thick-stemmed emergent vegetation and reduced the overall 

vegetative cover in stock ponds. 

In Montana, instances of Wilson’s phalaropes fatally colliding with a power transmission line 

were reported (Malcolm, 1982). Highest mortalities occurred in August and September during fall 

migration period. The power transmission line structure was constructed over a wetland that was 

intermittently flooded. The structure consisted of steel towers that supported six pairs of 25-mm 

diameter conductor wires in two layers below two static or lightning interceptor wires of 14-mm 

diameter. Distances from the water to the conductor wires ranged from 14 to 33 m. 

Recommendations

Several specific management actions and considerations pertinent to Wilson’s phalaropes were 

suggested by Dechant and others (2002): 

• Protect wetland complexes with both seasonal and semipermanent wetlands to provide 

suitable habitat during both wet and dry years (Kantrud and Stewart, 1984; Colwell and 

Oring, 1988c). Wilson’s phalaropes exhibit annual variation in nest site selection, moving to 

deeper, more permanent wetlands in dry years (Hohn, 1967; Colwell 1991). 

• Ensure the presence of wet-meadow areas near deeper wetlands during the breeding season 

(Colwell and Oring, 1988c). This may make it easier for adults to move young from nests to 

wetlands by decreasing overland travel distance. Wilson’s phalaropes nest in upland 

vegetation early in the breeding season and wet-meadow vegetation later in the season 

(Colwell and Oring, 1990). 

• Prevent diversion of water from saline lakes and wetlands in western staging areas (Colwell 

and Jehl, 1994). Preserve and/or restore wetlands (Johnson, 1996). 

• Consider shorebird needs when creating impoundments for waterfowl; provide nesting 

islands and beaches with gentle inclines (Colwell and Oring, 1988c). Wilson’s phalaropes in 
35

Alberta, Saskatchewan, and North Dakota nested on islands or in wet-meadow zones around 

lakes and wetlands (Bent, 1962; Hohn, 1967; Kagarise 1979). 

• Do not disturb (for example, drain, mow, burn, or heavily graze) nesting habitat during the 

breeding season, which generally extends from early May to late July (Kantrud and Higgins, 

1992). 

• Use prescribed burning to improve nesting habitat (Eldridge, 1992). 

• Defer livestock grazing (after 15 July) in pastures that contain wetlands important to 

breeding Wilson’s phalaropes (Prescott and others, 1993). Idle grasslands and previously 

grazed areas provide habitat for nesting, but areas with cattle present during the breeding 

season are less suitable (Renken, 1983; Renken and Dinsmore, 1987; Kantrud and Higgins, 

1992). In Alberta, Wilson’s phalaropes were present in deferred-grazed (grazed after 15 July) 

native pasture (Prescott and others, 1993). 

• Do not construct power lines through or within 1 km of known historical high-water marks 

of wetlands or dry basins known to hold water intermittently (Malcolm, 1982). Avoid 

constructing power lines through flight lines or heavily used waterbird migration routes. 

A conservation plan for Wilson’s phalarope was prepared in 2010 (Lesterhius and Clay, 2010). 

Many of the conservation strategies they suggested were broad and not specific to local areas. A few of 

their general recommendations included ensuring quality of wetland habitats on breeding grounds, and 

maximizing recruitment into the population by minimizing breeding failure (as a result of nest predation 

and abandonment to disturbance). Lesterhius and Clay (2010) pointed out the lack of information about 

the species during its migration outside of North America and on the South American nonbreeding 

(wintering) grounds. Other important areas of research they noted were the need to (a) quantify the 

exposure to and the likely impacts of agrochemicals, heavy metals, and other contaminants on 

phalaropes; (b) determine the extent to which agricultural land uses alter the ecology of wetland 

breeding and migration habitats; (c) determine methods for reducing/eliminating nest predation; (d) 

investigate the impact of the introduction of exotic species (for example, exotic fish species, cheatgrass); 

and (e) develop and refine models to explore the likely effects of climate change on breeding and 

nonbreeding habitats (Lesterhius and Clay, 2010). 

Research Priorities

Wilson’s phalaropes have been relatively well-studied on their breeding grounds and at several 

staging sites in North America (Lesterhaus and Clay, 2010). Therefore, research priorities found in the 

literature focused on the following: (a) how to better understand the use of saline lakes in the altiplano 

of Argentina, Bolivia, Chile, and Peru; (b) how to quantify the importance of wintering areas and 

habitats away from the altiplano saline lakes; and, (c) identification of the habitats used and flyways 

followed during the northbound migration (Lesterhaus and Clay, 2010). However, the U.S. Shorebird 

Conservation Plan research priorities listed for avocets in the previous section could also apply to 

breeding Wilson’s phalaropes in North America (Oring and others, 2000). 

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42

III. Sora (Porzana carolina) 

Introduction

Sora (Porzana carolina) is the most 

abundant and widely distributed North 

American rail (Melvin and Gibbs, 1996). 

Soras breed and winter primarily in freshwater 

marshes dominated by emergent vegetation, 

but they also occur in brackish coastal 

marshes during migration. Soras are more 

often heard rather than seen and have one of 

the most distinctive calls of any marsh bird, a 

loud descending whinny call. Soras feed 

primarily on seeds of wetland plants and 

invertebrates. Wild rice is a favorite food in 

the late summer and fall. Although appearing 

to be weak and reluctant fliers, soras migrate hundreds of kilometers each spring and fall between 

breeding and wintering wetlands. Many wetlands important to soras rank among the most threatened in 

the United States, including the coastal marshes of California, Florida, Louisiana, New Jersey, and 

Texas. The sora is legally hunted in 31 states and two Canadian provinces. Bag limits are generally 

liberal, but little information is available on harvest, population trends, or effect of harvest on 

populations. Key studies of the sora’s breeding biology, ecology, and behavior have been conducted in 

the midwestern United States by Walkinshaw (1940), Pospichal and Marshall (1954), Tanner and 

Hendrickson (1956), and Kaufmann (1983; 1987; 1989). Some researchers report that the sora 

population plunged in central North America during the past 30 years, largely the result of natural loss 

of wetlands from a long drought from which both wetlands and rail populations may yet recover 

(Conway and others, 1994). Others report that sora population trends are uncertain (Melvin and Gibbs, 

1994). 

A simple Google search on the Internet for “Porzana carolina” revealed 108,000 results and a 

Google Scholar search provided 2,070 results (accessed 11 July 2011). “Sora” revealed 111,000 results 

through Google and 1,610 results through Google Scholar (accessed 11 July 2011). 

Natural History

Morphology

The sora is a medium-sized rail (20 to 25 cm total length) with grayish-brown plumage, greenish 

legs, and long toes (Melvin and Gibbs, 1996). Adult mass averages 71.9 to 88.0 g and the mean mass of 

males is greater than that of females. Soras are the only North American rail with a combination of a 

stubby, yellow bill and black mask on the face and throat (Melvin and Gibbs, 1996). Adults of both 

sexes have similar plumage, but females have darker bills. 

Distribution

Soras breed in North America in suitable wetland habitat. The northern limits of their breeding 

range include north-central British Columbia (Campbell and others, 1990), northern Saskatchewan 

(Smith, 1996), northern Manitoba, northern Ontario, south-central Quebec, extreme southern 

Sora, courtesy of Dave Menke, U.S. Fish and Wildlife Service. 
43

Newfoundland (Godfrey, 1986), New Brunswick, Prince Edward Island, and Nova Scotia (Erskine, 

1992). Soras also breed in southeastern Alaska (Kessel and Gibson, 1978; Armstrong, 1995), extreme 

northwestern British Columbia (Campbell and others, 1990), and southwestern Northwest Territories 

(Godfrey, 1986). They may be more likely to breed near the northern edge of their range in Alaska and 

Canada during years when drought conditions exist in more southern breeding areas (Trapp and others, 

1981). The southern limit of the breeding range for soras includes southern Pennsylvania (Brauning, 

1992), southcentral Ohio (Peterjohn, 1989), southcentral Indiana, central Illinois, central Iowa, central 

Nebraska, eastern Colorado (Andrews and Righter, 1992), eastern and southern New Mexico (Hubbard, 

1978), central and northwest Arizona, and southern Nevada. Soras also breed in central Kansas 

(Parmelee and others, 1970; Tacha, 1975; Zimmerman, 1984; Thompson and Ely, 1989), northeastern 

West Virginia (Buckelew and Hall, 1994), and northwestern Virginia (Melvin and Gibbs, 1996). The 

western limit of the breeding range includes western California east of the Sierra from Mono County 

south to Inyo County, northeastern California (Small, 1994), Oregon and Washington mainly east of the 

Cascades (Gilligan and others, 1994), and central British Columbia (Campbell and others, 1990), 

northwestern Washington, around San Francisco Bay, and in the Central Valley of California.

Soras winter from the Atlantic Coastal Plain, throughout Florida, west along the Gulf Coast to 

coastal and southwestern Texas, southern New Mexico, southern and western Arizona, extreme 

southeastern Nevada and extreme southwestern Utah, and south throughout Mexico and Central 

America, the Caribbean, to northern South America (Melvin and Gibbs, 1996). Soras also winter along 

the Pacific Coast of North America. Most soras winter in areas that have a minimum January

temperature above -1o

 C (Root, 1988). Wintering populations extend into colder areas along the 

Colorado River in Arizona and along the Rio Grande in Colorado, where discharges of hot water from 

power plants may create open-water conditions. The densest wintering populations in the United States

are in southern Florida, the bayous of southern Louisiana, the coastal plain of Texas, and the lower 

Colorado River valley in southwestern Arizona (Root, 1988). 

Soras stay in Colorado from early April through early October (Griese and others, 1980). A few 

soras remain in Colorado in winter, but most winter from the southern edge of the United States to 

South America. The Breeding Bird Atlas Colorado (1998) reported soras breeding throughout the state 

where suitable wetlands were available. 

Migration

Little is known about migration behavior and patterns for the sora. Because this species is 

reclusive and difficult to observe, migrants are not easily detected (Melvin and Gibbs, 1996). They are a 

regular spring and fall migrant between their wintering and breeding areas. They are presumed to be 

nocturnal migrants (Stoddard and Norris, 1967; Avery and others, 1976; Robbins, 1991). It is unknown 

whether some populations are resident where breeding and winter ranges overlap. Vocalizations are 

useful in locating migrants during the springtime, but in other seasons, soras are much less vocal so they 

are more difficult detect (Melvin and Gibbs, 1996). Timing of migration, especially in the fall, may be 

determined in part by timing of frosts (Bent, 1926; Walkinshaw, 1940; Tanner and Hendrickson, 1956). 

In some areas of their range, soras appear to move from smaller breeding wetlands to congregate in 

larger wetlands with abundant food in August and early September (Pospichal and Marshall, 1954). 

Earliest spring migrants arrive in Colorado in early April (Griese and others, 1980). 

Haramis and Kearns (2007) conducted banding and telemetry studies of fall migrant soras in the 

historical rail hunting and exceptional stopover habitat of the Wild Rice marshes of the tidal Patuxent 

River. Their radio-telemetry results depicted soras as long-distance migrants with high stopover survival 

and a critical dependence on tidal freshwater marshes for migratory fattening. Their stopover period was 
44

frequently >40 days and the mean mass gain of +0.6 g per day suggested that soras accumulate large fat 

reserves for long-distance flight. Further radio-tracking confirmed that soras are strong flyers with a 

demonstrated overnight (10 hour) flight range of 700-900+ km. They could conceivably fly nonstop 

from the Patuxent River to Florida, the Bahamas, or even the Caribbean (Haramis and Kearns, 2007). 

Habitat

The Breeding Bird Atlas Colorado (1998) reported soras breeding in a variety of wetlands in 

Colorado, including cattail marshes, grass or sedge marshes, wet meadows, and irrigated hayfields 

especially in mountain parks (Pantle, 1998). Soras prefer cattails with shallow water for breeding and 

escape, and 80 percent of Atlas reports came from marshes, with cattails usually the dominant 

vegetation. Atlasers reported much less use of other habitats, such as hay meadows, open water, willow 

cars, lowland riparian thickets, and salt meadows. 

Across their breeding range, soras primarily use wetlands with shallow and intermediate water 

depths, dominated by robust or fine-leaved emergent vegetation, especially cattails (Typha spp.), sedges 

(Carex spp., Cyperus spp.), burreeds (Sparganium spp.), and bulrushes (Scirpus spp.) (Walkinshaw, 

1940, Pospichal and Marshall, 1954; Tanner and Hendrickson, 1956; Griese and others, 1980; Johnson 

and Dinsmore, 1985; 1986a; Gibbs and Melvin, 1990; Gibbs and others, 1991; Crowley, 1994). Soras 

may use wetland edges and upland fields, including row crops, adjacent to wetlands during broodrearing or postbreeding dispersal (Porspichal and Marshall, 1954; Johnson and Dinsmore, 1985). They 

may also occasionally nest in brackish or salt marshes along the Atlantic Coast, in habitat dominated by 

smooth cordgrass (Spartina alterniflora) or common reed (Phragmites communis) (Greenlaw and 

Miller, 1982). 

Soras breeding in prairie potholes in North Dakota and Iowa occur almost exclusively in 

wetlands classified as seasonal or semipermanent (Kantrud and Stewart, 1984; Johnson and Dinsmore, 

1986a). In northwestern Iowa, breeders seem to use emergent vegetation types in proportion to 

availability (Johnson and Dinsmore, 1985; 1986a). The highest breeding densities are in relatively 

shallow, shoreward portions of wetlands where water level instability produces diverse mosaics of fine 

and robust emergent vegetation. This habitat preference may be related to an increased prevalence of 

wetland plant seeds, especially sedges, which are important foods during the breeding season. High 

densities of floating and submerged residual vegetation may increase habitat quality by providing a 

substrate for invertebrates near the water surface and accessible to soras, which have short bills (Melvin 

and Gibbs, 1996). 

In Iowa, mean water depth in 71 sora breeding territories was 38.4 cm + 16.1 standard deviation 

(SD)(range 0-92) and was not significantly different from that for the Virginia rail (Rallus limicola) 

(Johnson and Dinsmore, 1986a). Mean vegetation height was 128.0 cm + 42.5 SD, and mean number of 

stems/m2 in territories was 121.9 cm + 80.9 SD. Soras did not use areas where heavy snow, ice, and 

high water had flattened emergent vegetation until early May, when new sedge growth was 20-30 cm in 

height and 80-100 stems/m2 provided some cover. 

In central Minnesota, sora nesting densities increase with increasing abundance of cattails 

(Pospichal and Marshall, 1954). In eastern Alberta, where cattails are rare, Carex sedges form 

predominant nesting habitat, and breeding densities are no greater where cattails do occur (Lowther, 

1977). In British Columbia, soras nest in wetter habitats and more often in cattails than Virginia rails 

(Campbell and others, 1990). Soras breeding in Massachusetts are present in wetlands with larger areas 

of cattails and greater interspersion of vegetation and water compared to wetlands where they are absent 

or where Virginia rails are present (Crowley, 1994). However, Johnson and Dinsmore (1986a) found no 

clear differences in breeding-habitat use between these two species in Iowa. Wetlands used by breeding 
45

soras in Maine had more fine-leaved and robust emergent and aquatic-bed vegetation (Cowardin and 

others, 1979) and fewer ericaceous shrubs than did wetlands where soras were not present (Gibbs and 

Melvin, 1990). Nests of soras in western New York were located in >70 percent emergent vegetation 

with a mean water depth of 24-56 cm and an average vegetation height that ranged from 69-133 cm (Lor 

and Malecki, 2006). 

Breeding soras appear to be area-independent in their selection of wetlands (Melvin and Gibbs, 

1996). Soras in Wisconsin may breed in wetlands with only a few hectares of shallow water and 

emergent vegetation (Robbins, 1991), and as small as <0.5 ha in Massachusetts (Crowley, 1994). In 

Iowa, soras occurred in 17 percent of wetlands <1 ha (Brown and Dinsmore, 1986). In one Maine study, 

soras occurred in 8 percent of wetlands <1 ha (Gibbs and others, 1991), but in another study, soras were 

absent from wetlands <5 ha but occurred in 39 percent of wetlands >5 ha (Gibbs and Melvin, 1990). In 

Massachusetts, soras occurred on 8 percent of wetlands 0.5-2.0 ha, compared to 22-42 percent of 

wetlands >2 ha (Crowley, 1994). Soras, however, may be more area-sensitive than Virginia rails. They 

occurred less frequently than Virginia rails on wetlands <5 ha in Iowa and Massachusetts (Brown and 

Dinsmore, 1986; Crowley, 1994) and in 1 of 2 Maine studies (Gibbs and Melvin, 1990; Gibbs and 

others, 1991). 

During migration, soras stop and rest in freshwater emergent wetlands and frequently in brackish 

or salt marshes. Stands of wild (Zizannia spp.) and cultivated (Oryza sativa) rice and flooded annual 

grasses or forbs provide important feeding habitats for migrants (Walkinshaw, 1940; Meanley, 1960; 

Odom, 1977; Fannucchi and others, 1986; James and Neal, 1986). Soras seem to prefer wetland sites 

that offer both tall, dense cover and shorter seed-producing plants (Meanley, 1965; Andrews, 1973; 

Griese and others, 1980; Rundle and Fredrickson, 1981; Sayre and Rundle, 1984). Migrants sometimes 

appear away from wetlands, including upland fields, pastures, gardens, and lawns, and, probably 

unintentionally, on roads, airport runways, and sandy beaches (Peterjohn, 1989; Campbell and others, 

1990; Amos, 1991). 

Little is known about habitat use by soras on their winter range, but habitats used include 

freshwater, brackish, and salt marshes (Eddleman and others, 1988; Small, 1994). Soras are probably 

the most abundant in marshes with good interspersion of shallow water and emergent vegetation 

(Gochfeld, 1972; Conway, 1990), but are also found in winter in vegetated canals and ditches, rice 

fields, impoundments, mangroves, wet pastures, overgrown and cultivated fields, and emergent 

vegetation along edges of small ponds and rivers (Gochfeld, 1972; Raffaele, 1989; Stiles and Skutch, 

1989; Amos, 1991; Rosenberg and others, 1991; Howell and Webb, 1995). At the northern fringes of 

their winter range, soras may use wetlands kept free of ice by sewer outlets, freshwater springs, or hotwater discharges from power plants (Root, 1988; Veit and Petersen, 1993). 

Foraging Habits

The main foods eaten by soras are the seeds of wetland plants and aquatic invertebrates (Melvin 

and Gibbs, 1996). Proportion of invertebrates in their diet increases in the spring (Martin and others, 

1951; Rundle and Sayre, 1983). Common plant foods include seeds of wild or cultivated rice, 

smartweeds, sedges, bulrushes, and grasses (Martin and others, 1951; Pospichal and Marshall, 1954; 

Rundle and Sayre, 1983). Principal animal foods include adults, larvae, and pupae of a variety of aquatic 

insects (Orders Coleoptera, Diptera, Hemiptera, Odonata) and snails (Gastropoda) (Pospichal and 

Marshall, 1954; Gochfeld, 1972). 

The microhabitat used for foraging consists of stands of robust emergent vegetation interspersed 

with shorter, seed-producing vegetation, and vegetative debris that provides good substrate for 

invertebrates near the water surface (Meanley, 1965; Rundle and Sayre, 1983; Sayre and Rundle, 1984; 
46

Johnson and Dinsmore, 1986a; Melvin and Gibbs, 1994). During brood-rearing and premigration 

periods, soras may be attracted to shallower portions of wetlands where seed-producing plants such as 

sedges, bulrushes, smartweeds, beggarticks (Bidens spp.), and grasses provide food (Johnson and 

Dinsmore, 1986a; Manci and Rusch, 1988). In the late summer, soras may leave wetlands for short 

periods to feed in upland fields and row crops (Horak, 1970; Johnson and Dinsmore, 1986a). Migrating 

soras feed in wetlands with tall emergent vegetation, where wild or cultivated rice or flooded annual 

grasses and forbs are preferred feeding habitats (Walkinshaw, 1940; Meanley, 1960; Odom, 1977; 

Fannucchi and others, 1986). Migrants in southeastern Missouri preferred water depths of 5-15 cm, but 

often used deeper water, up to 46 cm (Sayre and Rundle, 1984). Because of their foraging behavior and 

omnivorous diet, soras may be able to forage in a wider range of water depths than Virginia rails, which 

seem to occur more frequently at sites with shallow water (<15 cm deep) or saturated soils (Sayre and 

Rundle, 1984). 

Soras search for food by raking floating vegetation with their feet or pulling aside vegetation 

with their bill and visually searching for food (Kaufmann, 1989). They use their short, chickenlike bill 

to peck from substrates or water surfaces. Soras generally feed from a standing position. In autumn, 

soras strip seed heads of maturing smartweeds and annual grasses (Meanley, 1965; Rundle and Sayre, 

1983). They occasionally feed by swimming in open water like a coot (Pospichal and Marshall, 1954). 

Parents feed invertebrates to the chicks for 2-3 weeks (Kaufmann, 1989), although Pospichal and 

Marshall (1954) believed that the chicks were able to forage independently when only a week old. 

Chicks are precocial, follow adults, begging for food, and are fed by adults. They gradually develop the 

ability to eat food items placed in front of them, then capture their own food, and increase their 

consumption of plant material (Kaufmann, 1989). 

Breeding

In Colorado, soras lay their eggs from late May through early July. They have small territories. 

Six pairs of soras (and 10-18 pairs of Virginia rails) nested in a 10-ha marsh near Fort Collins, with only 

10 m separating the closest sora nests (Glahn, 1974). Soras lay a substantial 8-12 eggs; the black downy 

young leave the nest almost immediately upon hatching – as soon as their down dries (Pantle, 1998). 

Soras carefully conceal their nests, and Colorado Breeding Bird Atlasers based most confirmations of 

breeding on observations of fledglings. 

Demography & Population Status

There is no information on annual survival or mean life span for soras (Melvin and Gibbs, 

1996). Survival probability during the breeding season (August through April) of 23 radiomarked soras 

in southwestern Arizona was only 0.308 + 0.256 SE for all age and sex classes combined (Conway and 

others, 1994). This survival probability may have been lowered by the effects of the radio transmitters 

or underestimated because of emigration from the study area. Causes of mortality include predation, 

exposure, shooting and trapping, and collision with towers (Melvin and Gibbs, 1996). Pospichal and 

Marshall (1954) reported predation to be the primary cause of egg loss, although actual cases of 

predation were never observed. Confirmed or strongly suspected egg predators include marsh wrens 

(Cistothorus palustris; Allen, 1934), common grackle (Quiscalus quiscula), common crow (Corvus

brachyrhinchos), striped skunk (Mephitus mephitis), coyote (Canis latrans; Lowther, 1977), and 

raccoon (Procyon lotor; Tanner and Hendrickson, 1956). Porspichal and Marshall (1954) reported 2 

clutches of soras crushed by muskrats (Ondatra zibethicus) that used nests as feeding platforms. 

Lowther (1977) reported 2 nests trampled by cattle after water under the nests had dried. Confirmed 

predators on adult or juvenile soras include coyote (Lowther, 1977), bobcat (Felis rufus; James and 
47

Neal, 1986), feral house cats (F. silvestrus; Porspichal and Marshall, 1954), northern harriers (Circus

cyaneus), Cooper’s hawk (Accipiter cooperii; Craighead and Craighead, 1969), red-shouldered hawk 

(Buteo lineatus; Bent, 1937), peregrine falcon (Falco peregrinus; Barber and Barber, 1983), barn owl 

(Tyto alba; Bent, 1938; French and Wharton, 1975), great horned owl (Bubo virginianus; Errington and 

others, 1940), and short-eared owl (Asio flammeus; Bent, 1938). 

Although soras are considered the most abundant rail in North America (Melvin and Gibbs, 

1994), population size has not been estimated. Local density estimates of breeding soras include 0.1 

birds/ha in 1,321 wetlands in central North Dakota (Kantrud and Stewart, 1984), 1.3 birds/ha at Horicon 

Marsh in southern Wisconsin (Manci and Rusch, 1988), 1.6 birds/ha in 28 wetlands in northwestern 

Iowa (Tanner and Hendrickson, 1956), and 0.6 pairs/ha at Fossil Creek Reservoir in northern Colorado 

(Glahn, 1974). Breeding Bird Survey (BBS) estimates that populations of soras have declined 3.3 

percent + 1.7 SD annually from 1966-1991 in North America (p < 0.05, n = 528 BBS routes; Conway 

and others, 1994). Populations were stable 1982-1991 (p > 0.10, n = 149 BBS routes) in Canada, but 

declined 8.5 percent + 2.1 SD annually in the United States during the same period (p < 0.01, n = 241 

BBS routes) (Conway and others, 1994). Declines were most severe in central North America, where 

wetland loss has been the greatest (Tiner, 1984; Dahl, 1990; Dahl and Johnson, 1991). The Breeding 

Bird Survey regional trend analysis for soras along 30 Colorado routes from 1966 to 1999 showed a 

positive trend estimate of 0.5 (-2.8, 4.8; 95 percent CI) (Sauer and others, 2011). From 1999 to 2009, the 

trend was also positive and estimated to be 0.3 (-6.6, 8.4; 95 percent CI). However, the regional 

credibility measure for this species was in the “red” category, meaning that data have an important 

deficiency because there were overall low abundances of soras per route and/or the results were too 

imprecise to detect long-term trends. 

Threats

Shooting and Trapping

Soras are legally hunted in 31 states and in Manitoba and Ontario, and bag limits are liberal in 

most states (25 birds/day in 27 states; Melvin and Gibbs, 1994). From 1964 through 1976, an estimated 

13,400-47,200 soras were harvested annually by waterfowl hunters, and an equal number were 

estimated to have been shot by nonwaterfowl hunters (Martin, 1979). Only 14 states report recent 

estimates of numbers of sora hunters or harvest, and 9 of these lumped harvest estimates of soras with 

other rails (Melvin and Gibbs, 1994). Five states estimated harvests of >2,000 soras: Florida, Iowa, 

Louisiana, New Jersey, and Virginia. The effects of annual harvest on sora populations are unknown. 

Although low hunter interest and the difficulty of hunting rails probably keeps annual kill within 

sustainable levels, effects of incidental take by waterfowl hunters and other bird hunters on local or 

regional populations are unknown (Eddleman and others, 1988; Melvin and Gibbs, 1994). Much larger 

numbers were harvested early in the 20th century. Bent (1926) reported that hunters commonly shot 100-

150 rails/day along the Connecticut River in Connecticut, and estimated 1,000 rails killed at 

Longmeadow, Massachusetts, in 1908. Sora hunting is an old tradition in coastal marshes of 

Connecticut, New Jersey, Maryland, and Virginia, but numbers of both birds and hunters have declined 

substantially (Ripley, 1977). Fur trapping has also been known to cause accidental mortality of soras 

and other rails (Meanley, 1969; Linscomb, 1976; Eddleman and others, 1988). 

Other Potential Threats

There are no data for the effects of pesticides and other contaminants on soras (Melvin and 

Gibbs, 1996), but they may be a threat to soras in southern commercial rice fields and in wetlands in 
48

California and Arizona (Eddleman and others, 1988). Soras are potentially susceptible to mortality from 

ingested lead shot, which was present in the gizzards of 7.4-12.3 percent of soras collected in Maryland 

and 1.8 percent of birds from Missouri (Artmann and Martin, 1975; Stendell and others, 1980). 

Migrating soras are commonly killed at tall, lighted towers (Stoddard and Norris, 1967; Avery and 

others, 1976; Robbins, 1991) and probably in collisions with overhead wires (Sibley, 1993). At a TV 

tower in northern Florida, Stoddard and Norris (1967) found soras killed in all months except June, July, 

and August, and in greater numbers than other rails. 

In southcentral Colorado, decreasing water levels reduced suitable habitat by 95 percent in July 

and early August 1975 and 1976. Irrigation practices contributed to midsummer drying of wetlands and 

caused premature dispersal of rails to concentrate in wetlands with more water (Griese and others, 

1980). Porspichal and Marshall (1954) attributed an incident of nest abandonment to human 

disturbance. Harvesting of wild rice may disturb or reduce food available to local or migrant soras 

(Fannucchi and others, 1986). Porspichal and Marshall (1954) also reported no significant difference in 

mortality rates of marked and unmarked juvenile soras. Conway and others (1994) speculated that 

depredation of 3 of 26 radiomarked soras within 4 days of capture was the result of radiomarking, and 

that the low estimated survival probability during nonbreeding season may have reflected increased 

mortality of radiomarked individuals. 

Sora behavior changes in response to human disturbance. Burger and Gochfeld (1998) studied 

the effects of ecotourists on bird behavior at the Loxahatchee National Wildlife Refuge in Florida. They 

observed birds before people were near, while people were present, and following the departure of 

people. The variation in feeding behavior was largely explained by whether people were present, the 

number of people present, and the amount of noise made by people. Time devoted to feeding and 

number of strikes or pecks decreased while people were present. The percentage of time spent foraging 

and the number of strikes decreased as the noise by people increased (Burger and Gochfeld, 1998). 

Management

Response to Management

Mowing can be used to create more interspersion of habitats, which has been shown to be 

beneficial for marsh birds, including soras (Rehm and Baldassare, 2007). There is some general 

agreement that marshes that undergo cyclic vegetation changes (resulting from varying water levels) 

show maximum avian use and production during the period when emergent macrophytes and open 

water are present in equal proportions in an interspersed pattern (Weller and Spatcher, 1965; Weller and 

Fredrickson, 1974). One way to create interspersion of habitats is by mowing, which increases habitat 

edges. An increase in edge density was shown to be the best predictor of abundance of soras, and other 

marsh birds, in two marsh locations in NewYork (Rehm and Baldassare, 2007). 

Coastal wetlands in the Gulf Coast Chenier Plain are commonly managed intensively by fall or 

winter burning and constructing impoundments to improve winter habitat for waterbirds, reduce wetland 

loss, and create emergent wetlands (Gabrey and others, 2001). Gabrey and others (2001) conducted 

experimental burns in 4 types of impounded and unimpounded marshes on the Rockefeller State 

Wildlife Area in southwestern Louisiana. They found that vegetation characteristics in burned marshes 

did not differ from those of unburned marshes by the first summer post-burn, and that winter burning 

did not affect bird species richness or species composition. Bird surveys, including those for soras, 

showed no difference between burned and unburned marshes during the first or third summers postburn, but were on average 2 times greater in burned than in unburned marshes during the second 

summer post-burn (Gabrey and others, 2001). Controlled burning of marshes has also been used 
49

successfully to prevent woody invasion (Zimmerman, 1984). In a radio-telemetry study of soras at the 

USFWS Texas Mid-coast National Wildlife Refuge Complex, there was no direct mortality from 

prescribed burns (Grace and others, 2005). For the cases in which birds were tracked for the longest 

period, their ability to survive fire was documented. One sora was found alive under a clump of burned 

gulf cordgrass. The authors suggested that the rails that were unable to escape fire managed to survive 

by burying themselves underneath dead, damp vegetation. They also believed that the presence of 

standing water could aid in the survival of rails during prescribed burns. They suggested that managers 

conducting winter prescribed burns in this particular area should burn early in the winter when water is 

still present (Grace and others, 2005). 

Conway and others (2010) evaluated whether prescribed fire could be used as a surrogate 

disturbance event to help restore historical flooding conditions along the lower Colorado River basin for 

the benefit of Yuma Clapper rails and four sympatric marsh-dependent birds, including soras. They 

conducted call-broadcast surveys for marsh birds within burned and unburned (control) plots both preand post-burn and found that fire increased the number of Yuma Clapper rails and Virginia rails, but did 

not affect the numbers of black rails, soras, and least bitterns. They concluded that prescribed fire could 

be used to set back succession of emergent marshlands and help mimic the natural disturbance regime in 

the lower Colorado River basin (Conway and others, 2010). 

In Iowa, several studies have found that soras do not use restored wetland basins with any high 

degree of frequency (Delphey and Dinsmore, 1993; VanRees-Siewert, 1993; Schreiber, 1994). 

However, Schuster (1998) studied restored and natural wetlands and avifaunal use in the Prairie Pothole 

region of northern Iowa and found that both Virginia rails and soras used both wetland types equally. In 

the Columbia River Basin of eastern Washington, densities of soras were greater on unaltered wetlands 

than on wetlands that had been excavated to create openings for dabbling and diving ducks (Creighton 

and others, 1997). 

Linz and others (1997) studied the effects of herbicide-induced changes in wetland vegetation 

(largely cattails, Typha spp.) on densities of soras in northeastern North Dakota. In 1990 and 1991, 17 

cattail-dominated wetlands were randomly assigned to 0 percent (reference wetlands), 50 percent, 70 

percent, or 90 percent spray coverages with glyphosate-based herbicide. They found that one year posttreatment, reference wetlands harbored more soras than did the treated wetlands but sora numbers were 

similar among treatments two years post-treatment. They concluded that sora populations using 

semipermanent wetlands may be negatively influenced by fragmenting stands of dense cattails. 

Timmermans and others (2008) examined the associations between breeding marsh bird 

abundances and Great Lakes hydrology. From 1995-2002, Lakes Erie and Huron-Michigan experienced 

water-level declines. Abundance of soras declined within the marshes of these lakes during the same 

period. They concluded that long-term changes and annual water-level fluctuations are important abiotic 

factors affecting the abundance of some marsh-dependent birds in Great Lakes coastal marshes. 

Recommendations

Melvin and Gibbs (1994) summarized the management needs of soras. Most important is 

preservation of emergent wetlands that provide breeding, migration, and wintering habitats. Many 

wetland types most important to soras remain among the most threatened in the United States, including 

coastal marshes of California, Florida, Louisiana, New Jersey, and Texas, palustrine emergent wetlands 

in south Fordia and Prairie Pothole Region, and western riparian wetlands. Soras would benefit from 

policies and management that eliminate or minimize the effects of wetland draining and filling; siltation, 

eutrophication, and other forms of pollution; and invasion by exotic plants. 
50

Wetland habitat management to benefit soras should encourage diverse stands of both fineleaved and robust emergent vegetation, including sedges, bulrushes, and especially cattails, as well as 

moist-soil annuals around edges of wetlands (Rundle and Sayre, 1983; Johnson and Dinsmore, 1986). 

Periodic gradual drawdowns that encourage horizontal zonation of wetland vegetation may maximize 

habitat quality. Management should maximize interspersion of emergent vegetation and open water 

areas (Weller and Spatcher, 1965), because soras seem most abundant near edges of cover types or 

vegetation and open water (Walkinshaw, 1940; Porspichal and Marshall, 1954; Johnson and Dinsmore, 

1986; Crowley, 1994; Rehm and Baldassare, 2007). Wetland impoundments constructed with irregular 

or sloping bottoms will provide greater diversity of water levels and topography and will increase 

vegetation-water edge (Eddleman and others, 1988). Impoundments drawn down during the summer can 

be flooded in the late summer to provide habitat for post-breeding and migrant soras (Rundle and 

Fredrickson, 1981). When spring rail use is desired, impoundments with annual grasses and smartweeds 

should be de-watered over winter to protect vegetation from ice damage and waterfowl and to 

encourage growth of perennial vegetation. 

Research Priorities

Research needs and priorities for soras were summarized by Melvin and Gibbs (1994): 

• Develop reliable techniques for monitoring population trends. Indices of population trends using 

playback recording surveys will require identifying periods of peak response to tape-recorded 

calls, and obtaining better distributional information on breeding soras to allow stratification of 

surveys within states and provinces. 

• Develop a reliable harvest survey in the United States and Canada. 

• Determine the effects of habitat quality, predation, weather, and water-level fluctuations on 

reproductive success and breeding densities of soras. 

• Describe habitat use and distribution patterns of wintering soras. Determine management actions 

necessary to maximize carrying-capacity of winter habitats. 

• Describe chronology, habitat use, and behavior of soras during postbreeding dispersal and 

migration. Priority should be given to unstudied populations outside the Midwestern United 

States. 

• Characterize site fidelity, annual survivorship, life span, and age of first breeding. 

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Peterjohn, B.G., 1989, The birds of Ohio: Indiana University Press, Bloomington, Indiana. 

Pospichal, L.B., and Marshall, W., 1954, A field study of sora rail and Virginia rail in central 

Minnesota: Flicker, v. 26, p. 2-32. 

Raffaele, H.A., 1989, A guide to the birds of Puerto Rico and the Virgin Islands: Princeton University 

Press, Princeton, New Jersey. 

Rehm, E.M., and Baldassarre, G.A., 2007, The influence of interspersion on marsh bird abundance in 

New York: Wilson Journal of Ornithology, v. 119, p. 648-654. 

Ripley, S.D., 1977, Rails of the world: David R. Godine, Publishers, Boston, Massachusetts. 

Root, T., 1988, Atlas of wintering North American birds: an analysis of Christmas count data: 

University of Chicago Press, Chicago, Illinois. 

Rosenberg, K.V., Ohmart, R.D., Hunter, W.C., and Anderson, B.W., 1991, Birds of the lower Colorado 

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rails and shorebirds: Wildlife Society Bulletin, v. 9, p. 80-87. 

Rundle, W.D., and Sayre, M.W., 1983, Feeding ecology of migrant soras in southeastern Missouri: 

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55

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Journal of Field Ornithology, v. 55, p. 38-47. 
56

IV. White-faced Ibis (Plegadis chihi) 

Introduction

The white-faced ibis (Plegadis chihi) is a longlegged wader with a long, decurved bill and metallic 

bronze plumage. During the breeding season, adults 

have distinctive white feathers along the edge of their 

bare facial skin. The species is common, nesting in 

marshes in the western United States, and wintering in 

large flocks in Mexico, western Louisiana, and eastern 

Texas (Ryder and Manry, 1994). The species inhabits 

primarily freshwater wetlands, especially cattail (Typha

spp.) and bulrush (Scirpus spp.) marshes, although it 

feeds in flooded hay meadows, agricultural fields, and 

estuarine wetlands. No subspecies of this ibis are 

recognized, although the more cosmopolitan glossy ibis 

(Plegadis falcinellus) is considered by some to be 

conspecific, and the two species have produced hybrids 

in captivity. In Louisiana, Alabama, and possibly 

eastern Texas, 

however, the two forms nest in the same colonies, apparently without interbreeding. 

A simple Google search on the Internet for “Plegadis chihi” revealed 94,700 results and a 

Google Scholar search provided 1,080 results (accessed 11 July 2011). “White-faced ibis” revealed 

324,000 results through Google and 2,170 results through Google Scholar (accessed 11 July 2011). 

Natural History

Distribution

The white-faced ibis breeds in northern California, eastern Oregon, southern Idaho, southern 

Alberta, Montana, eastern North and South Dakota, Washington, and recently in northwestern Iowa 

south to Durango and Jalisco states in Mexico (Ryder and Manry, 1994; Tweit and Flores, 2006). 

White-faced ibises were found attempting to nest in Washington in 2001 (Tweit and Flores, 2006). The 

first record of nesting in Arizona occurred in 2000 at Cibola National Wildlife Refuge. The Mexican 

breeding distribution is poorly known. The largest breeding colonies are usually in Utah, Nevada, 

Oregon, and coastal Texas and Louisiana. Ibises winter primarily in coastal Louisiana and Texas south 

to the Mexican states of Guerrero, Puebla, and Tabasco (band recoveries), occasionally Guatemala 

(Land, 1970) and Costa Rica (Stiles and Skutch, 1989). Ibises also winter at Los Banos and the Salton 

Sea, California, other parts of southern California, and in the lower Colorado River Valley in Arizona 

(Rosenberg and others, 1991). 

Ibises are “boom” or “bust” species as breeding populations vary considerably from year to year 

depending on water levels in favored marshes (Ryder, 1967). In 1949, at least 12 pairs nested at Russell 

Lakes in the San Luis Valley (SLV) (Ryder, 1950); however, in 1965 only 10 pairs bred in the whole 

SLV (Ryder, 1967). During the Colorado Breeding Bird Atlas survey, an estimated 355 pairs nested on 

Alamosa NWR, and more than 470 pairs were documented in the SLV during 1993 (Rilling and 

Falzone-Schrim, 1993). 

White-faced Ibis, courtesy of Eleanor Briccetti.
57

Migration

Ibises are highly mobile, especially in the northern hemisphere. Northernmost populations 

undertake regular north-south migrations; Texas and Louisiana nesters are mainly resident year-round. 

All populations show some postbreeding wandering. They are well adapted to find new nesting areas 

when regular sites become dry in droughts or are drained by humans (Ryder, 1967). Most band 

recoveries of birds banded as nestlings in Utah (205 of 222) have been in Mexico during the 

nonbreeding season: 55 in Michoacan, 45 in Jalisco, 28 in Nayarit, 24 each in Guanajuato and Sinaloa, 

and 29 in 9 other states. Only 1 of 16 band recoveries from Texas bandings has been in Mexico. Most 

recoveries of birds banded in Oregon (5 of 7), Idaho (15 of 17), and Nevada (13 of 16) have occurred in 

Mexico during winter. Two of 4 recoveries of Colorado bandings have been from Mexico. Some east to 

west movements are also suggested by 2 recoveries of Utah birds in California and one recovery of a 

Nevada bird in California (Ryder and Manry, 1994). 

Most birds arrive at Utah and Idaho breeding areas in April and leave in September or October. 

A few stragglers linger in Utah as late as December (Ryder, 1967; Behle and Perry, 1975; Taylor and 

others, 1989). In Oregon, ibises begin migrating from Malheur National Wildlife Refuge (NWR) in 

early August and most have left by late September (Littlefield, 1990). Oregon birds return as early as 9 

April, but average spring arrival date is 3 May. The Colorado River Valley appears to be an important 

migration route between Great Basin nesting areas and Mexican wintering grounds (Ryder, 1967). 

During the Colorado Breeding Bird Atlas surveys (1998) (Ryder, 1998), migrant ibises were 

reported from all 28 latilongs of the state, but occur most often on the eastern plains and western 

plateaus. Most ibises leave Colorado in September, some as late as October. Based on band recoveries, 

most ibises from Colorado winter in Texas and Mexico (Ryder, 1993; 1994). 

Habitat

General habitats where ibises can be found regardless of region and season are shallow marshes, 

ponds, mudflats, and swamps (Ryder and Manry, 1994). Ibises prefer, almost exclusively, areas with 

emergent vegetation of some type. They inhabit primarily freshwater wetlands and marshes, especially 

those containing cattail (Typha spp.) and bulrush (Scirpus spp.) (Ryder and Manry, 1994), but they can 

inhabit estuarine wetlands and coastal areas as well. Ibises usually nest in emergent vegetation or low 

trees and shrubs over shallow water, but can also nest on the ground on small islands. Ibises need some 

degree of isolation by marshes or small islands for both nesting colonies and social night roosts (Trost, 

1989). At two locations in Utah, all nests were in hardstem bulrush (Kaneko, 1972) or alkali bulrush 

(Kotter, 1970). A nest’s height, above the water or ground, may vary depending upon substrate type or 

other conditions. In one Utah colony, nests ranged from 99.0-202.2 cm above water that was 61 cm 

deep (Kaneko, 1972). 

The white-faced ibis feeds in flooded hay meadows, agricultural fields, and wetlands (for 

example, pond and reservoir margins, mudflats, and marshes) with short, emergent vegetation (Ryder 

and Manry, 1994). Dominant plant species in foraging areas generally are sedges (Carex spp.), 

spikerushes (Eleocharis spp.), salt-tolerant grassworts (Salicornia spp.), saltgrass (Distichlus stricta), 

and greasewood (Sarcobatus vermiculatus) (Ryder and Manry, 1994). In Nevada, Colorado, Utah, 

Idaho, California, and Oregon irrigated crops of alfalfa, barley, and hay are important feeding sites for 

ibises. In Nevada, white-faced ibises feed in recently flooded agricultural fields where vegetation is <5 

to 90 cm high (Bray, 1986), and they show a strong preference for alfalfa fields. Strong preferences 

were also found for large (>30 ha), level (<5 percent slope) fields with clay or clay-loam soils and pools 

of standing water (Ryder and Manry, 1994). Laubhan and Gammonley (2000) speculated that 

differences in food abundance, vegetation structure and composition, and behavioral strategies during 
58

the breeding season, all influenced the ibis’ selection of foraging habitats in the SLV of Colorado. 

Along the Louisiana coast, flooded rice fields and livestock pastures over salt marshes are also used as 

feeding sites. In Wyoming, white-faced ibises have been noted in marshes, wet meadows, and along 

vegetated shorelines (Dorn and Dorn, 1999). Shallow marshes with emergent vegetation, lakes with 

vegetated shorelines, and mudflats are preferred habitat, and these areas are often surrounded by 

sagebrush-grassland or saltbush communities (Merrill and others, 1996). Foraging sites usually consist 

of shallow wetlands, agricultural fields, and shorelines of ponds and lakes (Ritter and Cerovski, 1990). 

Suitable summer or breeding habitat for the white-faced ibis includes shallow marshes with 

“islands” of emergent vegetation, spoil banks of lakes created by dredging, and locally flooded shoals 

and mangrove swamps (Ryder and Manry, 1994). Ibises generally nest in wetlands on “islands” of 

emergent vegetation that are either flooded or that extend over the water (Ryder and Manry, 1994). 

However, they can also nest inland in various vegetation types as well. Most inland nesting sites in the 

Colorado Plateau Province consist of sagebrush (Artemisia spp.) and saltbush (Atriplex spp.)(Bailey, 

1978). In the Great Basin, most nesting sites consist of hardstem bulrush (Scirpus acutus), Olney’s 

bulrush (S. olneyi), and alkali bulrush (S. paludosus). In Colorado, Baltic rush, sedge, Carex spp., and 

other wetland grasses and forbs are important (Dark-Smiley and Keinath, 2003). In California, saltcedar 

(Tamarix spp.) and Baltic rush (Juncus balticus) are primarily used (Ivey and Severson, 1984). At the 

Klamath River Basin in Oregon, Taft and others (2000) observed that white-faced ibis rely on early 

successional emergent plants such as hardstem bulrush for nesting. 

Habitat used by white-faced ibises during migration is similar to that used during the breeding 

season, and includes irrigated fields and mudflats along rivers and around lakes and isolated reservoirs 

(Taylor and others, 1989). In Oklahoma, they have been reported using a variety of habitats, from 

wooded streams to mudflats, and grassy fields bordering playas (Baumgartner and Baumgartner, 1992). 

In California, they have been found using sewage aeration ponds (Locatelli and Blankenship, 1973), and 

in Nebraska they have been documented in sandhill ponds and marshes (Ducey, 1988). 

Area requirements for individual ibises are not known for certain and no studies of what wetland 

complex size best supports successful, persistent colonies have been reported (Dark-Smiley and 

Keinath, 2003). Ibises typically defend an area of about one meter around the nest (Belknap, 1957; 

Kotter, 1970), and they may defend landing and preening perches up to three meters away from their 

nests (Belknap, 1957). Kotter (1970) noted that in Utah no specific spatial relationships were required 

for ibis nests. In large colonies, most nests are densely concentrated at the center and become more 

isolated at the periphery (Kotter, 1970; Kaneko, 1972). At a Utah colony in alkali bulrush, distances 

between ibis nests ranged from 0.5 to 10 m and averaged about 2 m in areas with the highest nest 

density (Kotter, 1970). In Carson Lake, Nevada, a 25 ha colony in hardstem bulrush averaged 75-150 

nests/ha (Herron and others, 1987), and in the SLV, Colorado, ibis nests in hardstem bulrush and cattails 

averaged 8 m from the nearest conspecific (n = 43; Schreur, 1987). Frederick and Ogden (1997) stated 

that large white ibis colonies (>1,000 pairs) need 800 km2

 or more of wetlands to flourish, and proposed 

that larger wetlands probably tend to sustain colonies for longer periods of time. Little is known about 

ibis’ home ranges outside the breeding season, but during the nesting period in Nevada, most birds 

foraged 3 to 6 km (but up to 18 km) from their breeding colony (Bray, 1986). In late summer, prior to 

fall migration, breeding adults and recently fledged young have been reported ranging from 40-48 km 

from colonies in Idaho (Trost, 1989). No studies have reported how foraging radius might be affected by 

breeding colony size or density. 

Earnst and others (1998) suggested that white-faced ibis would benefit from a landscape mosaic 

of “well-distributed peripheral wetlands and persistent colony sites.” Laubhan and Gammonley (2000) 

believed that hydrology is the single most important factor influencing ecosystem processes in wetlands,
59

as well as habitat suitability for the white-faced ibis. The landscape mosaic must consist of wetlands that 

have new vegetation growth and that contain suitable water levels (approximately 1 m in the SLV, 

Colorado) from late April through the end of July (Dark-Smiley and Keinath, 2003) to sustain ibises 

through the breeding season. In 2003, Monte Vista NWR experienced its driest year since 1906, and 

although some ibises were nesting late in May, they had deserted the site by the middle of June due to 

several factors, including the lack of new bulrush growth and low water levels (Dark-Smiley and 

Keinath, 2003). 

The availability of feeding areas within 15-25 km of breeding colonies is a very important 

landscape feature for the white-faced ibis (Dark-Smiley and Keinath, 2003). The importance of foraging 

areas was demonstrated in the summer of 2003 at the Monte Vista NWR in Colorado, where nesting 

colonies were deserted in the middle of the breeding season. It is believed that the primary reason for 

abandonment and low reproduction on the Refuge was the minimal or lack of shallow-water, wet 

meadows where ibises forage. 

Frederick and Ogden (1997) determined the longevity of white-faced ibis colonies based on 

nesting data from Frederick and others (1996). Their results showed that large colonies of white ibis 

(>1,000 pairs) persisted for one to 17 years and maintained at least 1,000 pairs during this time. They 

concluded that there may be a relationship between the size of a wetland and the longevity of ibis 

colonies. The colonies of greatest longevity that persisted up to 15 years, were associated with relatively 

large wetlands (100 to 150 km in dimension), and colonies that existed for only 10 years or less were 

associated with smaller wetlands (less than 10 km2

 in dimension) (Frederick and Ogden, 1997). This 

same study also reported that large, new ibis colonies are generally associated with unusually high 

abundances of food and exceptional breeding conditions, while abandonment of ibis colonies is 

associated with degraded breeding conditions and changes in water levels or water management.

Food Habits

The main foods taken by white-faced ibises are aquatic and moist-soil insects and earthworms 

(Ryder and Manry, 1994). Larval insects, leeches, and snails are also eaten (Petersen, 1953; Capen, 

1977; Bray, 1986; Bray and Klebenow, 1988; Alcorn, 1988). In Louisiana, crayfish, small fish, frogs, 

snails, and bivalves are consumed (Belknap, 1957). In Idaho, chironomid larvae (midges) may be an 

important food item (Taylor and others, 1989). Ibises choose shallowly flooded pond margins, 

reservoirs, and marshes to forage. Ibises also forage on recently flooded agricultural fields where 

vegetation is <5 to 90 cm high (Bray, 1986; Bray and Klebenow, 1988). Ibises feed in large flocks (up 

to >1,000). They leave their roosting and nesting marshes in the early morning to feed primarily in 

flood-irrigated agricultural fields, especially alfalfa. They use standing water to wash soil from prey 

items, although esophagi of birds collected in Nevada contained substantial amounts of soil (Bray, 

1986; Bray and Klebenow, 1988). Large flocks of ibises were observed foraging for coprophilic beetles 

at a poultry farm in La Paz, Baja California Sur, Mexico (Blanco and Rodriguez-Estrella, 1998). 

Breeding

Most ibises nesting in Colorado favor tall emergent vegetation such as bulrushes and cattails 

growing as “islands” surrounded by water more than 45 cm deep. As early as 1872, Aiken found ibises 

nesting in the SLV (Sclater, 1912) and in 1993 they still nested in at least 11 colonies. Ibises usually 

arrive in the SLV by mid-April and nest in May, but dates may vary from year to year. Both sexes take 

part in nest-building. Courtship occurs mainly in nesting colonies, with some calling and bickering 

between rivals. Nests consist of bulrush and cattail stems and average about 1 m above water. Colonies 

may consist of fewer than 5 pairs up to 100 pairs, usually intermixed with nesting Snowy egrets, black-
60

crowned night-herons, and in recent years, cattle egrets. In the SLV, incubation and hatching in 1984 

extended from early June to late July, yet from early May to late July in 1985 (Schreur, 1987). Young 

usually fledge by late July to mid-August. The Breeding Bird Atlas for Colorado confirmed that in 

Colorado most ibises continue to nest in the SLV, especially in marshes around Russell Lakes, between 

Head and San Luis Lakes, at Adams Lake, and on Alamosa and Monte Vista NWRs. Post-atlas, in 1996, 

two pairs nested at Arapaho NWR in North Park. 

Nest initiation seems to be synchronous with subgroups of ibis, and often in large colonies there 

are “subcolonies” of 20-30 nests, all of which are at a similar stage (Trost, 1989). Once nest initiation 

has occurred, it is very important that water levels do not drop below approximately 1 m, or else the 

ibises will desert the colony (Dark-Smiley and Keinath, 2003). 

The mating system of the white-faced ibis is poorly known, but they are thought to be 

monogamous. There have never been more than 2 adult attendants observed at any active nest (Belknap, 

1957; Kotter, 1970). The events leading up to pair formation, and the duration of pair bonds are not 

known either (Ryder and Manry, 1994). In Utah, adults generally begin to forage in pairs one or two 

weeks before onset of nesting (Kotter, 1970), and in Louisiana pairs arrive at the colony “already 

mated” (Belknap, 1957). 

As stated previously, breeding can be highly synchronous within a large colony, with courting, 

nest-building, incubation, and fledglings occurring at the same time (Belknap, 1957). Seven pairs of 

white-faced ibis nesting in the SLV all completed hatching within a 4-day period, indicating that nest 

initiation was highly synchronous (Schreuer, 1987). Ibises may re-nest if initial breeding attempts fail 

early in the season, but distinguishing between second nesting efforts and late nesting efforts is very 

difficult (Kotter, 1970; Bryant, 1983). The breeding season in the northern Great Plains seems to be up 

to 1 month later than in the southwestern United States, and this is thought to be the result of cooler 

early-spring temperatures in the northern areas (Kotter, 1970; Steele, 1980). 

Eggs are laid from the 2nd week of April to the last week of June in Utah, and egg laying 

extended over a period of 41 days in a larger colony (Kotter, 1970). Eggs are usually laid in the morning 

(Caspen, 1977) on every other day, and clutches are completed over a period of 5 to 8 days (Kotter, 

1970; Capen, 1977). In Colorado at Monte Vista NWR, ibises usually have 1-3 eggs by the third week 

of May and young leave the nest in mid- to late July (Dark-Smiley and Keinath, 2003). Egg-laying dates 

from June 4 to July 25 in eastern Wyoming and Iowa (Findholdt, 1984; Dinsmore and Dinsmore, 1986) 

indicate that breeding occurs several weeks later in the northern Great Plains than in Utah, Nevada, and 

Colorado. This is probably due to the colder weather in northern areas during the early spring months. 

Hatchlings usually emerge within 24 hours of initial pipping (Kotter, 1970). Sequential eggs in 

the clutch usually hatch one or two days apart, although they can hatch on the same day or up to three 

days later (Kotter, 1970). Young are born semi-altricial (Trost, 1989). Adults brood and shade the young 

continuously from the time they hatch up to two weeks of age (Ryder and Manry, 1994), after which 

young are left unattended for up to three hours at a time until brooding ceases completely during week 

three (Belknap, 1957; Kotter, 1970). Both sexes brood and feed the young (Ryder and Manry, 1994). 

Young are independent at eight weeks of age (Kotter, 1970). There is very little information regarding 

dispersal. Band recoveries suggest that no yearlings and few second-year birds return to natal areas to 

breed in the next breeding season (Capen, 1977). Adults may leave the breeding colony after the young 

have fledged, but wandering seems more prevalent during prebreeding compared to postbreeding 

(Ryder, 1967). 
61

Demography & Populations

The oldest white-faced ibis known in the wild was 14 years and 6 months (Clapp and others, 

1982). In captivity, the oldest bird was at least 14 years (Stott, 1948). Based on 111 band recoveries of 

birds banded as nestlings in Utah, all birds died by 9 years of age (Ryder, 1967). 

In the 1960s and 1970s, nesting populations and numbers of colonies in North America 

decreased precipitously because of pesticide contamination (mainly from organochlorines such as DDT) 

and loss of habitat to drought and drainage (Ryder, 1967; King and others, 1980). In the 1980s and early 

1990s, distribution of the species increased as habitat improved and organochlorines were banned, 

reclaiming historical breeding range and expanding north to Alberta and east to North and South 

Dakota, Iowa, and Alabama. A single nest at Lake Washington, Florida (Brewster, 1986) and at least 2 

nests at Heron Lake, Minnesota, in both 1894 and 1895 (Peabody, 1896) suggest a wider breeding 

distribution in the late 19th century (Ryder and Manry, 1994). 

Population numbers for white-faced ibis are difficult to assess because of a lack of census data 

(Ryder and Manry, 1994). Total breeding populations in the Great Basin (Idaho, Nevada, Utah) 

averaged about 7,500 pairs in the late 1970s and early 1980s (Sharp, 1985). In southern Idaho, the total 

nesting population was estimated at nearly 2,000 pairs in the mid-1980s (Taylor and others, 1989). 

Annual nesting populations ranged from 450 to 8,690 pairs at the Great Salt Lake, Utah, during the 

1970s (Capen, 1977; Steele, 1984), from 1,400 to 5,000 pairs in Lahontan Valley, Nevada, during 1980-

1988 (Henny and Herron, 1989), and from 600 to 2,475 pairs in Harney Basin, Oregon, during 1980-

1987 (Ivey and others, 1988). Along the Texas coast, annual breeding populations fluctuated between 

1,520 and 6,500 pairs during 1973-1990 (Mullins and others, 1982). 

Breeding range and populations have expanded in the last two decades (Ryder and Manry, 

1994). This is owed in part to improved nesting habitat management on Federal and state Refuges, 

increased planting of alfalfa used by feeding ibises (in the western United States), the banning of DDT 

and other pesticides in the 1970s, and improved breeding success at major nesting centers. Since 1970, 

nesting populations have increased dramatically in Idaho (Taylor and others, 1989), Oregon (Ivey and 

others, 1988), and Nevada (Henny and Herron, 1989), with more modest gains in the SLV (Ryder and 

Manry, 1994), California (Ryder, 1967; Ivey and Severson, 1984, Sharp, 1985), and Wyoming (Ritter 

and Cerovski, 1990). The first state nesting records were confirmed in Montana in 1970, in North and 

South Dakota in 1978 (Schmidt, 1980), and in Iowa in 1986 (Dinsmore and Dinsmore, 1986) as 

breeding range expanded into the northern Great Plains. The first state nesting record in New Mexico 

was in 1973 (Hundertmark, 1974). In the late 1970s, the breeding range expanded eastward through 

coastal Louisiana to southwestern Alabama, widening the zone of nesting overlap with the glossy ibis 

(Duncan and Johnson, 1977; Jackson and Cooley, 1978). However, Texas coastal breeding populations 

declined from 6,500 to 2,300 pairs between 1981 and 1990 (Ryder and Manry, 1994). In coastal 

Louisiana, ibis populations (predominately white-faced) appear to have declined from 12,495 to 6,255 

breeding adults between 1976 and 1990 (Portnoy, 1977; Martin and Lester, 1990). 

In the western United States, breeding populations fluctuate markedly at major breeding centers 

from year to year in response to changing habitat conditions. Breeding adults appear to be nomadic, 

relocating when drought or floods render traditional sites temporarily unusable (Ryder, 1967; Thompson 

and others, 1979; Steele, 1984; Ivey and others, 1988; Henny and Herron, 1989; Taylor and others, 

1989). Similarly, birds may shift between Texas and Louisiana coastal breeding sites according to 

habitat conditions (King and others, 1980). Within the major breeding areas (for example, the Great Salt 

Lake, Harney Basin, Lahontan Valley), colony sites change from year to year depending on the local 

water conditions, as some sites become desiccated in dry years or are flooded in wet years (Capen, 

1977; Steele, 1984; Ivey and others, 1988; Henny and Herron, 1989). 
62

Annual or biannual censusing of breeding colonies of white-faced ibis is comprehensive and 

consistent in Nevada (Henny and Herron, 1989), Oregon (Ivey and others, 1988), and Texas (Mullins 

and others, 1982), but is sporadic and incomplete in Idaho (Taylor and others, 1989), Wyoming (Ritter 

and Cerovski, 1990), Louisiana (Martin and Lester, 1990), Utah (Sharp, 1985), California, and other 

states (including Colorado). In view of the species’ nomadism and apparent tendency to relocate 

breeding centers across state lines in some years, monitoring long-term population status would require 

coordinated survey efforts covering the entire breeding range (Booser and Sprunt, 1980; Henny and 

Herron, 1989) that are repeated on a regular basis. Population surveys and status assessment in Mexico 

are also urgently required (Sharp, 1985). 

The Breeding Bird Survey (BBS) regional trend analysis for white-faced ibis along 17 Colorado 

routes from 1966 to 1999 showed a positive trend estimate of 11.6 (-1.6, 28.6; 95 percent CI) (Sauer and 

others, 2011). From 1999 to 2009, the trend was also positive and estimated to be 12.5 (-8.2, 33.8; 95 

percent CI). However, the regional credibility measure for this species was the category “red,” meaning 

that data were deficient because of overall low abundances of ibises per route and/or the results were too 

imprecise to detect long-term trends. 

Threats

Causes of mortality include exposure and predation (Ryder and Manry, 1994). Small nestlings 

left unattended may die of exposure in cold, wet weather, and large numbers of dead chicks may be 

found in colonies after severe thunderstorms (Kotter, 1970; Capen, 1977). Extremely high temperatures 

also can be fatal to nestlings (Tyler, 1933; Kingery, 1988). Abandoned or temporarily unguarded eggs 

and chicks are vulnerable to predation by Franklin’s, California, and laughing gulls (Kotter, 1970; 

Capen, 1977; Alford, 1978; King, 1978), grackles (Belknap, 1957), and black-billed magpies (Kotter, 

1970), although the magnitude of these losses is difficult to measure because disturbance by 

investigators is often a contributing factor. Kotter (1970) reported at least 28 percent egg destruction by 

Franklin’s gulls breeding in the same colony, whereas Capen (1977) and Steele (1980) concluded that 

Franklin’s gulls posed little or no threat to nesting ibises. Mink, weasels, striped and spotted skunks, 

raccoons, and coyotes sometimes raid colonies, especially when water levels are low, consuming eggs 

and chicks and sometimes killing adults (Kotter, 1970; Kingery, 1976; 1980; 1988; Capen, 1977; Ryder 

and others, 1979; Herron and others, 1987). Kotter (1970) reported “surplus killing” at one colony, 

possibly by mustelids, which ate very few of the eggs and chicks that were destroyed. Capen (1977) 

reported a similar episode in which California gulls were suspected: eggs and chicks in nearly all 600+ 

nests at a single colony were destroyed but left uneaten. Steele (1980) concluded that large colonies 

suffer higher predation rates than smaller ones, possibly because they are more conspicuous. Hatching 

eggs and small chicks may be killed by fire ants (Solenopsis spp.) on islands off the Texas coast (Burger 

and Miller, 1977). An adult ibis was attacked and killed by a northern harrier in Kansas (Rivers, 2000). 

Nests, eggs, and young suffer from human disturbance, overgrazing by livestock, and, especially 

in the SLV, heavy predation from magpies, ravens, and raccoons. Breeding populations in Texas, Utah, 

and Oregon declined in the 1970s probably due to pesticides, but increased in the 1980s and 1990s. The 

USFWS has at various times designated the white-faced ibis as a “sensitive species,” a “species of 

management concern,” and a “category 2 candidate” for possible listing as a threatened species, based

on the birds’ limited number of consistent breeding sites and uncertain status (Sharp, 1985). Ryder 

(1998) urged Refuges in Colorado to adopt management controls to keep the white-faced ibis population 

in the state viable.

Hurricanes, thunderstorms, and prolonged heat waves also can disrupt breeding at individual 

sites (Belknap, 1957; Bryant, 1983; Kingery, 1988). Botulism regularly sickens and kills ibises in Utah 
63

(Wetmore, 1918). Until 1918, white-faced ibises were harvested legally in California, Oregon, Texas, 

and elsewhere in the United States for market and personal consumption (Bent, 1926; Pearson, 1936; 

Ryder, 1967). The current extent of illegal hunting in the United States is unknown, but is probably 

negligible. Hunting on wintering grounds in Mexico, despite legal protection (Ryder, 1967), continues 

to this day (recent band returns), although the impact on populations is unclear. 

DDT-DDE contamination causes eggshell thinning and abnormal eggshell ultrastructure, leading 

to cracking, denting, or crushing of eggs during incubation. DDE residues > 3-4 parts per million (ppm) 

in white-faced ibis eggs have been linked to high incidences of damaged eggs, lower hatching success, 

reduced reproductive output, and possibly delayed breeding (Ryder and Manry, 1994). Other effects of 

DDT-DDE on adults is unclear. At Carson Lake, Nevada, DDE contamination apparently reduced the 

potential fledging production by at least 20 percent in 1985 and 1986 (Henny and Herron, 1989; Henny, 

1997). Despite the 1972 ban on the use of DDT in the United States, DDE concentrations have remained 

high in white-faced ibis populations nesting in the Great Basin, most likely as a result of continued DDT 

use on birds’ wintering grounds in Mexico (Capen, 1977; Henny and others, 1985; Henny and Herron, 

1989). DDE-induced eggshell thinning in white-faced ibis is still considered a continuing problem in the 

western United States (King and others, 2003; Yates and others, 2010). 

High concentrations of dieldrin in the brains and others tissues were linked to extensive 

mortality of nestling and juvenile white-faced ibises on the Texas coast in the early 1970s, before the 

use of dieldrin (aldrin) in local rice fields was phased out in the mid-1970s (Flickinger and Meeker, 

1972; King and others, 1980). Potentially hazardous levels of mercury and selenium, accumulated 

within the breeding range, were detected in ibises nesting at Carson Lake, Nevada, but the effects of 

these contaminants on survival and reproduction were unknown (Henny and Herron, 1989). 

Diversion of natural water supplies away from existing wetlands to irrigation projects, especially 

during severe droughts, has led to temporary or permanent abandonment of traditional colony sites for 

ibises (Weller and others, 1958; Herron and Lucas, 1978; Ryder and others, 1979; Henny and Herron, 

1989). After severe drought, emergent vegetation may require several years to regenerate before it can 

be used by nesting ibises (Ryder, 1967; Herron and others, 1981). In the Great Basin, loss of wetland 

feeding habitat may be partly offset by the creation of irrigated farmlands, in which ibises forage 

extensively (Capen, 1977; Bray, 1986; Alcorn, 1988; Bray and Klebenow, 1988). The effects of wetland 

destruction on ibis populations in Mexico are unknown but are cause for concern (Sharp, 1985). 

In Utah and Nevada, some breeding sites have been severely damaged by cattle grazing and 

trampling (Weller and others, 1958; Herron and Lucas, 1978). In Texas and Louisiana, vegetation of 

some colony sites has been damaged by the introduced nutria (Myocaster coypus), resulting in delayed 

nesting or disuse of the site (Belknap, 1957; Ogden, 1978). In some Federal and state wildlife Refuges, 

burning of emergent vegetation to create open water conditions for waterfowl has reduced nesting 

habitat for ibises (Sharp, 1985). 

Humans entering active colonies may cause partial or total desertion, particularly during nestsite selection, nest-building, and incubation. Unattended eggs and small chicks are highly vulnerable to 

avian predators and to chilling or overheating (Belknap, 1957; Kotter, 1970; Kaneko, 1972; Capen, 

1977; King, 1978; Schreur, 1987). Biologists working in active colonies should carefully schedule and 

limit visits to avoid disturbance during sensitive nesting states.
64

Management

Response to Management

There was not a lot of information in the literature about white-faced ibis responses to different 

management schemes such as mowing, haying, and burning. Water-related management and agricultural 

practices were the most frequently encountered management themes. Competition over water rights

amongst irrigation projects, municipalities, and private hunting clubs, especially during droughts, poses 

a continuing threat to several important colony sites in the Great Basin, including some sites in state and 

Federal Wildlife Refuges. Beginning in the late 1980s, negotiations among users have improved water 

supplies to important breeding areas in Lahontan Valley, Nevada (Sharp, 1985; Colby and others, 1991) 

and Harney Basin, Oregon (Ryder and Manry, 1994). During the 1989 drought, limited water allotment 

to Carson Lake, Nevada, was concentrated in a single management unit to accommodate ibis nesting; 

breeding success and total fledgling production were exceptionally high despite drought conditions 

(Herron and others, 1989). 

Irrigated fields can be valuable feeding-sites for white-faced ibises (Bray and Klebenow, 1988). 

In areas where historical feeding-sites of ibises have diminished, flood irrigation of agricultural fields 

could be maintained or created for their use. Ibises showed a greater preference for alfalfa fields and less 

preference for corn, wheat-barley, pastures, and oat fields in the Lahontan Valley, Nevada. Farming 

practices can also enhance earthworm populations, a preferred food source for ibises (Bray and 

Klebenow, 1988). Applying organic fertilizers to fields would be favorable for earthworms, but 

pesticide spraying, burning, disking, and mowing of fields should be minimal (Edwards and Lofty, 

1977). However, in Utah, ibises are blamed by ranchers for harming alfalfa crops by trampling irrigated 

fields and causing soil compaction (Ryder and Manry, 1994). Alcorn (1988) described ibises flattening 

standing alfalfa in Nevada, making it difficult or impossible to cut alfalfa for hay. 

Recommendations

The U.S. Fish and Wildlife Service provided management guidelines for the Great Basin 

population of white-faced ibises (Sharp, 1985). Their overall goal was to maintain an average Great 

Basin population of at least 7,500 breeding pairs, emphasizing the maintenance of large breeding 

colonies. Specifically, they recommended providing habitat, including sufficient water to maintain 

emergent vegetation growth at Carson Lake Pasture, Stillwater NWR, Malheur Lake NWR, and the 

Great Salt Lake. A selection of their recommendations include: 

• At Carson Lake Pasture, maintaining a 400-hectare stand of dense emergent vegetation with 5.7 

million to 8.6 million cubic meters of water should provide adequate habitat of 3,000 nesting 

pairs. 

• Work cooperativly with other land management agencies and private landowners to maintain 

emergent vegetation sufficient for approximately 500 pairs at Malheur Lake (NWR). 

• At Stillwater NWR, integrate ibis production with diving duck production objectives and refrain 

from burning emergent growth in ibis colonies, especially in the spring. 

• In all areas, organize and manage water delivery to maintain at least some stable hardstem 

bulrush (Scirpus acutus) habitat in extremely wet and extremely dry years. 

• Where possible, provide shallow water feeding areas (depth up to 15 cm), especially when 

young have hatched and are being fed (June and July). 

• Eliminate cattle grazing from ibis nesting marshes

• Incoroporate nesting requirements for ibises into marsh/grassland/fire management plans 
65

• Monitor ibis nesting efforts on an annual basis. 

Schreur (1987) studied the reproductive success of colonially-nesting Ciconiiformes, including 

the white-faced ibis, and the effects of public viewing in the SLV, Colorado between April 1984 and 

August 1985. The following recommendations were made: 

• “Heronries in the SLV should be censused once a year to monitor breeding population trends. 

The number of nests in each heronry should be counted on one trip during the last week of May 

or the first week of June. All heronries should be censused to account for movement of breeding 

adults among heronries. A nest census is the preferred method of monitoring breeding 

Ciconiiformes populations over measuring reproductive success because it causes less 

disturbance and is less expensive. In addition, the reliability of reproductive success estimates 

are uncertain due to unidentified renests, unmeasured mortality of nestlings after 10 days, and 

unknown effects of observers. 

• No recreational uses of wetlands containing heronries should be allowed from 1 April through 

15 May. Ciconiiformes are selecting nest sites, constructing nests, and laying eggs during this 

time. Ciconiiformes will desert nests much more readily during this time than during the nestling 

stages (Vos, 1984). 

• Public access to heronry wetlands from 16 May through 15 August should be by guided tour or 

in motor vehicles only. Both of these methods of public viewing restrict visitor disturbance and 

access. There is evidence that motor vehicles are tolerated at closer distances than persons on 

foot. 

• Water levels should be maintained at a constant depth from 1 April through 15 August to 

prevent flooding or abandonment. Wetlands where Ciconiiformes forage should be drawn down 

periodically to ‘recharge’ the soil. Drawdowns decrease anaerobic soil conditions and increase 

the productivity of wetlands. 

• Carp should be controlled to increase water clarity and aquatic plant production. Clear water 

enhances the visibility of prey to visually orientated predators. Enhanced aquatic plant densities 

will result in more prey availability.

• Hardstem bulrush should be encouraged to grow in small ‘island’ stands (from 20 m to 100 m in 

diameter) and stands larger than 0.5 ha should include openings at least 20 m in diameter. 

Hardstem bulrush is preferred to other emergent because of superior support characteristics.” 

Research Priorities

The general behavior of white-faced ibises is not known, including communication, pair 

formation, pair-bond duration, breeding and natal dispersal, and time budgets (Ryder and Manry, 1994; 

Dark-Smiley and Keinath, 2003). Monitoring of pesticides and other contaminants should also be 

maintained, and specifically the effects of these contaminants on ibis mortality and reproduction should 

be studied (Ryder and Manry, 1994; Dark-Smiley and Keinath, 2003). Dark-Smiley and Keinath (2003) 

also suggested a need for quantifying the resources necessary for colony establishment and successful 

breeing of ibises. For example, studies that would determine the thresholds of wetland size, hydrology, 

and water quality that can support successful colonies would greatly benefit land managers seeking to 

maintain colony health and facilitate colony formation for ibises (Dark-Smiley and Keinath, 2003). 
66

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V. Black Tern (Chlidonias niger) 

Introduction

The black tern (Chlidonias niger 

surinamensis) has been a species of concern in 

North America because of continental

population declines, particularly since the 1960s 

(Shuford, 1999). This species was proposed for 

listing under the Federal Endangered Species 

Act, but the U.S. Fish and Wildlife Service 

(USFWS) concluded that there was not enough 

information to make a listing determination 

(U.S. Fish and Wildlife Service, 1991). The 

black tern has a Heritage status ranking of 

“imperiled’ in the state of Colorado (Naugle, 

2004). Loss of wetlands on breeding grounds and migration routes is probably a major cause of 

population declines in this species, but food supplies may have also been reduced through agricultural 

control of insects and overfishing in marine winter ranges (Dunn and Agro, 1995). The black tern has 

been a nontarget beneficiary of protection programs aimed at conserving habitat for upland nesting 

waterfowl. The most comprehensive of such habitat programs is the North American Waterfowl 

Management Plan (NAWMP) (http://www.nawmp.ca/; accessed 13 June 2011). No Breeding Bird 

Survey trend analyses are available for the black tern along Colorado routes from 1968 to 2009 (Sauer 

and others, 2011). 

A simple Google search on the Internet for “Chlidonias niger” revealed 432,000 results and a 

Google Scholar search provided 2,450 results (accessed 11 July 2011). “Black tern” revealed 761,000 

results through Google and 4,120 results through Google Scholar (accessed 11 July 2011). 

Natural History

Systematics and Morphology

The black tern has a holarctic distribution with two subspecies: Chlidonias niger niger

(European black tern) and C. niger surinamensis (Gmelin)(North American black tern) (Cramp, 1985; 

Dunn and Agro, 1995; American Ornithologists’ Union, 1998). It is in the Order Charadriiformes, the 

Family Laridae, and subfamily Sterninae. The species is a small (23-26 cm, 50-60 g) dark tern, 

unmistakable in alternate plumage with the head, neck, and underparts blackish (blacker in males), the 

wings (paler on coverts), back, rump, and tail smoky grey, the undertail coverts white, and the 

underwings whitish; the leading edge of the inner wing is white (Novak, 1992; Dunn and Agro, 1995; 

Howell and Webb, 1995). In the late summer and fall, the underparts are blotched black and white 

during prebasic molt. The bill is black, eyes are dark, and legs are dark reddish brown. In basic 

plumage, the head, neck, and underparts are white with a black crown patch extending down onto the 

auriculars and a blackish patch occurring at the sides of the chest. The upperparts are smoky grey, 

darker on the mantle and lesser upperwing coverts; the underwings are pale grey. Juvenile plumage is 

similar to basic plumage, but the upperparts are washed brown, the upperwing coverts are edged pale, 

and the underwing coverts are whitish. The first summer plumage resembles basic adult plumage, but 

some birds have blackish patches on the underparts. 

Black Tern, Montana, courtesy of Jessi Miller.
73

Distribution

In North America, the black tern breeds from southeastern Yukon, southwestern Northwest 

Territories, central and northeastern British Columbia, northern Alberta, northern Saskatchewan, 

northern Manitoba, northern Ontario, southern Quebec, southern New Brunswick and central Nova 

Scotia south to south-central California, northern Nevada, northern Utah, Colorado, Nebraska, northern 

Iowa, northeastern Illinois, northern Indiana, north-central Ohio, northwestern Pennsylvania, northern 

New York, northwestern Vermont, and Maine (Shuford, 1999). The species’ population is generally 

patchily distributed on the fringes of its range, particularly in the Northeast and in arid portions of the 

West. The largest populations are concentrated in zones of highly productive wetlands, particularly in 

the prairies of Alberta, Saskatchewan, Manitoba, North Dakota, South Dakota, and Minnesota (Dunn 

and Agro, 1995; Peterjohn and Sauer, 1997). 

In Colorado, breeding populations of black terns were likely more common in the past (Nelson, 

1998). A “large” colony nested in a “marshy tract along the Arkansas River” at Fort Lyon in 1964 

(Coues, 1874). Eggs presented to the Colorado College Museum documented breeding near Greeley in 

1902 (Sclater, 1912). In 1907-1908, Hersey and Rockwell found two nests along the chain of marshes 

north of Barr Lake (Rockwell, 1911; Niedrach and Rockwell, 1939), and a few pairs regularly nested at 

this site at least through 1938 (Nelson, 1998). Black terns no longer nest in these areas (Nelson, 1998). 

The most recent Colorado Breeding Bird Atlas listed probable nesting sites in all but two of the eastern 

latilongs, but no confirmed sites in the eastern part of the state. Definite breeding was confirmed only 

for San Luis State Wildlife Area and Alamosa and Arapaho NWRs (Nelson, 1998). 

The black tern winters mainly in marine and marine-coastal areas in the Americas along the 

Pacific Coast from southern Mexico east and south to Peru, and on the Atlantic coast from eastern 

Panama east along northern South America to French Guiana (Dunn and Agro, 1995; Howell and 

Webb, 1995; AOU, 1998). 

Migration

During migration, the black tern is found throughout the interior of North America south of the 

breeding range, along both coasts and the interior of Middle America, along the Atlantic coast from 

Nova Scotia south to Florida, the West Indies, and Trinidad, and in northern South America east to 

French Guiana and south to Ecuador and Peru, and often far out at sea (Dunn and Agro, 1995; AOU, 

1998). The species is casual to accidental in Hawaii, Alaska, Newfoundland, Prince Edward Island, 

Clipperton Island, Bermuda, Chile, Brazil, Uruguay, and northern Argentina. 

Black terns gather at favorite feeding sites after young fledge and then migrate singly or in small 

groups, mainly inland on a broad front through the United States (Dunn and Agro, 1995). Large flocks 

of up to thousands may form, probably where food is concentrated, particularly in coastal or marine 

habitat south of the breeding range and at a few favored sites in the interior of the western U.S 

(Carmona and others, 2011). No data are available on the length of time terns remain at migratory 

stopover sites (Dunn and Argo, 1995). Fall movements may begin by late July, and most birds leave 

northern breeding grounds by mid- to late August and are scarce in the United States after September. 

Spring migration is more rapid and less coastal than in fall, occurring primarily on a broad front through 

the interior of the United States. Spring movements occur primarily from mid-April through late May 

with stragglers into June. The degree to which terns exhibit fidelity to migratory stopovers or particular 

portions of the marine winter range are not known (Dunn and Argo, 1995). 
74

Habitat 

Black terns nest in shallow, highly productive wetlands with emergent vegetation in freshwater 

(sometimes brackish or alkaline) marshes, along prairie sloughs, lake margins, edges of islands or slowmoving rivers, wet meadow, bogs, shrub-swamps, and, in California, cultivated ricefields or flooded 

fallow fields (Dunn and Agro, 1995). Weller and Spatcher (1965) reported that during wetland 

succession the hemi-marsh stage (roughly 50:50 open water and vegetation) was ideal for most marshnesting birds, including black terns. The interspersion of water and vegetation and the size of the water 

area was perhaps of more importance than the ratio of water to cover. Studies by Tilghman (1980), 

Chapman-Mosher (1986), Rabenold (1986; 1987), Hickey (1997), Hickey and Malecki (1997), and 

Mazzocchi and others (1997) also support the conclusion that black terns generally select nest sites with 

an approximate 50:50 vegetation cover to open water ratio that is well interspersed with water; open 

water, though, can range from 5 to 95 percent of the area of breeding marshes (Powell, 1991). 

Black terns can rapidly colonize restored marshes or ones rejuvenated after drought, and 

numbers may initially increase as vegetative conditions improve and later decline as the vegetation 

deteriorates (Weller and Spatcher, 1965; Weller, 1979; Hemesath and Dinsmore, 1993). The 

relationship of marsh size and landscape features to numbers of black terns is complex (Shuford, 1999). 

Brown and Dinsmore (1986) found black terns preferred marshes >20 ha and only rarely used smaller 

marshes (5 ha minimum) unless they were part of larger wetland complexes. Naugle (1997) used GIS 

techniques to develop a model of important black tern habitat in eastern South Dakota and found that 

terns nested in larger basins and wetland complexes. He used a 50 percent probability of occurrence 

(from a logistic regression analysis) as a conservative estimator, and delineated the minimum area 

requirement for the black tern as a semipermanent wetland basin of 12.4 ha. Naugle (1997) also found 

that the area requirements of black terns varied in response to the structure of the wetland landscape. 

Black terns did not widely use wetland landscapes with a low density of primarily small wetlands, 

where few nesting wetlands occurred and potential food sources were spread over large distances. 

Instead, their wetland area requirements were small (6.5 ha) in high wetland density landscapes with a 

mixture of large and small wetlands; 32.6 ha in landscapes of predominately large wetlands; and, 15.4 

ha in landscapes with mostly small wetlands. Black terns were also more likely to occur in wetlands 

with surrounding grasslands that were <50 percent tilled for agriculture (Naugle, 1997). The average 

one-way foraging distance of black terns from colonies range from about 1 to 4 km (Chapman-Moser, 

1986), which may suggest why smaller marshes support terns only when those small marshes are part of 

a larger mosaic of marshes. Hickey and Malecki (1997) considered black terns to be area dependent 

because of the species’ loose coloniality and scattered nest spacing.

Black terns use a wide variety of vegetation types in nesting areas, including cattails (Typha

spp.), bulrushes (Scirpus spp.), burreed (Sparganium spp.), sedges (Carex spp.), rushes (Juncus spp.), 

spikerush, pickerelweed (Pontederia spp.), smartweed (Polygonum spp.), reed-canary grass (Phalaris

arundinacea), tufted hairgrass (Deschampsia caespitosa), arrowhead (Sagittaria spp.), spatterdock 

(Nuphar spp.), water lilies (Nymphaea spp.), wild rice (Zizania aquatica), marsh horsetail (Equisetum

fluviatile), cultivated rice (Oryza spp.), and, in northern areas, buttonbrush (Cephalanthus occidentalis) 

and willows (Salix spp.) in shrub-swamps (Dunn and Agro, 1995). In Colorado, black terns breed in 

large cattail marshes next to open water (Andrews and Richter, 1992). Rockwell (1911) described a nest 

at Barr Lake as built on a dense carpet of cattails blown over by the wind about 3 m from shore in a 

sparsely covered spot amid dense cattail growth and over about 15 cm of water, another on a large 

round top of a duck blind floating in almost waist deep water at the edge of cattails in a small “rushbound” pond, and a third on a mass of dead floating cattails. 
75

Ideal nest site characteristics reduce nest loss from wind and waves and provide adequate cover 

for chicks (Chapman-Mosher, 1986; Dunn and Agro, 1995), but presumably also provide camouflage to 

incubating adults without greatly hindering their entry to the site or reducing their visibility of 

approaching predators. Nest sites allow for easy access of chicks and adults to open water even after 

vegetation density has increased by seasonal growth; dense homogeneous stands of vegetation are 

generally avoided (Shambaugh, 1995). Vegetation around nest sites generally varies from sparse to 

moderately dense (Shuford, 1999). Emergent vegetation at nests grows from about <0.25 to 0.5 m high 

at nest initiation to >1 m before hatching (Dunn and Agro, 1995). Nest sites are usually adjacent to 

expanses of open water; mean distances of nests from open water have ranged from 0.1 to 35.9 m 

(Dunn, 1979; Novak, 1990; Mazzocchi and Hickey, 1997; Hickey and Malecki, 1997). Black terns may 

choose an optimum distance from both the water edge and upland edge that reduces the effects of wind 

and wave action and predators (Hickey and Malecki, 1997). Wetland habitat with lots of patch 

fragmentation and increased edges might negatively influence reproductive success. Water depth 

averages 25 to 134 cm, but depth varies considerably among sites, habitats, and years (Gould, 1974; 

Stern, 1987; Novak, 1992; Laurent, 1993; Faber, 1996; Hickey and Malecki, 1997; Mazzocchi and 

others, 1997). Areas of open water or sparse vegetation are used for foraging (Chapman-Mosher, 1986; 

Shambaugh, 1995). Water clarity at nesting and foraging sites may also be an important habitat need 

(Richardson, 1996). 

Nest success is highly variable among regions, years, and colonies, and this has made it difficult 

to identify consistent relationships between nest success and habitat characteristics (Mazzocchi and 

others, 1997). In British Columbia, experiments showed that nests surrounded by vegetation or on 

platforms suffered the least from wind and wave action or fluctuating water levels (Chapman-Mosher, 

1986). Nests in reed canary grass (Phalaris arundinaceae) survived water-level fluctuations better than 

those in other habitats, and fledging success was greatest in areas with the shortest plants, which 

included horsetail (Equisetum spp.). In northern New York, water levels at successful nests were not 

significantly different than levels at failed nests (Mazzacchi and others, 1997). Conversely, in 

Wisconsin and Minnesota, Laurent (1993) and Faber (1996) found that failed nests had significantly 

lower minimum water depths than those that successfully hatched young. In Oregon, nest success and 

fledging rates did not differ among habitats of varying vegetative composition and water depth (Stern 

and others, 1985). Several other studies failed to find a correlation between nest success and nest-site 

characteristics, including Bergman and others (1970), Dunn (1979), and Hickey (1997). 

Naugle and others (1999) conducted bird surveys throughout South Dakota in 1995 and 1996 to 

assess the effects of woody vegetation encroachment on nongame wetland birds. They found that 

wetland bird species richness decreased as the extent of woody vegetation encompassing wetland 

perimeters increased. Logistic regression analyses indicated that black terns were less likely to occur in 

wetlands surrounded by trees than in more open wetlands. 

Migrating black terns in the United States use freshwater lakes, rivers, and other interior 

wetlands (Shuford, 1999). They also forage over plowed fields and coastal wetlands (Dunn and Agro, 

1995). In arid areas of the West in fall, thousands of black terns congregate at highly productive interior 

wetlands, such as Malheur NWR, Lower Klamath NWR, and the Salton Sea (Shuford, 1999). At the 

Salton Sea, large numbers of terns forage over irrigated fields. South of the United States, black terns 

are found mainly over offshore marine waters where water is 30 to 450 m deep (up to 2000 m) in areas 

of high productivity, such as the Panama upwelling and edges of the Gulf of Mexico currents (Dunn and 

Agro, 1995). Black terns in the Gulf of Mexico concentrate over the freshwater plume of the Mississippi 

River because of the low surface-water salinity and high productivity (Ribic and others, 1997). 
76

During the winter season, black terns are found mostly in marine waters within 30 km of land, 

with occasional sightings to as much as 3500 km offshore (Dunn and Agro, 1995). They also use coastal 

areas and productive freshwater lakes near coasts.

Foraging Habits

Breeding black terns are mainly insectivorous, but fish make up a large part of the diet in some 

habitats and regions (Dunn and Agro, 1995). Fish may dominate the diet by mass and provide an 

important source of calcium (Beintema, 1997). Black terns circle low over foraging areas with slow, 

shallow wingbeats and bill pointed downwards (Goodwin, 1960). The birds often flock where food is 

concentrated. They may hover briefly before suddenly dropping or swooping to the surface to capture 

prey on the water surface, or pick insects off vegetation (Dunn and others, 2009). Sometimes, black 

terns hunt from a perch over water (Welham and Ydendberg, 1993). They may also catch insects in the 

air, especially when insect swarms occur over land. Prior to pesticide use, black terns commonly 

foraged for insects behind plows and over grain fields (Gabrielson, 1914; Pittman, 1927; Roberts, 1932; 

Brewer, 1969; James and Neal, 1986). Larson and Leyva (2007) observed 3,000 black terns foraging 

among hundreds of breaching manta rays one kilometer offshore the coast of Guerrero, Mexico. 

Breeding

Black terns nest semicolonially, placing their nests in clusters in favorable areas of marshes; 

clusters are typically about 11-50 nests but can range from two to hundreds (Dunn and Argo, 1995). 

Nests are usually 5 to 20 m apart, but sometimes as close as 1 m. Up to 25 percent to 30 percent of birds 

nest “solitarily,” or from 20 to 30 m up to 600 m from other nests. Birds from these subgroups in a 

wetland area may behave as a single colony in their flock behavior. Territories of nests are defended to 

about 2 m. Nests are small cuplike gatherings of aquatic vegetation usually built on floating substrates 

of matted or decaying marsh vegetation and commonly mixed with mud (Shuford, 2004). Floating 

substrates can also be detached root masses, logs and boards, muskrat (Ondatra zibethicus) feeding 

platforms or clippings, algae or peat mats, lily pads, dried cowpies, and old nests of grebes, American 

Coots (Fulica americana), and Forster’s terns (Sterna forsteri). These substrates are usually anchored to 

or lodged in emergent vegetation or dense beds of submerged rooted aquatics (Novak, 1992; Dunn and 

Agro, 1995). Nonfloating substrates include muskrat lodges, raised mud patches, marshy hummocks, 

rooted flattened vegetation, and upturned tree roots with attached vegetation. 

Nest site selection and nest building are rapid, with the time from colony occupation to egg 

laying being as little as 4 days (Dunn and Argo, 1995). Initial nesting attempts are fairly synchronous, 

but renesting frequently prolongs the breeding season. Egg-laying has occurred as early as 11 May in 

lowland California, where the majority of clutches are initiated in the second or third week of May. At 

higher elevations or more northerly latitudes, most clutches are initiated from the third week of May 

through the first week of June. Average clutch size is 2.6 (n=2297); 3-egg clutches account for about 65 

percent of nests, 2-egg clutches for 25 percent of nests, 1-egg clutches are uncommon, and clutches of 4 

to 6 are rare (Dunn and Argo, 1995). Eggs require 19 to 23 days to hatch (Novak, 1992; Dunn and Argo, 

1995). Earliest hatching is in early June, but most occurs in late June and early July. Chicks can fly as 

early as 18 to 19 days, most at 20 to 24 days, and some probably at 25 days (Dunn and Agro, 1995). 

Most young fledge in mid- to late July. 
77

Demography and Populations

Little is known about the demography of black tern populations in North America (Nisbet, 

1997). Nesting success varies widely across their range, but on average there is probably less than one 

chick raised per nest (Dunn and Agro, 1995). Birds can renest at least 42 km from their original nest, 

which makes estimating annual reproductive success of birds that fail on their first attempt extremely 

difficult (Mazzocchi and Muller, 1993). Nests or eggs are often lost to bad weather, effects of wind and 

waves, or changing water levels (Dunn and Agro, 1995). Chick loss is also highly variable and its 

causes are generally unknown. 

The North American Waterbird Conservation Plan provides the only available population 

estimate for black tern in North America – from 100,000 to 500,000 breeding individuals (Kushlan and 

others, 2002). This is a speculative estimate based on the judgment of species experts and information 

from the literature (Naugle, 2004). The largest breeding populations are likely in North Dakota (83,000 

to 86,000 as estimated by Igl and Johnson, 1997), South Dakota (Peterjohn and Sauer, 1997), and 

Minnesota (Baker and Hines, 1996). In Canada, black tern populations may be larger than in the United 

States (Naugle, 2004). An estimate of from 2,873 to 14,996 breeding pairs in Ontario (Austen, 1994) is 

the only regional estimate for any province in Canada. The best information on continental population 

trends (since 1966) is from the North American Breeding Bird Survey (BBS; Robbins and others, 1986). 

While the BBS is designed to survey mostly passerine species, data are collected on all species 

encountered during surveys. Interpreted cautiously because BBS data are inherently biased, there was 

indication that black terns declined survey-wide at an average rate of 3.1 percent annually from 1966-

1996. This period coincided with extensive wetland losses over much of the range (Dahl, 1990; 2000). 

Fewer declines in black terns were evident after 1980 (Peterjohn and Sauer, 1997). The BBS trends for 

black terns from 1980 to 2007 in Region 6 of the U.S. Fish and Wildlife Service were actually found to 

be increasing (an adjusted trend index of 5.26; P<0.1) with 53 routes surveyed across the region (Sauer 

and others, 2011). There were no results available on black tern trends in the BBS specifically for the 

state of Colorado (data were too sparse). However, during the Colorado Breeding Bird Atlas surveys, 

the largest number of black terns were found in the San Luis Valley (SLV) (Shuford, 1999). 

Threats

Sources of Mortality

Predators of black tern eggs or chicks are the great blue heron (Ardea herodias), black-crowned 

night-heron (Nycticorax nycticorax), great horned owl (Bubo virginianus), mink (Mustela vison), and 

Norway rat (Rattus norvegicus) (Dunn and Agro, 1995). Other potential predators of eggs or chicks are 

northern harrier (Circus cyaneus), ring-billed gull (Larus delawarensis), American crow (Corvus

brachyrhynchos), common raven (Corvus corax), raccoon (Procyon lotor), muskrat, long-tailed weasel 

(Mustela frenata), and snapping turtle (Chelydra serpentine) (Gerson, 1988; Novak, 1992; Dunn and 

Agro, 1995). Adult birds have been taken by northern harrier and a large fish (Dunn and others, 2009). 

Adults have also been attacked by common ravens (Dunn and others, 2009).

Nest losses from flooding are common in black tern breeding colonies (Gilbert and Servello, 

2005). Mosher (1986), Macikunas (1993), and Hickey (1997) reported nest losses from flooding during 

each year of the four, seven, and two years of their studies, respectively. Heath (2004) suggested that 

predation was the primary factor limiting chick survival in Maine colonies of black terns.
78

Loss and Degradation of Habitat

The major factors cited as causes for black tern population declines throughout North America 

are loss, degradation, isolation, and fragmentation of habitat via drainage for agriculture or development 

(Shuford, 1999). Wetland loss in the United States has already been addressed earlier in this document. 

Rabenold (1987) expressed concern that habitat loss resulted in local marshes that were too small or 

were not part of larger marsh complexes and therefore were not suitable habitat for black terns. 

Degradation of habitat can occur by plant succession, increasing or decreasing water levels, introduction 

of exotic species, and reduction in water quality, which may alter both the food web and vegetative 

structure of wetlands (Novak, 1992). Competition for scarce water supplies in the arid West may impact 

breeding marshes. Mazzocchi and Muller (1995) suggested that as the quantity and quality of nesting 

habitat declines, the negative impact of factors such as predation, human disturbance, and adverse 

weather may increase, thereby reducing productivity. Little is known about loss or degradation of winter 

or migratory staging habitat. Nisbet (1997) observed that because black terns spend 8 to 9 months of the 

year on salt water, factors affecting the wintering grounds could be equally or more important than ones 

on the breeding grounds. For instance, black terns may be adversely affected by overfishing in their 

wintering habitats (Dunn and Agro, 1995). 

Other threats to black tern populations cited in the literature include human disturbance, disease, 

predation, pesticides or other contaminants, introduced species, and inadequacy of existing regulatory 

mechanisms (Shuford, 1999). Human disturbance can expose chicks to adverse weather or directly 

destroy the nests. Black terns sometimes die of botulism, but this disease or the various parasites 

harbored by the species do not seem to be major causes of mortality (Novak, 1992; Dunn and Agro, 

1995). Predation could also limit reproductive success, and Brewer (1991) suggested that increasing 

predator populations might be one of the multiple factors responsible for population declines in 

Michigan. Dunn and Agro (1995) and Weseloh and others (1997) reviewed the literature on 

concentrations of contaminants, such as organochlorine, PCBs, and metals, in black tern eggs, but found 

no evidence of effects of such contaminants oon reproduction. However, pesticides may indirectly affect 

black terns by reducing insect foods (Shuford, 1999). 

The introduced marsh plant purple loosestrife (Lythrum salicaria) crowds out native emergent 

and forms stands too dense for nesting black terns (Seyler, 1991; Whitt and others, 1999; Blossey and 

others, 2001). Terns are apparently responding favorably to recent vegetation removal in the Columbia 

Basin of Washington, however, where purple loosestrife and phragmites (Phragmites australis) have 

been choking out marshes (Shuford, 1999). 

There are currently no mandates or regulatory mechanisms for protecting the species and its 

habitats on the winter range (Shuford, 1999). 

Management

Response to Management

Little is known about the effects of burning, mowing, or grazing on black terns (Zimmerman and 

others, 2002). Because wetland suitability varies with yearly fluctuations of water levels and subsequent 

changes in vegetation, black terns use different wetlands or different locations within wetlands from 

year to year (Bailey, 1977; Carroll, 1988; Dunn and Agro, 1995). In South Dakota, black tern presence 

was positively correlated with the total area of semipermanent wetlands in the surrounding landscape 

(Naugle and others, 1999; 2000; 2001). Thus, it is important to protect and maintain wetland complexes 

(Dunn and Agro, 1995). In Nebraska, successional changes (for example, wooded vegetation 
79

encroachment) within the channel of the Platte River have reduced the quality of river edge vegetation 

formerly occupied by black terns (Faanes and Lingle, 1995). 

In the northern Great Plains, cattails (Typha spp.) have overgrown many wetlands, contributing 

to a decline in the numbers of black terns (Linz and others, 1994; Linz and Blixt, 1997). Since 1991, 

wildlife agencies have applied glyphosate-based herbicide to cattail-dominated wetlands with the 

objectives of increasing duck abundance and reducing crop-depredating blackbird populations. In 1990 

and 1991, glyphosate-based herbicide was aerially applied to cattail-dominated wetlands in North 

Dakota to assess the influence of habitat changes on birds. Numbers of black terns were positively 

correlated with open water and dead cattails. Black tern numbers were also positively related to the 

number of mallards, blue-winged teal, redheads, and yellow-headed blackbirds, suggesting that there 

were some common habitat requirements among these species. Researchers concluded that cattail 

management programs designed specifically to enhance duck use and decrease red-winged blackbird 

numbers would probably benefit black terns (Linz and others, 1994; Linz and Blixt, 1997). 

Restored wetlands can provide nesting habitat for black terns (Svedarsky, 1992; Delehanty and 

Svedarsky, 1993). Black terns nested in a newly restored wetland in Minnesota each year of a 3-year 

study. The rapid colonization of the restored wetland was attributed to the availability of flooded, dead 

vegetation (Delehanty and Svedarsky, 1993). In Iowa, black terns were present, but did not nest in 

wetlands that had been restored for 1-4 years; the study did not examine restored wetlands older than 4 

years (VanRees-Siewert, 1993; VanRees-Siewert and Dinsmore, 1996). Preliminary analysis of data 

from North Dakota and South Dakota indicated that black terns are more common on natural wetlands 

than restored wetlands (Naugle and others, 2001). 

A study conducted by Shealer and Haverland (2000) revealed that neither hatching success 

(proportion of nests from each treatment group that hatched >1 egg) nor fledgling success (proportion of 

chicks fledged from all eggs that hatched) were negatively affected by investigator disturbance 

(repeated nest visits and trapping and banding of adults). Black tern mortalities due to utility wire 

collisions have been reported, but appear to occur infrequently (Thompson, 1978; Dunn and Agro, 

1995). 

Recommendations

To provide suitable habitat for black terns in most regions of North America, managers should 

maintain or create emergent marshes with an approximate 50:50 vegetation: open water ratio with a 

good interspersion of vegetation and water (Shuford, 1999). Brown and Dinsmore’s (1986) work also 

suggests that black terns would benefit from creation or restoration of marshes >20 ha or marshes of >5

and <20 ha situated within a wetland complex. Naugle’s (1997) research indicated that the overall 

minimum area requirement for black terns is 12.4 ha of semipermanent wetland, but this requirement 

can be lowered to 6.5 ha in landscapes of high wetland density that contain a mixture of large and small 

wetlands. For western New York, Hickey (1997) suggested that in impoundments managed for black 

terns, habitat should be in unbroken patches >10 ha, especially in marsh units <20 ha. In large marsh 

units, maintaining habitat patches of >20 ha should be attempted. 

Below is a selection of citations describing more responses to different management schemes 

and recommendations found in the literature: 

• The Tonawanda complex in western New York is managed for black terns by draining and 

discing to favor burreed, an important nesting habitat, and muskrats, which provide nest sites 

and open up the marsh vegetation (Adams, 1990; Hickey, 1992; Seyler, 1991; Hickey, 1997; 

Hickey and Malecki, 1997). Marshes are drawn down in May, disced in July or August, and 

subsequently reflooded. Hickey (1997) recommended that marshes managed for black terns 
80

should be placed in a 4- to 6-year cycle of drawdown followed by flooding in years 2 to 5. In the 

first year following flooding, water levels should be kept higher than normal to inhibit 

undesirable vegetation, such as purple loosestrife, and to allow muskrat populations to build up. 

After intentional or natural removal of water for one or more seasons, black terns almost always 

colonize impoundments the year following reflooding, and peak numbers usually occur in the 

second and third years after reflooding (Hickey and Malecki, 1997). In the first year after 

drawdown, vegetation responds, muskrat populations grow, and black tern nesting is probably 

limited by lack of suitable nesting substrates. In the second and third year, muskrat feeding and 

house-building activities remove vegetation, improving the interspersion of vegetation to water 

and providing nesting substrates. However, some researchers have questioned the extent to 

which muskrats are beneficial to black terns, noting that muskrats often cut down large areas of 

protective marsh vegetation, making those areas unsuitable for tern nest sites, and usurp floating 

nests or artificial nest platforms (Shuford, 1999). 

• The black tern was chosen as an indicator species in the Lake Ontario-St. Lawrence River 

hydrosystem for its response to water-level changes (Desgranges and others, 2006). Black terns 

were more abundant in marshes with reduced water-level fluctuation. Because black terns build 

crude, typically floating nest platforms of dead aquatic vegetation, they can be particularly prone 

to inundation during flooding. Where water-level increases were rapid or moderate, observed 

breeding populations of black terns were reduced by 84 percent or more (Desgranges and others, 

2006). 

• Timmermans and others (2008) examined associations between breeding marsh bird abundances 

and Great Lakes hydrology. From 1995-2002, Lakes Erie and Huron-Michigan experienced 

water-level declines. Abundance of black terns declined within the marshes of these lakes during 

the same period. They concluded that long-term changes and annual water-level fluctuations are 

important abiotic factors affecting the abundance of some marsh-dependent birds in Great Lakes 

coastal marshes.

• Gilbert and Servello (2005) suggested that management actions that lessen large water-level 

increases during the period of tern nesting would reduce the proportions of nests lost to flooding 

in Maine. Early spring water levels that are exceptionally low may increase the risks of nest 

losses to flooding if nests are established on more stable substrates lower in the wetland basin. 

Floating nesting platforms placed in nesting areas prior to the breeding season may be beneficial.

• In Minnesota, black terns colonized a marsh the year following reflooding and peak populations 

occurred in the second and third years after reflooding (Delehanty and Svedarsky, 1993). 

• Seyler (1991) and Hickey (1997) felt that elevated perches would benefit black terns because 

they use them for copulation, resting, and sites for feeding recently fledged young. 

• Linz and Blixt (1997) recommended the use of aerially-sprayed herbicides to open up cattaildominated marshes to manage for wetland features that benefit black terns on a local scale. 

Aquatic herbicides could be used with a creativity and precision difficult to achieve with other 

methods, and the time between treatments may be from 8 to 10 years. They also suggested that 

vegetation management treatments be staggered within and among wetland complexes to 

diversify successional stages of emergent vegetation and maintain avian diversity on a more 

regional scale. 

• Other alternatives for vegetation management of benefit to black terns that could be considered 

include regulation of muskrat populations, water-level control, livestock grazing, prescribed 

burning, mowing, discing, crushing, excavating, and blasting with explosives (Hickey, 1997; 

Linz and Blixt, 1997). 
81

• Removal of purple loostrife by uprooting plants, water-level manipulation, mowing, burning, 

flooding, or herbicide application can eliminate small and young stands but is costly, requires 

continued long-term maintenance, and, in the case of herbicides, is nonselective and 

environmentally degrading (Malecki and others, 1993). 

• Faber (1996) studied use and reproductive success of black terns on artificial nest platforms 

along the Mississippi River from 1989 to 1991. Platforms were significantly more successful at 

hatching young than were natural nests in 1990, but results were equivocal for the combined 

period 1989 to 1991 (Faber, 1996). In northern New York, Mazzocchi and Hickey (1997) found 

nest success on artificial platforms was higher than on natural substrates in 1995 and lower in 

1996. In perhaps the first attempt at providing artificial nest platforms in 1925, R. J. Niedrach 

scattered some square pieces of board in a cattail marsh at Mile High Duck Club in the Barr 

Lake drainage, and half a dozen nests were constructed on them (Bailey and Niedrach, 1965). 

Shealer and others (2006) studied the effect of floating nest platforms on breeding performance 

of black terns and found that eggs laid on platforms were significantly larger than those laid on 

natural substrates, suggesting that the platforms were occupied by healthier birds. More research 

is needed on the effects of artificial nest platforms and breeding success of black terns (Shuford, 

1999). 

Research Priorities

The following research priorities on black terns were found in the literature: 

• Additional research on the effects of artificial nest platforms and breeding success of black terns 

(Shuford, 1999). 

• Regional population and habitat surveys (Naugle, 2004). 

• Collection of population demographic data (Naugle, 2004). This includes measurement of adult 

survival rates, characterizing chick survival rates, and the contribution of renesting to annual 

productivity (Naugle, 2004). 

• Ecology and survival of black terns on the wintering grounds where they spend six to seven 

months of each year (Naugle, 2004). 

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VI. Short-eared Owl (Asio flammeus) 

Introduction

The short-eared owl (Asio flammeus) is one of the 

world’s most widely distributed owls (Holt and others, 

2006). This owl is an open country, ground-nesting 

species that inhabits marshes, grasslands, and tundra 

throughout most of North America and Eurasia. It also 

breeds in South American grasslands and on islands such 

as Iceland, the Hawaiian chain, and the Galapagos. These 

far-ranging owls are prone to nomadic wanderings as 

well, which has resulted in far-flung island populations 

and occasional sightings of these owls on ships at sea 

(Gray, 1945). In Colorado, short-eared owls range over a 

variety of open habitats including grasslands, marsh 

edges, shrub-steppes, and agricultural lands (Holt and 

Leasure, 1993). Reproduction and population dynamics of this species are closely linked to the density 

of its primary prey, small mammals such as Microtus (voles) (Wiggins and others, 2006). Like other 

birds that depend on such a fluctuating food source, the short-eared owl shows considerable local 

variation in numbers and reproductive success. Although the literature on its diet in the nonbreeding 

season is extensive, few studies have thoroughly addressed other facets of its biology, ecology, and life 

history. This owl is active both day and night and it tends to hunt low above the ground. It generally 

uses acoustical cues to locate prey, but it also can rely on vision as well. 

Historical references generally portray short-eared owls as common breeders on the Great Plains 

(Cadman and Page, 1994). They were the most common raptor noted during explorations of the Dakotas 

in the late 1800s (Allen, 1875). In Kansas, they were noted as common in the late 1800s (Goss, 1886). 

Although such references must be interpreted with caution due to the possibility of observations made 

during years of peak owl abundance, the general picture drawn from studies on the northern Great 

Plains is that short-eared owls were formerly a common species. In recent decades, the numbers of 

short-eared owls have declined in many areas of North America, especially the northeastern United 

States (Wiggins and others, 2006). The major cause appears to be habitat loss owing to human activities. 

It could also be susceptible to increased levels of predation because of ground-nesting. Although this 

species is listed as a Bird of Conservation concern by the USFWS both nationally and within Regions 3, 

4, 5, and 6, as well as a priority species within many state and regional Partners in Flight bird 

conservation plans, no research programs on the conservation status of short-eared owls are currently 

underway (Wiggins, 2004). The USDA Forest Service (USFS) Region 2 (Rocky Mountain Region) 

classifies the short-eared owl as a sensitive species. The species is listed as a Priority Species in 

wetlands in both the Colorado and Wyoming Partners in Flight (PIF) plans (Beidleman, 2000; Cervoski 

and others, 2001). 

A simple Google search on the Internet for “Asio flammeus” revealed 294,000 results and a 

Google Scholar search provided 4,320 results (accessed 11 July 2011). “Short-eared owl” revealed 

493,000 results through Google and 6,070 results through Google Scholar (accessed 11 July 2011). 

Short-eared Owl, courtesy of Eleanor Briccetti.
90

Natural History

Morphology & Systematics

The short-eared owl is medium-sized (33-43 cm) with females about 20 percent larger than 

males (Wiggins and others, 2006). The adult plumage is dorsally mottled brown and buff and resembles 

dried grasses. Ventrally, it is whitish to rust-colored with dense vertical streaking on the breast, thinning 

out on the sides and flanks. The short-eared owl is morphologically similar to the long-eared owl (Asio

otus), but is larger overall and multilocus protein electrophoresis indicates a large genetic distance 

between the two species (Randi and others, 1991). These two species are usually separated by diel 

activity period, with the long-eared owl being primarily nocturnal and the short-eared owl more 

crepuscular (Wiggins and others, 2006). 

Ten subspecies of short-eared owls are recognized following Holt and others (1999) and 

Dickinson (2003). Two of these subspecies occur in North America, one breeding and another only a 

visitor to southern Florida. A. f. flammeus (Pontoppidan, 1763) is the subspecies that breeds across 

North America, Europe, northern Africa, and northern Asia. 

Migration 

In cooler parts of their range, including Colorado, short-eared owls migrate seasonally, and 

Colorado hosts more birds in winter than in summer. Sclater (1912) characterized this species as a 

winter visitor, and several later publications also mentioned short-eared owls only as winter visitors and 

transients (for example, Aiken and Warren, 1914, Niedrach and Rockwell, 1939). The species is 

considered a partial migrant but is thought to show more long-distance migrations in the northern part of 

its range (Wiggins and others, 2006). Although present year-round at mid-latitudes, USFWS band data 

indicate seasonal north/south migration and west/southwest movements for such individuals (Clark, 

1975). Migrants and nomads may flock in large numbers in response to concentrations of, or changes in, 

prey populations (Bent, 1938). For example, 60 owls seen at Monte Vista NWR in mid-August 1958 

probably represented a post-breeding flock (Boyle, 1998). 

Habitat

The short-eared owl is almost always associated with open country that supports small mammals 

(voles, lemmings), typically in large expanses of prairie and coastal grasslands, heathlands, shrubsteppe, and tundra. However, the species also uses agricultural areas. The Breeding Bird Atlas of 

Colorado recorded short-eared owls most often in grasslands, prairies, and wetlands (68 percent of 

habitat observations), less commonly in shrublands, and rarely in scrub oak or pinyon-juniper 

woodlands (Boyle, 1998). 

On the Great Plains, short-eared owls prefer to nest within large patches of relatively tall (30 to 

60 cm), dense, ungrazed grassland (Duebbert and Lokemon, 1977; Fondell and Ball, 2004). In Illinois, 

Herkert and others (1999) in Illinois found that the owls required relatively short grass (<50 cm tall) and 

most grasslands in the region needed to be mowed or burned to maintain this condition. Nests on the 

arctic plain typically are on elevated patches of tundra vegetation (Pitelka and others, 1955a). At 

Churchill, Manitoba, they often nest at the base of small birch or willow (Jehl, 2004). In coastal 

Massachusetts, short-eared owls nest in grassy dune vegetation (Holt, 1992). In Cuba, short-eared owls 

were found nesting in pasturelands, rice fields, and sugar-cane plantations (Garrido, 1984). In 

Hispaniola, one nest was found in an abandoned agricultural field, below a small bush (Guerrero, 2005). 

This species also will nest on slight ridges and mounds if present. 
91

During the nonbreeding season, short-eared owls use habitats similar to those used during the 

breeding season, but may also use large open areas within woodlots, stubble fields, fresh and saltwater 

marshes, weedy fields, dumps, gravel pits, rock quarries, and shrub thickets (Clark, 1975). In Missouri, 

the short-eared owl prefers dense grasslands less than 30 cm high (Weller and others, 1955). 

Occasionally, this species will roost in trees during the winter (Clark, 1975; Bosakowski, 1986). Winter 

areas often turn into breeding areas if food is plentiful (Clark, 1975). In coastal southeastern 

Massachusetts, short-eared owls hunt in salt marshes and the edges of salt marshes, where grass 

transitions to low brush (Iva sp.). They also have been documented foraging in open pastures and fields 

with low woody vegetation (Rosa, Prunus sp.) (Wiggins and others, 2006). 

Foraging Habits

Small mammals, particularly Microtus, are the main foods taken by short-eared owls (Craighead 

and Craighead, 1956; Clark, 1975; Colvin and Spaulding, 1983; Holt, 1993a; 1993b; Wiggins and 

others, 2006). Less frequently, this species will hunt birds. There is a strong positive correlation 

between vole abundance and owl abundance in several areas of this species’ distribution, including 

Scotland (Village, 1987) and Finland (Korpimaki and Nordahl, 1991). Short-eared owls hunt day and 

night, probably dictated by hunger, prey density, and nestlings’ hunger. Winter birds, however, may be 

essentially crepuscular (Clark, 1975). Erinaro (1973) reported that diel activity of short-eared owls 

coincided with activity periods of prey. The extent of diurnal foraging may be over emphasized, 

facultative, or due to the ease in which this owl is observed in open habitats (Wiggins and others, 2006). 

Short-eared owls hunt primarily on the wing, approximately 0.3 to 3.0 m above ground (Clark, 1975; 

Village, 1987). They also hover 2-30 m above ground or, less frequently, perch on poles or hills. They 

forage similarly to northern harriers (Circus cyaeneus) and can quickly adjust flight to drop down on 

prey. 

Breeding

Male short-eared owls may begin courtship displays during late winter and continue into the 

nesting stage (Pitelka and others, 1955a; Hammerstrom and others, 1961; Clark, 1975). Pair formation 

begins in mid-February and continues through June (Urner, 1923; 1925; DuBois, 1924; Pitelka, 1955b; 

Hammerstrom and others, 1961; Beske and Champion, 1971; Trann, 1974; Clark, 1975; Holt, 1992). 

Communal winter roosts may turn into local breeding territories (Clark, 1975; Wilson, 1995). Shorteared owls are thought to be colonial nesters, but coloniality is probably dependent on prey, nesting 

cover, and area. The first brood of the season occasionally appears as early as late March, but more 

often in April through June (Pitelka, 1955b). This species is reported to produce two broods in a season, 

but more data are needed on marked individuals to confirm this (Mikkola, 1983). 

One of the only owls to build their own nests, female short-eared owls make a shallow scrape on 

the ground and usually construct a sizable nest of grass blades, forbs, and downy feathers (Clark, 1975; 

Holt and Leasure, 1993). Owls may transport vegetation from some distance, making it possible to 

confirm breeding by observing them carrying nest materials. Because ground-nesting makes females 

and young vulnerable to predation, incubating females often sit tight until intruders approach very 

closely, making nests hard to find by flushing nesting females. Young are more precocial than those of 

most other owls; they gain weight rapidly and often leave the nest on foot when only 14-17 days old, 

well before fledging. 

The Colorado Breeding Bird Atlas verified previous reports that short-eared owls are scarce and 

local breeders in Colorado (Boyle, 1998). Breeding populations occur in North Park, the SLV, and the 

northern Great Plains (Boyle, 1998). 
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Demography & Populations

Short-eared owls may lay large clutches in response to the abundance of prey in the nesting area. 

Thus, the reproductive potential of this species is high (Wiggins, 2004). Although available data suggest 

that pre-fledging survival is relatively high for a ground-nesting species, there is little information on 

post-fledging survival. No study has measured adult or juvenile survival rates for short-eared owls. 

North American Bird Banding Laboratory recovery records are limited (n = 47), but the longevity 

record for a wild short-eared owl is 4 years, 2 months in the United States (Wiggins and others, 2006). 

In Europe, the longevity record is 12 years, 9 months for a wild bird (Cramp, 1985). 

Population status for this species is difficult to assess because individuals are nomadic and prone 

to annual fluctuations in numbers. Consequently, fixed-area census projects such as the Breeding Bird 

Survey (BBS) do not adequately sample for short-eared owls (Wiggins, 2004). An abundance map 

generated from BBS data shows low abundance (< 1 per route) over most of the breeding range 

(Wiggins and others, 2006). Numbers fluctuated between 0 (in 3 years) and one pair (in 2 years), with a 

peak of 28 pairs during a year when the lemming population was high (Pitelka and others, 1955a). In 

southeastern Massachusetts, 1985 estimates showed a total of 20-25 pairs breeding annually on islands 

off the coast (Holt, 1986a). Christmas Bird Count totals for wintering populations in New Jersey, 1976-

1997, show numbers fluctuating on 4-year cycles (likely reflecting prey cycles farther north), with highs 

statewide of about 80 individuals and lows <20 (Walsh and others, 1999). In Lowland Mainland, British 

Columbia, Christmas Bird Count numbers from 1972-1987 ranged from 170 to 5 individuals, often in 

the space of just a year or two (Campbell and others, 1990). 

Historically (1960s-early 1980s), of 7 North American regions, three (Central Southern, Prairie 

Provinces, Middle Pacific Coast) reported that short-eared owls were greatly down in numbers (Tate, 

1986). Declines were also noted in the Midwest, California, and southern Ontario (Tate, 1992). Melvin 

and others (1989) considered this owl to be the rarest and most threatened species of owl in the 

northeastern United States. The BBS regional trend analysis for short-eared owls along 19 Colorado 

routes from 1966 to 1999 showed a positive trend estimate of 5.3 (-2.7, 19.0; 95 percent CI) (Sauer and 

others, 2011). From 1999 to 2009, the trend was also positive and estimated to be 14.3 (-9.4, 72.1; 95 

percent CI). However, the regional credibility measure for this species was the category “red,” meaning 

that data were deficient because of overall low abundances per route and/or the results were too 

imprecise to detect long-term trends. 

Threats

Data are limited on causes of death of short-eared owls in North America, but occasionally the 

birds are hit by cars and airplanes, shot, or trapped (Clark, 1975; Holt, 1992). They may also suffer from 

collisions with barbed-wire fences, but the extent of this problem is unknown (Wiggins and others, 

2006). A road-killed bird in England was found be infected with Tuberculosis (Mycobacterium

tuberculosis avium; Harrison, 1943). In California, secondary poisoning from fowl cholera (Pasteurella

bacteria) passed through waterfowl, rodents, and raptors, killing 44 short-eared owls (Rosen and Morse, 

1959). Single cases of mortality due to the West Nile Virus in the midwestern United States (Fitzgerald 

and others, 2003) and avian cholera (Pasteucella multocida) in the Texas panhandle (Taylor and Pence, 

1981) have been reported. 

Short-eared owls generally show low concentrations of pesticides and contaminants with little 

effect on egg viability (such as eggshell thickness and breakage, tissue damage, or embryo mortality) or 

health of adults. Mean levels of DDE in 5 short-eared owl eggs from the Canadian prairies in the 1960s 

was 0.5 ppm (Keith and Gruchy, 1972). Although organochlorine residues magnify with successive 

trophic levels in the food chain, birds of prey (such as short-eared owls) that eat mostly herbivorous 
93

mammals usually accumulate low levels of the compounds. There is some evidence that short-eared 

owls are susceptible to organophosphate contamination. In Israel, five short-eared owls were found dead 

and four were paralyzed in a field sprayed with Azodrin (an organophosphate) to control a vole outbreak 

(Mendelssohn and Paz, 1977). 

Nesting habits and nomadism make this species particularly vulnerable to habitat loss during any 

season (Wiggins and others, 2006). Habitat loss includes conversion of open habitats to agriculture, 

grazing, recreation, housing, and resort development. Reforestation in some areas may also contribute to 

habitat loss. Predation by striped skunks (Mephitis mephitis) on eggs and nestlings was suspected in the 

decline of short-eared owls on Martha’s Vineyard Island, Massachusetts (Melvin and others, 1989) and 

disturbance of nests by domestic and feral cats and dogs is known (Tate, 1992). Interspecific 

competition with barn owls (Tyto alba) may also occur; successful nest box programs to attract barn 

owls have coincided with the decline of the short-eared owl on Martha’s Vineyard and Nantucket 

islands, Massachusetts.

Short-eared owls appear to be particularly sensitive to habitat fragmentation as they require 

relatively large tracts of grassland and are ground nesters, making them susceptible to the increased 

predation pressure that is typical within fragmented landscapes near rural developments. The 

development of conservation/management plans for this owl has been hampered by the difficulty in 

accurately assessing their local and regional status, and by a lack of information on reproductive 

success. The loss of open grassland habitats on the Great Plains and along the Pacific and northern 

Atlantic coasts has been cited in a number of studies as the primary causative factor behind population 

declines. On the Great Plains, the primary source of habitat loss has been the conversion of native 

prairie to agriculture and overgrazing of remaining grasslands (Wiggins and others, 2006). Along 

coastal areas, which includes many wintering sites, recreational use and land development have caused 

losses of nearshore marsh and oldfield habitats. 

Wiggins (2004) summarized the current and historical threats to the viability of short-eared owl 

populations in the Rocky Mountain region (Region 2, USFS) and ranked the threats as follows: 

• Loss of native grassland and wetland habitat – This has been a problem historically, but recent 

efforts (for example, Playa Lakes Joint Venture, Conservation Reserve Program [CRP]) may be 

slowing habitat loss. However, the extent to which CRP land represents suitable breeding habitat 

for short-eared owls is still unclear.

• Degradation of existing grasslands due to overgrazing by livestock – This represents a 

continuing problem on most land in the Rocky Mountain region. Efforts to reduce or to 

eliminate grazing would clearly benefit short-eared owls. 

• Degradation of grassland habitat due to fragmentation – As short-eared owls require large 

(minimum 100 ha) tracts of land for successful nesting, fragmentation of grassland habitats has 

likely had significant negative impacts on populations in the Rocky Mountain region. 

Management

Response to Management

Although disturbance generally has a negative effect on short-eared owls, periodic disturbances 

may be necessary to maintain suitable habitat. In tallgrass areas, burning or mowing every 3-5 years is 

recommended to maintain habitat for voles, the principal small rodent prey of short-eared owls (Lemen 

and Clausen, 1984; Kaufman and others, 1990). Berkey and others (1993) suggested that dense nesting 

cover in uplands could be hayed periodically to stimulate plant growth. In Illinois, short-eared owls 

preferred rotary-mowed grasslands that were 30-40 cm tall to idle grasslands with taller vegetation 
94

(Herkert and others, 1999). In tallgrass areas, mowing or other techniques are recommended to reduce 

grass heights. In Missouri, short-eared owls preferred medium to tall grasslands that were moderately 

grazed or idle (Skinner and others, 1984). Bock and others (1993) suggested that short-eared owls 

respond negatively to grazing in the Great Plains and western shrubsteppe. In North Dakota, short-eared 

owls used upland sites with tall, dense vegetation and 2- to 8-year accumulations of residual vegetation 

(Duebbert and Lokemoen, 1977). Kantrud and Higgins (1992) found short-eared owls nesting in both 

idle native and seeded grasslands in North Dakota. Fields idled during the current growing season had 

more nests than fields under long-term grazing; no nests were found in fields grazed during the breeding 

season (Kantrud and Higgins, 1992). Data from 12 years of nesting studies on annually grazed 

grasslands indicated that short-eared owls (along with American bitterns and marsh hawks) did not use 

grazed habitats for nesting (Kirsch and others, 1978). Johnson and Horn (2008) examined the effects of 

rotational grazing on rodents and raptors in a coastal grassland in northwestern California. They found 

that grazing reduced short-eared owl activity and this response was presumably driven by changes in the 

availability of rodent prey, not by a direct response to the presence of cattle. 

The impact of prescribed burning on ground-nesting birds was investigated in northwestern 

North Dakota (Kruse and Piehl, 1986). Prescribed burns occurred in 1982 and 1984 and 69 percent of 

the active clutches survived the fire and 37 percent eventually hatched. One short-eared owl nest was 

active during the fire, but no new nests were initiated after the fire. Although it is generally not 

considered good management practice to prescribe fires during the breeding season, burning an area 

during the nesting season does not necessarily eliminate total production of ground-nesting birds for the 

year. Kruse and Piehl (1986) suggested that land managers who burn during the nesting season and are 

concerned about ground-nesting birds should consider partial burns as a management tool. Partial burns 

will probably have less impact on total vegetation changes, but should result in higher recruitment of 

birds than would complete burns. 

Recommendations

Management actions for short-eared owls should be focused on protecting large open spaces for 

hunting, and maintenance of nesting cover in grasslands and shrublands. The Colorado Partners in 

Flight Bird Conservation Plan suggested that conservation efforts should focus on landscapes of prairie 

with suitable habitat for short-eared owls and should assure that large areas of this habitat are available 

(Beidleman, 2000). Beidleman (2000) also suggested that the best locations for this approach in 

Colorado would be in North Park and the San Luis Valley (SLV). Habitat management schemes for 

nesting waterfowl and upland gamebirds generally benefit short-eared owls (Beidleman, 2000). 

Populations of the principal prey species (voles) should be maintained at levels compatible with 

economic activities on the land. Habitat restoration programs, such as the Conservation and Wetland 

Reserve Programs, have shown some success in restoring suitable habitat for short-eared owls on 

private land. Such programs not only provide suitable nesting and wintering habitat, but they may also 

help to restore small mammal populations, which are the key resource responsible for the population 

fluctuations of this owl. 

In Massachusetts (Massachusetts Natural Heritage Program), management suggestions have 

included a standardized survey protocol, maintaining large continuous tracks of habitat for owls and 

their prey, monitoring predation and human disturbance, public education, and research. The short-eared 

owl has benefitted indirectly from protection of nesting cover for waterfowl (Larsen, 1987). It has also 

benefitted from reclaimed and replanted strip-mines and dikelands (Tate, 1992). Burning and 

maintenance of grasslands for gallinaceous birds and waterfowl provides nesting and foraging cover for 

this owl (Millsap and others, 1987). 
95

Wiggins (2004) suggested the following management approach for short-eared owls in the 

Rocky Mountain region: 

• Improving existing grassland/wetland quality by limiting livestock grazing to achieve a balanced 

mix of short/medium/tall grass heights. Eliminating or restricting livestock grazing during the 

owl breeding season (April through July) on known or potential breeding plots. 

• Pursuing habitat conservation measures that will increase the amount of native grassland and 

wetland habitats, particularly on private land holdings. 

• Supporting a research program that would monitor the reproductive success of short-eared owls 

within the region, relative to the success of owls nesting on other land management types. 

Dechant and others (2003) made the following management recommendations for the shorteared owl:

• Create and protect large open areas for short-eared owls and their prey (Holt and Leasure 1993). 

Because short-eared owls are nomadic, they may be present only sporadically, but suitable 

habitat should be maintained (Clark 1975, Holt and Leasure 1993). 

• Preserve native grassland. Collaborate with ranching and farming advocates to maintain native 

pasture and rangeland (Johnson 1996). Maintain a mosaic of grasslands and wetlands so that 

some units are available for nesting, while others are being treated to halt succession (Ryan 

1990, Murphy 1993). 

• Protect grasslands through conservation easements, land purchases, and development of farm 

programs with wildlife habitat conservation priorities (Johnson 1996). Continue the 

Conservation Reserve Program to preserve nesting habitat (Sinclair 1990, Murphy 1991, Bock et 

al. 1993). 

• In North Dakota, periodically burn, mow, or graze, to maintain the 2-8 yr old accumulations of 

residual vegetation preferred by short-eared owls in this region (Duebbert and Lokemoen 1977, 

Berkey et al. 1993, Murphy 1993). In tallgrass areas, burning, mowing, or grazing every 2-5 yr 

is recommended to maintain habitat for small mammal prey (Leman and Clausen 1984, 

Kaufman et al. 1990). In Illinois, burn, mow, or use other techniques to reduce grass height and 

maintain vegetation 30-40 cm tall (Herkert et al. 1999). 

• Increase the amount of western public rangeland from which livestock are excluded, especially 

in the U. S. Forest Service National Grasslands (Bock et al. 1993). 

• To prevent mortality or injury from collisions with fences, remove unused fences (Fitzner 1975). 

Increase visibility of fences by hanging pieces of ribbon or foil. 

Research Priorities

One of the areas in the United States where short-eared owls appear to breed regularly is at 

Monte Vista NWR (Wiggins, 2004). Such a breeding population may allow for a study of a number of 

problems that are not possible to study elsewhere, including long-term changes in the number of 

breeding pairs, as well as responses to various management activities (for example, grazing, prescribed 

fire, and water regulation). The apparently higher site fidelity in such areas could be used to study two 

important demographic parameters: survival and philopatry (Wiggins, 2004). Although this species is 

nomadic, the extent to which local habitat characteristics (for example, availability of different 

grassland successional stages) influence site fidelity deserves further research. Holt and Leasure (1993) 

suggested that noncyclic, insular vole populations on offshore Massachusetts islands resulted in greater 

population stability and higher philopatry there. Thus, it is possible that habitat management programs 

(for example, wetland inundation schemes, prescribed fire) carried out on National Wildlife Refuges 
96

may be providing an optimal mosaic of grassland/wetland cover for short-eared owls, which in turn may 

be responsible for the higher frequency of breeding activity in those areas (Wiggins, 2004). Studies that 

involved individually marking males and females could also provide conclusive evidence of the extent 

of site fidelity at breeding and wintering sites. Wiggins (2004) also suggested that detailed, long-term 

studies of site occupancy and owl reproductive success would be useful as well as studies of how habitat 

fragmentation effects occupancy and reproductive success. Wiggins (2004) also suggested that a 

quantitative study of foraging habitat choice, in summer and winter, would provide useful information 

for the conservation of short-eared owls in the Rocky Mountain region. 

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VII. Brewer’s Sparrow (Spizella breweri) 

Introduction

The Brewer’s sparrow (Spizella breweri) is a 

sagebrush obligate that is often the most abundant 

songbird in sagebrush shrubsteppe habitats (Holmes and 

Johnson, 2005). The winter range for the Brewer’s 

sparrow extends from Death Valley to west-central 

Texas, and south through the western Mexico states of

Jalisco and Guanajuato. Their summer range extends

into the three western provinces of Canada. 

Brewer’s sparrow is considered globally 

“secure” by the Natural Heritage Program because of its 

wide distribution across North America. According to 

the Breeding Bird Survey (BBS), however, populations 

have declined by over 50 percent during the past 25 

years. Brewer’s sparrow populations within the states of 

the Rocky Mountain region have exhibited similar long-term declines; in fact, declines in Colorado and 

Nebraska have outpaced national trends (Holmes and Johnson, 2005). In South Dakota and Kansas, the 

species is considered of conservation concern by the states’ Natural Heritage Programs. The Brewer’s 

sparrow is listed as a priority bird species in the Colorado and Wyoming Partners in Flight Bird 

Conservation Plans and as a species of special concern by the USFWS. As a result of these regional 

declines and the species’ vulnerability to habitat loss, the U.S. Forest Service Region 2 listed the species 

as a regional sensitive species. 

A simple Google search on the Internet for “Spizella breweri” revealed 36,200 results and a 

Google Scholar search provided 1,160 results (accessed 11 July 2011). “Brewer’s sparrow” revealed 

105,000 results through Google and 1,060 results through Google Scholar (accessed 11 July 2011). 

Natural History

Morphology & Systematics 

The Brewer’s sparrow is a small, slim, long-tailed sparrow with a thin conical bill and a notched 

tail (Rotenberry and others, 1999). It has a finely streaked brown crown and a median crown-stripe that 

is frequently absent. The breast is unstreaked as an adult although sometimes the flanks are streaked. 

There is little or no seasonal change in plumage. There are two recognized subspecies: S. b. breweri

(Brewer’s sparrow) and S. b. taverneri (timberline sparrow) (Rotenberry and others, 1999). The 

timberline sparrow was originally described as a separate species and is recognized as such by some 

authorities. These subspecies differ in morphology, song, and breeding habitat (Swarth and Brooks, 

1925; Oberholser, 1974; Pyle and Howell, 1996; Doyle, 1997). Breeding habitat for Brewer’s sparrow is 

sage scrub or shrubsteppe and breeding habitat for the timberline sparrow is subalpine-alpine transition 

with low canopy cover (Oberholser, 1974; Doyle, 1997). 

Migration

Brewer’s sparrows are Nearctic-Neotropical migrants. Most individuals breed in and around the 

Great Basin and winter in the Sonoran and Chihuahuan deserts of southwestern United States, western 

Brewer’s Sparrow, courtesy of Dave Menke, U.S. Fish 

and Wildlife Service.
105

Mexico (including Baja California peninsula), and the Mexican Plateau (Rappole and others, 1993). 

Northerly populations move the farthest south and some southerly populations probably move only a 

short distance to wintering grounds (Rotenberry and others, 1999). Primary migratory routes are northsouth oriented through the Intermountain West. Brewer’s sparrows are early spring migrants: first 

arrivals on breeding grounds in the Great Basin are in mid- to late March (Paine, 1968). The bulk of 

migration is in April with some stragglers arriving into late May. Fall migration sometimes begins as 

early as late July (Paine, 1968), but mid-August through October is more typical (Rotenberry and others, 

1999). As with many sparrows, migration is nocturnal (Rotenberry and others, 1999). 

Habitat

Brewer’s sparrow (Spizella breweri) is a sagebrush-steppe obligate species (Sauder, 2002). 

Brewer’s sparrows breed in shrublands and the average canopy height is usually <1.5 m (Rotenberry 

and others, 1999). Throughout most of its range in the United States, it is most closely associated with 

landscapes dominated by big sagebrush (Artemisia tridentata), where it is usually the most abundant 

bird species (Wiens and Rotenberry, 1981a). The Brewer’s sparrow overlaps extensively with the sage 

sparrow (Amphispiza belli). Brewer’s sparrows may also occur in large openings in pinyon-juniper 

(Pinus edulis – Juniperus spp.) woodlands (Sedgwick, 1987) or large parklands within coniferous 

forests. In northwestern Canada, this sparrow occurs at high elevation above timberline and in shrubby 

montane valleys dominated by low-growing willow (Salix spp.), dwarf birch (Betula spp.), and shrubs 

(Doyle, 1997). Considerable research has been conducted on habitat associations, habitat selection, and 

density-habitat correlations during the breeding season (for example, Best, 1972; Rotenberry and Wiens, 

1980; Wiens and Rotenberry, 1981a; 1985; Larson and Bock, 1986; Rotenberry, 1986; Wiens and 

others, 1986; 1987a; Petersen and Best, 1987; Knopf and others, 1990; Knick and Rotenberry, 1995). 

Sauder (2002) studied factors influencing avian presence and abundance in sagebrush-steppe 

communities of southeast Idaho. He found that Brewer’s sparrow density was significantly correlated 

with total shrub cover, and capture rates were significantly correlated with shrub cover and dead wood 

cover. Brewer’s sparrows were positively influenced by factors indicating more structurally complex 

habitat. 

Seventy-five percent of the Breeding Bird Atlas confirmations for Colorado occurred in 

sagebrush (Lambeth, 1998). Ligon (1961) considered Brewer’s sparrows to be associated with 

rabbitbrush and big sage. They have also been found nesting in willow in Colorado (Lambeth, 1998). 

Habitat characteristics correlated with dense populations of this species include a dominance of shrubs, 

but little diversity in shrubs, more forbs, less grass, more bare ground, less duff, and relatively level 

ground with little rockiness (Short, 1984). Breeding Bird Atlas locations for Brewer’s sparrow indicate 

preference for sagebrush of middle heights. Brewer’s sparrows usually take the foliage-feeding position, 

and sage and vesper sparrows, their close associates, forage mostly on the ground (Wiens and others, 

1990). After fledging, Brewer’s sparrow families often move to higher elevations, mixing with other 

species, especially chipping sparrows. 

Foraging Habits

The main foods taken by Brewer’s sparrows are small insects, mainly gleaned from foliage and 

bark of shrubs and dwarf trees (Rotenberry and others, 1999). They also eat seeds, which are usually 

taken from the ground. Brewer’s sparrows forage mostly in shrubs (>75 percent of over 600 observation 

periods) and less commonly on open ground between shrubs or at the base of bunchgrasses (Wiens and 

others, 1987b). Shrubs selected by foraging sparrows differ significantly from those randomly available; 

they are larger, more vigorous, and more likely to consist of sagebrush rather than green 
106

(Chrysothamnus viscidiflora) or gray (C. nauseosus) rabbitbrush (Rotenberry and Wiens, 1998). They 

intersperse singing with foraging within bouts of foraging and songs are delivered from widely spaced 

perches (Wiens and others, 1990). 

Breeding

Brewer’s sparrows start to arrive in Colorado in mid-April, with full numbers present at the end 

of the month (Lambeth, 1998). They sing during migration and on the nesting ground, but once paired 

the birds become quieter (Best and Petersen, 1985). This variable, plus annual fluctuations in nesting 

schedules due to weather, complicates population monitoring of this species. Increased singing from 

delayed pairings or failed nesting might falsely indicate relatively high populations. In wetter years, they 

have larger clutch sizes. Clutch size is typically 3-5 eggs (Harrison, 1979). Brown-headed cowbirds 

sometimes parasitize this species although there are no records of this in Colorado (Lambeth, 1998). 

Brewer’s sparrows produce replacement clutches and frequently double broods (Rotenberry and others, 

1999). Renesting begins soon after the loss of the first nest. Second nests are initiated approximately 10 

days after the first brood fledges. 

In southeastern Idaho, about 25 percent of color-banded males returned to nesting habitat used in 

the previous year (Petersen and Best, 1987). Males and females forage together during the first week 

after pairing; then the females become more secretive during the beginning of the nesting process 

(Nordin and others, 1988). Compared with surrounding habitat, nests are located in significantly taller, 

denser shrubs (primarily big sagebrush) with reduced bare ground and herbaceous cover (Petersen and 

Best, 1985). Available cover may be the primary factor in nest-site selection (Best, 1972). Nests are 

located toward the edge of the densest portion of sagebrush relative to nests of the sage sparrow and the 

sage thrasher (Rich, 1980). Brewer’s sparrows prefer nest shrubs to be entirely alive or mostly alive 

(Petersen and Best, 1985). The foliage of live shrubs provides concealment from predators and 

protection from the elements. Occasionally Brewer’s sparrows nest in herbicide-killed shrubs with no 

foliage but with dense branching and plenty of grass cover (Best, 1972). 

Demography & Populations

There was significant annual variation in the average number of eggs per clutch at a site in 

Oregon during a 5-year period (1976-1980; Rotenberry and Wiens, 1989). The increase in clutch size 

was strongly correlated with increasing precipitation during the preceding winter, implying that 

Brewer’s sparrows respond opportunistically by increasing initial reproductive investment when food is 

more plentiful (Rotenberry and Wiens, 1991). However, a study in Idaho from 1980 to 1985 found no 

significant annual variation in clutch size (Petersen and Best, 1987). Proportion of males that 

successfully acquire mates varies among years and sites (Rotenberry and others, 1999). In Idaho from 

1976-1977, only 23 percent of males were successful (Reynolds, 1981). At the same site in 1980-1985, 

86 percent were successful (Petersen and Best, 1987). Number of fledglings produced per nest also 

varies annually and geographically. 

In southeastern Idaho, the proportion of banded males returning to the same breeding area from 

one year to the next was 25 percent (Petersen and Best, 1987). Return rates were comparable for this 

species in Oregon (Rotenberry and others, 1999). However, these rates represent a lower limit, since 

they do not account for individuals alive but breeding elsewhere. 

Population numbers for this species are highly variable depending on habitat conditions and year 

(Rotenberry and others, 1999). Although the species is most common in sagebrush-dominated habitats, 

local densities are not significantly correlated with variation in percent cover of sagebrush. However, 

densities were observed to decline on plots on which coverage of sagebrush had been reduced, either 
107

through experimental manipulation or by wildfire (Best, 1972; Schroeder and Sturges, 1975; Castrale, 

1982; Kerley and Anderson, 1995). Brewer’s sparrow density is usually correlated with some aspect of 

total shrub cover, or “shrubbiness” (Rotenberry and Wiens, 1980; Wiens and Rotenberry, 1981a; 

Dobler, 1994; Knick and Rotenberry, 1995). 

Population numbers of Brewer’s sparrows are negatively influenced by increasing landscapelevel fragmentation of shrublands, and those numbers appear to be more sensitive to variation in 

landscape-level attributes than in local-scale attributes (Knick and Rotenberry, 1995). 

BBS data show the highest abundance for Brewer’s sparrows (30-40 individuals per route) in 

central Nevada northward to southeastern Oregon, with other centers of abundance in southeastern 

Idaho and southwestern Wyoming (Sauer and others, 1997). The bulk of the breeding population 

appears to be mostly contiguously distributed. Winter counts show the highest abundance in the United 

States in southern Arizona, New Mexico, and western Texas (Sauer and others, 1997). There is no 

information about distribution and abundance in Mexico (Rotenberry and others, 1999) and no density 

estimates reported for wintering birds. 

Brewer’s sparrows appear to have undergone statistically significant (p < 0.05) declines (from 

about -3 to -6 percent per year) throughout the BBS survey area during 1966-1996 (Sauer and others, 

1997). Surveywide, decline has been -3.7 percent per year. Individual states with significant declines are 

California (-3.3 percent per year), Idaho (-6.0 percent per year), and Wyoming (-2.8 percent per year). 

Declines were more pronounced from 1980-1996 than in 1966-1979. No state or physiographic region 

shows significant upward trends. In contrast, surveywide trends in winter abundance are positive, but 

not significant (0.2 percent per year). From 1959-1988, there was a significant increase noted in Texas 

(6.7 percent per year; Sauer and others, 1997). 

Although Brewer’s sparrows are considered common in western Colorado (Andrews and 

Righter, 1992), trend estimates show significant decreases from 1966 to 2002 (Sauer and others, 2004). 

Declines were most pronounced between 1966 and 1979. An analysis of detection frequencies on BBS 

routes for the periods 1968-1983 and 1983-2001 (Dobkin and Sauder, 2004), indicated detection 

frequencies on routes in western, southern, and eastern Colorado declined, while in north-central 

Colorado they increased. The BBS regional trend analysis for Brewer’s sparrows along 96 Colorado 

routes from 1966 to 1999 showed a negative trend estimate of -1.9 (-3.8, 0.4; 95 percent CI) (Sauer and 

others, 2011). From 1999 to 2009, the trend was also negative and estimated to be -2.0 (-5.2, 1.4; 95 

percent CI). The regional credibility measure for this species was the category “blue.” A blue category 

reflects data with at least 14 samples in the long term, of moderate precision, and of moderate 

abundance on routes. 

Threats

Sources of Mortality

There are few disease threats reported for Brewer’s sparrows. Of four adults examined, two were 

infected with the blood protozoan Haemoproteus and one with Haemoproteus and Trypanosoma

(Greiner and others, 1975). Brewer’s sparrows are host to several ectoparasites including biting lice, 

flesh flies, and blowflies (Rotenberry and others, 1999). Predation of adult Brewer’s sparrows by 

loggerhead shrikes can be a significant source of mortality (Reynolds, 1979; Woods, 1994). Predation 

by other avian predators (for example, American kestrels, sharp-shinned hawks, and Cooper’s hawks) is 

also a potentially significant cause of mortality. The gopher snake (Pituophis melanoleucus) and the 

Townsend’s ground squirrel (Spermophilus townsendii) are documented predators of eggs and nestling 

Brewer’s sparrows (Holmes and Johnson, 2005). In Oregon, intense, episodic predation of nests was 
108

attributed to a marked increase in the density of Townsend’s ground squirrels. The irruption of ground 

squirrels appeared to be triggered by a pattern of drought year followed by two relatively wet years, and 

nest predation peaked when ground squirrels were most abundant (Rotenberry and Wiens, 1989; 

Rotenberry and others, 1999). Other potential nest predators include the common raven, black-billed 

magpie, longtailed weasel (Mustela frenata), least chipmunk (Tamius minimus), and snakes other than 

gopher snakes, such as the western rattlesnake (Crotalus viridis) (Reynolds, 1979; Petersen and Best, 

1987; Rotenberry and Wiens, 1989; Rotenberry and others, 1999). These nest predators likely prey on 

adults as well (Rotenberry and others, 1999). In addition, a coachwhip snake (Masticophis flagellum) 

has been reported eating adults (Paine, 1968; Rotenberry and others, 1999). 

Brewer’s sparrow nests are hosts to parasitic brown-headed cowbirds (Molothrus ater) (Holmes 

and Johnson, 2005). Parasitized nests are usually abandoned by the host during incubation; this 

desertion usually results in the loss of the entire clutch (Rotenberry and others, 1999). Also, because 

cowbirds usually remove at least one host egg (Rich, 1978), the number of host young is typically 

reduced (Rotenberry and others, 1999). The frequency of brood parasitism varies geographically. In 

southern Idaho, 13 percent of nests were parasitized (Rich, 1978); in southeast Washington, 5 percent 

(Vander Haegen and Walker, 1999); in central Oregon, 0 percent (Rotenberry and Wiens, 1989); and in 

southeast Alberta, 52 percent (Biermann and others, 1987). There is no information on the impact of 

brood parasitism on host productivity at the population level (Rotenberry and others, 1999). 

Habitat Loss & Fragmentation

Degradation of habitat is the most important threat to Brewer’s sparrows (Rotenberry and others, 

1999). Large-scale fragmentation and reduction in area of native shrublands has been occurring 

throughout the Intermountain West (Braun and others, 1976), and may be responsible for declines in 

Brewer’s sparrows reflected in BBS data (Rotenberry, 1998). A conservative estimate suggested that 

through the mid-1970s, 10 percent of native sagebrush shrublands had been converted to agriculture 

(Braun and others, 1976). In Colorado, 30 percent of sagebrush was altered between 1900 and 1974 

(Braun and others, 1976). In addition to agriculture, a number of other activities have contributed to 

habitat loss and fragmentation, including urban and suburban development, development of road and 

powerline rights of way, and range improvement programs that remove sagebrush by burning, herbicide 

application, and replacing sagebrush with annual grassland to promote forage for livestock (Holmes and 

Johnson, 2005). The pace of loss has accelerated because of the complex interactions among agriculture, 

livestock grazing, and invasion of exotic annual plants, especially cheatgrass (Bromus tectorum; 

Rotenberry, 1998). Brewer’s sparrows belong to a group of species that are most typical of undisturbed 

shrubsteppe; as such, they appear to be especially sensitive to the negative effects of habitat 

fragmentation (Knick and Rotenberry, 1995a; Rotenberry, 1998). Fragmentation may also reduce 

sagebrush patch size below a minimum suitable for Brewer’s sparrow occupancy (Knick and 

Rotenberry, 1995a). 

Nest predation is probably the principal determinant of reproductive success (Rotenberry and 

others, 1999), and predation rates are apparently affected by fragmentation of sagebrush shrubsteppe 

habitat. Predation on artificial and natural nests of shrubsteppe birds was higher in fragmented sites than 

in continuous shrubsteppe sites in eastern Washington, likely due to an increase in corvid populations 

associated with agricultural and other human-modified habitats (Holmes and Johnson, 2005). Also, 

nests might be more difficult to locate in extensive stands of shrubsteppe than in fragmented sites 

(Vander Haegen and others, 2002). 
109

Agriculture

The predominant impact of agricultural development is direct habitat loss due to conversion of 

shrubland areas to grasslands and croplands, and the subsequent fragmentation of once contiguous 

shrublands. These losses are long-term and include entire landscapes that are unlikely to be returned to 

shrublands in the foreseeable future (Dobler and others, 1996; Vander Haegen and others, 2000; Knick 

and Rotenberry, 2002). In addition to outright habitat loss, agricultural development has a less obvious 

but destructive role in the introduction and spread of alien plants into natural habitats; this may now be 

the most serious threat to these habitats in many areas (Rotenberry, 1998). Agricultural areas, and their 

associated roads, serve as continually renewable sources for immigrant alien plant species (Janzen, 

1986; Alberts and others, 1993; Rotenberry, 1998). Furthermore, agricultural areas apparently extend 

the landscape-level distribution of brown-headed cowbirds, which are avian brood parasites, and corvids 

such as common ravens and American crows, which can be major predators of songbird nests (Marzluff 

and others, 1994; Robinson and others, 1995; Rotenberry, 1998; Vander Haegen and others, 2000). 

Invasion of Exotic Annual Grasses

Agricultural development, livestock grazing, off-road vehicle use, and road building disturb soil, 

which promotes the germination of annual plant seeds and promotes the invasion of exotic annual plants 

into otherwise undisturbed areas. This process has resulted in perhaps the greatest impact on western 

shrublands: the establishment of the invasive, exotic cheatgrass (Holmes and Johnson, 2005). 

Cheatgrass first appeared in the early to mid-1980s, probably as a contaminant in grain seed (Mack, 

1981). It then spread into undisturbed shrublands through widespread livestock grazing and the soil 

surface disturbance. Cheatgrass now occupies millions of hectares of western rangelands, has greatly 

increased fire frequency, and has substantially, and perhaps permanently, altered postfire successional 

pathways (Whisenant, 1990; Rotenberry and others, 1999). The principal impact has been to alter the 

fire ecology of the shrubsteppe ecosystems. Cheatgrass provides a continuous surface cover of relatively 

fine fuel that carries fire into and over much larger areas than likely occurred historically (Shienant, 

1990). It matures and dries earlier than native bunchgrass, increasing the chance of fire earlier in the 

season (Knick and Rotenberry, 1997). Also, because cheatgrass does not catch and hold snow like a 

diverse perennial stand of vegetation, the site becomes drier. Shrublands infested with cheatgrass are 20 

times more likely to burn than lands without cheatgrass (Stewart and Hull, 1949, Whisenant, 1990). 

Although cheatgrass and the increased fire frequency associated with its invasion are threats to 

sagebrush habitats and the species dependent on them, complete fire suppression also can be a threat to 

healthy sagebrush ecosystems (Holmes and Johnson, 2005). In some areas, fire suppression and the loss 

of fine fuels to livestock grazing have resulted in much longer fire-return intervals, which has altered the 

dominant process (for example, fire) that controlled the shifting temporal and spatial mosaic of 

grasslands and shrubland characteristic of these landscapes (Dobkin and Sauder, 2004). In Colorado, the 

ecological integrity of sagebrush shrublands has been widely compromised by the invasion of exotic 

cheatgrass or native pinyon and juniper plant species (Biedleman, 2000). 

Prescribed Fire

Burning over large areas to eradicate sagebrush is detrimental to Brewer’s sparrows because it 

removes shrub cover, fragments large tracts of sagebrush, and can reduce patch size to levels not used 

by Brewer’s sparrows (Holmes and Johnson, 2005). It also promotes changes in the vegetative 

community. Although there is some disagreement over the frequency and spatial scale of fires prior to 

Euro-American settlement, there is uniform agreement that fire frequencies in the Intermountain West 
110

have been greatly altered over the past 150 years (Dobkin and Sauder, 2004). Altered fire frequencies in 

combination with the ubiquity of livestock grazing continue to drive the loss of native plant community 

structure and composition upon which shrubsteppe birds depend. 

Mining and Oil/Gas Development

Energy development and natural resource extraction directly alter sagebrush habitats at the site 

of operation (Braun and others, 2002; Knick and others, 2003). Associated road networks, pipelines, and 

power transmission corridors fragment habitat and create soil conditions facilitating the spread of 

invasive species (Braun, 1998; Gelbard and Belnap, 2003; Knick and others, 2003). The cumulative 

effects of energy development have not been assessed. However, the density of sagebrush-obligate birds 

within 100 m of roads constructed for natural gas development was 50 percent lower than at greater 

distances (Ingelfinger, 2001; Knick and others, 2003). In addition, the increased numbers of corvids and 

raptors associated with powerlines also increase the potential impact of predation on sagebrush-breeding 

birds (Knick and others, 2003). Within the Rocky Mountain region, increasing rates and densities of oil 

and gas development constitute expanding threats to sagebrush ecosystems. In Wyoming, oil and gas

wells have been located primarily in habitats dominated by sagebrush. 

Livestock Grazing

Livestock grazing has affected most of the sagebrush habitats across the range of the Brewer’s 

sparrow. The effects of livestock grazing in shrubland habitats are complex, depending on grazing 

intensity, season, and duration and the extent of alteration to native vegetation. Livestock grazing 

impacts sagebrush habitats in several interrelated ways (Holmes and Johnson, 2005). Its greatest impact 

is soil disturbance that promotes the germination of annual plant seeds and promotes the invasion of 

exotic annual plants, as previously mentioned. In eastern Washington, Brewer’s sparrow abundance was 

significantly lower at sites with poor range condition (less than 25 percent cover in climax vegetation) 

than at sites with fair condition (25 to 50 percent cover in climax vegetation), but abundance did not 

differ between fair and good sites (greater than 50 percent cover in climax vegetation) (Holmes and 

Johnson, 2005). 

Management

Response to Management

The influence of management activities and disturbances of Brewer’s sparrows habitat have led, 

in many areas of the West, to large-scale conversion of shrublands to grassland habitats dominated by 

exotic annuals (Holmes and Johnson, 2005). The conversion of landscapes to exotic annual grasslands 

with high fire frequencies results in changes in the avian composition from communities composed of 

mostly shrubland obligates (such as the Brewer’s sparrow) to those composed of grassland species (such 

as meadowlarks [Sturnella spp.] and horned larks [Eremophila alpestris]) (Knick and Rotenberry, 

1997). Sagebrush habitat has been found to support more species, have a greater bird density and a 

greater individual density for most species than cheatgrass habitat types (Schuler and others, 1993). This 

continuing conversion process will likely result in loss of bird species richness and decreased numbers 

of shrubland-obligate species, including the Brewer’s sparrow (Knick and Rotenberry, 2002). 

Holmes (2007) measured songbird abundance and vegetation cover in and around a 420-ha 

prescribed burn in a mountain big sagebrush community in Oregon. Brewer’s sparrows occurred in 

lower abundance on sites that were largely or completely burned relative to sites that were outside the 
111

fire perimeter or within unburned islands of shrubs. The number of Brewer’s sparrow detections was 

linearly related to remaining sagebrush cover. Reported responses of Brewer’s sparrows to fire are 

consistently negative (Castrale, 1982; Bock and Bock, 1987; Knick and Rotenberry, 1999; Petersen and 

Best, 1999; Reinkensmeyer, 2000; McIntyre, 2002; Welch, 2002). 

Recommendations

The overriding essential element for the conservation of Brewer’s sparrow is healthy sagebrush 

shrubsteppe. Nonnative grasses and agricultural and urban conversion now dominate most western 

shrublands, making it especially important to sustain the remaining native sagebrush communities in a 

healthy state to support native wildlife, including Brewer’s sparrows populations (Paige and Ritter, 

1999; Rotenberry and others, 1999). At a regional scale, the maintenance of Brewer’s sparrows depends 

on the existence of extensive tracts of sagebrush shrublands and associated habitat physiognomy, while 

on a more local scale, the occurrence and abundance of Brewer’s sparrows depend on high sagebrush 

cover, large patch size, habitat heterogeneity, low disturbance, and little fragmentation. The minimum 

patch size and degree of patch isolation for breeding has not been determined. However, Brewer’s 

sparrows do appear to be area-sensitive, and isolated stands of sagebrush smaller than 2 ha are not likely 

to be used as nesting habitat (Knick and Rotenberry, 1995a). 

Management Approaches

Holmes and Johnson (2005) suggested the following management approaches:

• The maintenance of the largest, most continuous stands of sagebrush that exist within the SLV 

complex of Refuges will benefit Brewer’s sparrows, as will the maintenance of small openings 

(for example, <1 ha) of short vegetation surrounded by sagebrush (Paige and Ritter, 1999). The 

first step would be to identify and protect areas of healthy sagebrush shrubsteppe (Holmes and 

Johnson, 2005). Practices that permanently convert shrublands to nonnative grasslands or 

agricultural lands should be avoided, as should practices that further fragment areas, such as 

road-building. The creation of edges with converted habitat should be minimized. Burning to 

eradicate sagebrush over large areas (for example, >20 to 40 ha) should be avoided, and range 

fires that threaten to eradicate large areas of sagebrush should be suppressed. Providing for “no 

net loss of sagebrush habitats” (Paige and Ritter, 1999) across landscapes will help to achieve 

the desired condition of large areas of healthy sagebrush in a landscape that mimics a natural 

matrix. 

• Disturbed sites should be restored or rehabilitated to native plant communities, if possible. 

Management should move toward the re-establishment of natural disturbance processes and the 

careful management of practices, such as livestock grazing, which often degrades sagebrush 

steppe. In some areas, simply protecting areas from degradation may allow for the reestablishment of native plants. In large undisturbed areas, however, sagebrush and native 

perennial grasses may need to be reseeded to shorten the recovery time and to prevent 

dominance by nonnative grasses and forbs. Ecologically appropriate native plant species should 

be used in all shrubsteppe restoration projects (Knick and others, 2003). 

• Sources of soil disturbance (for example, road building, off-road vehicle use, heavy grazing) 

should be minimized to maintain biological soil crusts, which are sensitive to trampling by 

humans, vehicles, and livestock (Paige and Ritter, 1999). Roads that fragment otherwise 

continuous sagebrush should be closed or re-routed. All vehicles should be kept on established 

roads and trails or confined to areas established specifically for off-road recreation. These areas 
112

should be established so as not to influence sensitive sagebrush habitats and species. 

No single grazing strategy is appropriate for all shrubland habitats, and grazing management 

should be tailored to the condition and potential of each grazing unit (Paige and Ritter, 1999). 

Management plans should consider other grazers such as elk and deer, and their influence on the 

vegetation. Stocking levels should enable the stabilization or increase of native perennial groundcover 

and reduce disturbance to soil crusts. In addition, grazing seasons and the distribution and types of 

grazers should be managed to promote native groundcover and to maintain herbaceous cover for nest 

concealment (Paige and Ritter, 1999). For instance, in stands where cheatgrass and native perennial 

grasses are mixed, grazing during the dormant period may favor perennial species (Vallentine and 

Stevens, 1994; Young, 1994; Paige and Ritter, 1999). Managing grazing time and intensity can reduce 

or eliminate trampling of ground nests and nests in low shrubs, such as Brewer’s sparrow nests. 

Practices and activities that concentrate livestock, such as corrals, feedlots, and watering sites, 

during the Brewer’s sparrow breeding season may increase the impact of brown-headed cowbird brood 

parasitism on Brewer’s sparrow nesting success (Robinson and others, 1995). Rotating livestock to rest 

units from cowbird concentrations in alternate years may give local songbird populations breeding 

opportunities without high parasitism pressure (Paige and Ritter, 1999). 

Research Priorities and Information Needs

Birds can be good indicators of ecosystem health because they tend to have dynamics that 

parallel those of the ecosystem; they are sensitive enough to provide an early warning of change and to 

provide continuous assessment over a wide range of stresses. Birds also have dynamics that can be 

linked to either natural cycles or anthropogenic stressors. In addition, bird populations are distributed 

over wide geographical areas and/or are often relatively numerous, so they can be accurately estimated, 

have costs of measurement that are not cost prohibitive, have a low impact to measure, and can provide 

measurable results that are repeatable with different personnel. Because of the relative ease of detection 

of songbirds such as the Brewer’s sparrow, both individual species and communities can be monitored. 

Thus, birds can serve as informative organisms for measuring natural and anthropogenic changes and 

for guiding and measuring management and restoration actions (Greenwood and others, 1993; Hutto, 

1998; Hutto and Young, 2002). 

Prior to monitoring, an inventory of the biological resource of concern is generally conducted. 

Bird inventories seek to identify species presence/absence, range, distribution, and relative abundance, 

and they can provide a baseline for monitoring. Bird monitoring programs are generally developed on 

this baseline inventory. Commonly, monitoring programs fit into one of two categories: those designed 

to monitor long-term population trends and those designed to monitor the effects of land management 

actions. The goals of the monitoring program will affect the methods used. The current Partners in 

Flight landbird monitoring strategy (Bart and Ralph, 2001) recommends integration of long-term 

monitoring programs and short-term assessments of habitat associations and land-use effects (Hutto and 

Young, 2002). 

Designing a monitoring program that includes habitat information with sample locations will 

facilitate the discovery of habitat associations and make a much more effective monitoring program than 

one based on long-term population trends alone (Hutto and Young, 2002). An integrated approach to 

monitoring both vital rates and population trends of Brewer’s sparrows, and relating them to habitat 

characteristics across the landscape, is critical for understanding causes of population changes and for 

identifying, as well as testing, management actions and conservation strategies to reverse population 

declines (Fancy and Sauer, 2000). Therefore, habitat inventory and habitat monitoring should be 
113

conducted concomitant with a Brewer’s sparrow inventory and monitoring program. Vegetation and 

habitat should be characterized at multiple spatial scales, and measurements of both horizontal and 

vertical structure should be taken in areas where sparrows are detected and where they are not. Habitat 

measurements can be coupled with bird inventories to establish species habitat selection within the area 

of concern. This information can be used as baseline information in subsequent monitoring of long-term 

avian population trends, and the effects of land use and management actions. 

A review written by experts (Knick and others, 2003) discussed the threats that birds in 

sagebrush habitats face, synthesized existing information regarding avifauna of sagebrush habitats, and 

identified their management and conservation needs. They concluded that little is understood regarding 

shrubsteppe bird-habitat dynamics. To manage and conserve birds in sagebrush habitats there is an 

urgent need for more research on four major topics: 

• Identification of primary land-use practices and their influence on sagebrush habitats and birds, 

• Understanding of bird responses to habitat components and disturbance processes of sagebrush 

ecosystems, 

• Hierarchical designs for surveying and monitoring programs, and 

• Linking bird movements and population changes during migration and wintering periods to 

dynamics on sagebrush breeding areas.

This review found that we do not have a comprehensive understanding of how habitat 

fragmentation influences the productivity, density of breeding adults, size of home range, probability of 

predation, or brown-headed cowbird parasitism on Brewer’s sparrows. The general breeding distribution 

of the Brewer’s sparrow is relatively well known. However, within its geographic distribution, 

distribution patterns, habitat associations, and demographic information within a range of habitat 

conditions are not well known. Without information on habitat associations, reproductive success, adult 

and juvenile survivorship, and dispersal patterns, it is impossible to understand and predict the effects of 

different management options and conservation actions on source-sink dynamics (Knick and others, 

2003). The Brewer’s sparrow appears to be an area-dependent species, and studies that improve our 

knowledge of how landscape context influences Brewer’s sparrows’ sensitivity to habitat fragmentation, 

nest predation, and parasitism rates are also needed. Such studies will provide information that can 

guide conservation planners in determining adequate sizes of sagebrush conservation areas, how these

areas should be spatially arranged, and into what type of landscapes they should be placed (Herkert and 

Knopf, 1998). 

In order to study the effects of management approaches on Brewer’s sparrows , such as grazing, 

experiments having strong statistical designs that include treatments and controls at spatial and temporal 

scales relevant to understanding the impacts to vegetation, soils, and the dynamics of recovery are 

needed (Knick and others, 2003). Long-term studies incorporating a widespread system of exclosures 

and the ability to control treatment levels are necessary to measure the effects of land use on habitats 

and birds. Knick and others (2003) suggest that treatment projects planned by management agencies, 

and the large number of areas to be treated, represent a tremendous opportunity to design a sound 

experimental approach. Establishing such studies and combining them with monitoring at appropriate 

scales would provide the feedback to evaluate treatment effects and provide the basis for adaptive 

management strategies (Walters, 1986; Morrison, 2002; Knick and others, 2003). Such information is 

critical for identifying causes of population changes and for testing management actions and 

conservation strategies to reverse population declines. 
114

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VIII. Savannah Sparrow 

(Passerculus sandwichensis) 

Introduction

The savannah sparrow (Passerculus

sandwichensis) is widespread and abundant in 

open habitats throughout North America. Like 

other grassland birds, it is often overlooked 

because of its secretive habits. During the breeding 

season its persistent buzzy song can be heard in 

agricultural fields, meadows, marshes, coastal 

grasslands, and tundra. Most populations are 

migratory, although salt marsh populations in the 

southwest are mainly resident. Adults tend to 

reoccupy the same breeding site. Such philopatry 

has resulted in reproductive isolation and substantial geographic variation; 17 subspecies are currently 

recognized (Wheelwright and Rising, 2008). The species was named for the town of Savannah, Georgia, 

where the type specimen was collected. Savannah sparrows have been used as model organisms for the 

study of avian migratory orientation, with major contributions in the scientific literature by Able and 

coworkers (Able, 1991; Able and Able, 1990; 1996). Several biological features of this sparrow have 

made it attractive for studying comparative ecology and evolution, including the species’ broad 

geographic distribution, occurrence in a variety of habitat types, high breeding densities, natal 

philopatry on islands, and evolutionary divergence of populations. 

Savannah sparrows breed across the continent from the Arctic Ocean south to the Great Basin, 

Rockies, and Appalachians. In Colorado, historical accounts refer to savannah sparrows as common 

summer residents. Although they breed throughout Colorado, savannah sparrows concentrate in wet 

streamside meadows and irrigated hayfields in the mountainous regions of the state. North, Middle, and 

South parks, the upper Yampa Valley, the upper Gunnison Valley, and the San Luis Valley (SLV) 

contain the bulk of the Colorado’s Breeding Bird Atlas confirmations and the largest aggregations of 

records (Toolen, 1998). Populations increased significantly in western North America from 1969 to 

1989 (BBS unpublished data in Wheelwright and Rising, 1993) while declining slightly across the 

continent from 1966-1996. Forest clearing and more intensive agriculture may explain increases in the 

West. Irrigated hay meadows are an important component of savannah sparrow habitats in mountain 

parks and western valleys in Colorado. Prior to settlement these areas were likely already wet grassy 

meadows, so the loss of irrigated hay meadows to urban and resort development for uses such as golf 

courses, residences, and ski area base facilities may negatively affect savannah sparrow populations in 

the future. 

A simple Google search on the Internet for “Passerculus sandwichensis” revealed 115,000 

results and a Google Scholar search provided 3,680 results (accessed 11 July 2011). “Savannah 

sparrow” revealed 243,000 results through Google and 4,970 results through Google Scholar (accessed 

11 July 2011). 

Savannah Sparrow, courtesy of Eleanor Briccetti.
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Natural History

Morphology & Systematics

Savannah sparrows are small and sexually monomorphic in plumage, but these two traits can 

vary geographically (Wheelwright and Rising, 2008). In basic plumage, the nape, back, and rump are 

brownish or grayish, and streaked to varying degrees. Throughout much of the species’ range, the crown 

usually has a pale beige median stripe. The eyebrow stripe is yellow, especially anteriorly, and a little 

yellow is generally concealed in the bend of the wing. The throat, breast, and belly are whitish to pale 

beige; the breast is generally streaked with brown, often with a small central spot; and the throat and 

belly can be with or without streaks (Wheelwright and Rising, 2008). 

On average, savannah sparrows from eastern North America are darker and larger than those 

from the Great Plains or West, with many populations blending into each other morphologically, 

particularly east of the Rocky Mountains (Wheelwright and Rising, 2008). Some 28 different subspecies 

have been named, but just 17 of these are diagnostically distinct (Wheelwright and Rising, 2008). The 

subspecies in Colorado is P. s. nevadensis (Wheelwright and Rising, 2008). This subspecies breeds 

from eastern British Columbia east through Manitoba and south to eastern California east across the 

Great Basin and northern Great Plains. 

Migration

All populations of this species migrate except for those resident along the coast of California, 

and those that breed in Mexico (Wheelwright and Rising, 2008). There is little direct information on 

migration for savannah sparrows; however, birds that breed on the Mexican Plateau, Arizona, and New 

Mexico probably move south or to lower elevations in winter. Before 1987, only four migratory 

individuals banded on breeding grounds had been recovered in winter, and only one banded in winter 

had been recovered on the breeding grounds (Bird Banding Lab; Wheelwright and Rising, 2008). 

In spring, savannah sparrows arrive by late February in Kentucky (Mengel, 1965), early March 

in Missouri (Robbins and Easterla, 1992), early April in South Dakota (South Dakota Ornithologists’ 

Union, 1991), late March in southern Ontario, early May in northern Ontario (Speirs, 1985), mid-April 

in Nova Scotia (Tufts, 1986), late March in Oregon (Gabrielson and Jewett, 1940), and early May in 

Alaska (Gabrielson and Lincoln, 1959). In fall, savannah sparrows depart Alaska in mid- to late 

September (Gabrielson and Lincoln, 1959), South Dakota by early November, northern Ontario by midOctober (Speirs, 1985), New Brunswick by mid-September, and Nova Scotia by late October (Stobo 

and McLaren, 1975). Migrants appear in Kentucky from early September through mid-November 

(Mengel, 1965), in Georgia by early September (Georgia Ornithological Society, 1986), in Florida by 

late September (Kale and Maehr, 1990), in Missouri from mid-September through November (Robbins 

and Eastera, 1992), in Arkansas by mid-September (James and Neal, 1986), in Oregon from late 

September through early November (Gabrielson and Jewett, 1940), and in southern California by late 

August (Unitt, 1984). 

Habitat

Breeding range habitat for savannah sparrows is open country, including grassy meadows, 

cultivated fields (especially alfalfa), lightly grazed pastures, roadsides, coastal grasslands, sedge bogs, 

edge of salt marshes, and tundra (Wheelwright and Rising, 2008). Savannah sparrows avoid areas with 

extensive tree cover, and generally occupy fields with some herbaceous plants or weeds. In arid parts of 

their range, they are generally restricted to irrigated areas or to the grassy margins of ponds. Savannah 
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sparrows prefer dense ground vegetation, especially grasses, and moist microhabitats are favored 

(Wiens, 1969). In Colorado, moist, grassy, mountain meadows are favored breeding habitats, but 

irrigated hay meadows and alfalfa are also important breeding areas throughout Colorado (Toolen, 

1998). Colorado’s Breeding Bird Atlas data confirmed these habitat preferences for Colorado, with 

croplands, montane grassland, and emergent marshes accounting for 58 percent of habitats reported. Salt 

meadows, mountain sagebrush, montane carr, and willow carr made up another 22 percent of sightings. 

Winter habitats for savannah sparrows are cultivated fields, pastures, golf courses, roadsides, dumps, 

dune grass, and salt marshes (Wheelwright and Rising, 2008). 

A 4-year study in tallgrass prairie in Minnesota and North Dakota found it difficult to define 

“optimal” habitat for savannah sparrows. Breeding density varied among regions and years and was 

influenced by landscape effects independent of the size of the grassland patches; the influence of patch 

size and landscape composition on nesting success was unclear (Winter and others, 2006). In North 

Dakota mixed-grass prairie habitats of various successional stages, savannah sparrows are ubiquitous 

and generalized in habitat preferences (Madden and others, 2000). In Saskatchewan, the presence or 

absence of breeding savannah sparrows in different-sized tallgrass pastures varied among years, 

suggesting the species was relatively area-insensitive (Davis, 2004). Field size also proved to be a poor 

predictor of breeding density in several north-central states in the United States (Johnson and Igl, 2001; 

Ribic and Sample, 2001). In contrast, area-incidence functions indicated that savannah sparrows require 

open areas of 10-15 ha in grasslands and blueberry barrens (O’Leary and Nyberg, 2000) and in South 

Dakota, savannah sparrows were found to be area-sensitive and used only large grassland patches 

(Bakker and others, 2002). 

Foraging Habits

During the breeding season, savannah sparrows forage primarily on adult and larval insects, 

spiders, seeds and fruits (Wheelwright and Rising, 2008). Occasionally, they will eat insect eggs, 

millipedes, isopods, amphipods, decapods, mites, and small mollusks (Baird, 1968; Maher, 1979; 

Meunier and Bedard, 1984). Savannah sparrows forage on the ground in short vegetation in pastures, 

sedge bogs, tundra, beach wrack, mud flats (near cover), and dunegrass (Wheelwright and Rising, 

2008). In winter, savannah sparrows feed primarily on small seeds of grasses and other herbs, smallseeded fruits, insects, and spiders (Desmond and others, 2008). They forage on the ground in grasslands, 

cultivated fields, roadsides, dunegrass, and salt marshes (Ginter and Desmond, 2005). 

Breeding

Although territories that the males defend had very small dimensions (0.11 ha) in Michigan, the 

defended area often had unoccupied space around it (Ryser, 1985). Males spend only about 66 percent 

of their time on the territory and feed in nonterritorial common areas. In contrast, females feed mostly 

within their mates’ territories. Pairs whose members survive typically breed together in subsequent 

years (Wheelwright and Rising, 1993). The female alone builds the nest and it is usually well hidden on 

the ground. Nests are concealed under grass clumps with a canopy of dead grasses and herbs. Some 

populations have second broods, but this depends on climate and whether the male assumes postfledging care of the first brood. Clutch size ranges from two to six eggs across their range, but with a 

latitudinal gradient from north to south of larger clutches in the northern parts of the range 

(Wheelwright and Rising, 2008). Only the female incubates the eggs (Potter, 1974; Stobo and McLaren, 

1975; Baird, 1968; Bedard and Meunier, 1983). 

Brown-headed cowbirds may lay eggs in savannah sparrow nests where the two species co-occur 

(Freidmann and others, 1977). However, recorded frequency of brood parasitism is low (Wheelwright 
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and Rising, 2008). Four records of brown-headed cowbird nest parasitism exist for Colorado, including 

two records from Colorado’s Breeding Bird Atlas (Toolen, 1998). 

Demography & Populations

Age at first breeding is one year for both males and females (Wheelwright and Rising, 2008). 

Overall reproductive success is apparently higher on islands because of an absence of mammalian 

predators (Stobo and McLaren, 1975), and reproductive output is low during years of inclement 

weather. 

Annual survivorship of breeding adults in Nova Scotia averaged 35 percent (28-42 percent; 

Stobo and McLaren, 1975); Quebec, 31-45 percent (Bedard and LaPointe, 1984a; 1984b); Kent Island, 

New Brunswick, 45-50 percent (Dixon, 1972); and, Michigan, 68-70 percent (Potter, 1974). About 40 

percent of adults are estimated to return to the same wintering area in successive years (Odum and 

Hight, 1957). Adult survival not only varies geographically, but varies widely among years (37-73 

percent; Wheelright and others, 1992). 

Densities of savannah sparrows are greatest on islands, possibly because of extreme philopatry, 

fewer competing species, and less predation because of fewer mammalian predators (Wheelwright and 

Rising, 2008). On Kent Island, New Brunswick, as many as 20 females breed per hectare in some 

habitats, although more typical densities are 4-5 pairs/ha. Density of savannah sparrows declines with 

plant succession as shrubs invade open areas (Potter, 1974; Weatherhead, 1979a). Winter densities in 

habitat such as agricultural fields was reported to be 4.0 birds/ha, but as high as 30 birds/ha where food 

and cover were plentiful (Norris, 1960). In addition to islands, greatest densities of savannah sparrows 

are found along the St. Lawrence River, around the Great Lakes, in northern Minnesota, Saskatchewan 

and Manitoba, and in the Pacific Northwest (Wheelwright and Rising, 2008). 

Populations of savannah sparrows increased significantly in western North America from 1969 

to 1989 (BBS unpublished data in Wheelwright and Rising, 1993) while declining slightly across the 

continent from 1966-1996. However, when BBS results were combined from 1966 to 2005, savannah 

sparrows showed significant population declines in the United States and in Canada. In the Eastern and 

Western BBS survey regions, populations also declined, but populations increased significantly in the 

Central region (Wheelwright and Rising, 2008). Populations declined notably around the Great Lakes 

and Illinois east to western Pennsylvania (Sauer and others, 2011). As with many other grassland birds, 

more intensive agriculture, the shift to cash crops (corn, wheat, soybean), and declines in dairy farming 

have resulted in decreasing population sizes, based on BBS data (Jobin and others, 1996). 

The Breeding Bird Survey (BBS) regional trend analysis for savannah sparrows along 71 

Colorado routes from 1966 to 1999 showed a slight positive trend estimate of 0.2 (-1.2, 1.8; 95 percent 

CI) (Sauer and others, 2011). From 1999 to 2009, the trend was also positive and estimated to be 1.9 (-

0.8, 4.9; 95 percent CI). The regional credibility measure for this species was the category “blue.” A 

blue category reflects data with at least 14 samples in the long term, of moderate precision, and of 

moderate abundance on routes. 

Threats

Sources of Mortality

In 3-4 percent of broods of savannah sparrows in eastern Quebec, nestlings were infested 

subcutaneously with larval blow flies (Calliphoridae), which retarded feather growth and, in one case, 

caused the loss of an eye (Bedard and McNeil, 1979). Ectoparasites are uncommon on adult birds 

(Wheelwright and Rising, 2008). Avian louse flies (Hippoboscidae; Baird, 1968) were found on 10 
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percent of recently independent juveniles; were less commonly found on adults. Fewer than 1 percent of 

adults carry ticks and when found, they are usually around the face. There are various blood protozoans 

known to infect savannah sparrows (Baird, 1968). 

Nest predation is the primary source of nest failure for grassland passerines, including savannah 

sparrows (Koford, 1999; Jones and Dieni, 2007). Small mammals, considered the primary threat to 

ground nests in the northern Great Plains (Pietz and Granfors, 2000; Davis, 2003), opportunistically find 

grassland bird nests while foraging for invertebrates. Data from Kent Island, New Brunswick, indicated 

the daily mortality rate of eggs and nestlings was 11.9 percent and 2.1 percent, respectively, and the 

author determined that predation was the major factor affecting reproductive success (Dixon, 1978). In 

this location, the main nest predators were common crows and herring gulls. 

Effects of Human Activity

Across their range, savannah sparrows have generally benefitted from human activity, at least 

until the past few decades (Wheelwright and Rising, 2008). Twentieth century savannah sparrow 

densities are probably greater than at any time in the past because of this species’ dependence on open 

habitats (fields, blueberry barrens, hay fields, cropland) for breeding and over-wintering. Post-war 

urbanization in the eastern United States and reversion of marginal farms to forest may partly explain 

population declines in the Northeast and around the Great Lakes. Landscape changes caused by the shift 

in agricultural practices from dairy-farming to cash crops (for example, corn, wheat, soybean) have been 

associated with declines in savannah sparrow densities in Quebec (Jobin and others, 1996). Forestclearing and more intensive agriculture in the west may explain population increases there until 

recently.

Breeding savannah sparrows are relatively tolerant of human disturbance at the nest (Baird, 

1968; Welsh, 1975). Pairs nest readily in human-made grasslands such as reclaimed surface mines, but 

reproductive success in such habitats may be so low that they effectively serve as population “sinks” 

(Wray and others, 1982). Pesticide use in agricultural fields probably has negative effects on savannah 

sparrows, but so far such effects have been documented only in the case of dimethoate, an 

organophosphate which reduced mean daily activity levels of the birds, which then had a recovery time 

from intoxication of 11.5 days (Brunet and Cyr, 1992). Savannah sparrows appear more likely than 

other birds to ingest pesticide granules applied to cornfields (Wheelwright and Rising, 2008). 

Management

Mowing/Haying

According to Wheelwright and Rising (2008), grassland management, particularly mowing 

regimes, is critical for maintaining mainland populations of savannah sparrows. Mowing of grassland 

fields has at least four effects on grassland birds: (1) it alters the vegetative characteristics of the field, 

changing its suitability for individual grassland species (Owens and Myres, 1973; Dale and others, 

1997); (2) it displaces individuals from a field, causing an egress of birds to nearby fields (Bryan and 

Best, 1991; Igl, 1991); (3) mowing in the breeding season destroys or causes abandonment of a large 

proportion of the nests that were active just prior to mowing (Bollinger and others, 1990; Frawley and 

Best, 1991; Kershner and Bollinger, 1996; Dale and others, 1997) and repeated mowing may not allow 

birds to complete their nesting cycles (Bollinger and others, 1990; Herkert and others, 1960; Koford and 

Best, 1996); and, (4) fields mowed during the breeding season may have lower breeding densities the 

following year due to greater nest failure experienced previously and reduced site fidelity (Bollinger, 

1995; Dale and others, 1997). Rodenhouse and others (1995) synthesized the effects of mowing on 
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grassland birds and suggested that mowed grassland fields would be less suitable for some species (for 

example, bobolinks) than idled fields, whereas mowed fields would be more suitable for other species 

(for example, savannah sparrows). 

Earlier and more frequent harvests in agricultural grasslands in recent decades have been 

implicated as major cause of population declines in savannah sparrows and other grassland songbirds 

(Kirsch and others, 1978; Wheelwright and Rising, 2008). Hayfields are unique breeding habitats in that 

dramatic structural changes occur due to harvesting during the breeding season (Zalik, 2007). These 

harvesting processes, along with the subsequent predation of exposed nests, can cause nearly 100 

percent failure of all active nests (Perlut and others, 2006). In Saskatchewan, breeding densities were 

greater in idle hayfields than in annually mowed fields; mowing resulted in an 80 percent decline in 

productivity (Dale and others, 1997). However, in North Dakota on Conservation Reserve Program 

lands, savannah sparrows were more abundant in portions of fields that were mowed the previous year 

(Horn and Koford, 2000). Haying before June 11 caused 99 percent nest failure in Vermont; delaying 

haying until at least June 21 results in 2.7-fold increase in the number of fledglings per female per year 

(Perlut and others, 2006). In Nova Scotia, a one-week delay in hay-cutting increased fledging success 

from 0 to 56 percent, with minimal loss of hay nutritional value (Nocera and others, 2005). 

Zalik (2007) studied savannah sparrows in both harvested and unharvested hayfields in the 

Champlain Valley of Vermont. The removal of vegetation through haying likely influences the 

invertebrate community, which is a primary diet item for both adults and young (Vickery and others, 

2001; Schmidt and others, 2005; Britschgi and others, 2006). Any effects on these food sources may 

influence the fitness of young and adult savannah sparrows. Zalik (2007) studied food habits of breeding 

savannah sparrows and found that caterpillars and sawfly larvae were consistently used more than 

expected based upon their availability. This was true, even later in the season, particularly on cut fields, 

when larvae constituted only a small proportion of available prey, which lead to the conclusion that 

adults may have been able to provide their nestlings with relatively more larvae on cut fields. 

Perlut (Perlut, 2007; Perlut and others, 2008) studied the effects of hayfield management on 

grassland songbirds in the same area as Zalik (2007). The ecology of savannah sparrows was studied in 

four grassland management types: (1) early-hayed fields harvested in late-May or early-June and again 

in mid-July; (2) middle-hayed fields harvested in late-June or early-July; (3) late-hayed fields harvested 

after 1 August; and, (4) rotationally-grazed pastures, a matrix of small paddocks where cows were 

moved after the grass in a paddock was eaten to a low point. This study found that early-hayed fields 

were low-quality habitats for reproduction. Female savannah sparrows on middle- and late-hayed fields 

fledged 2.6 – 2.7 times more young each year than females on early-hayed fields, even though females 

on early-hayed fields built more nests and laid more eggs. Predation rates between the first and second 

haying were highly variable. In extreme years, predation caused 59-69 percent of the failures on earlyhayed fields, while in other years only 9-18 percent of failed nests were depredated. High predation 

resulted in asynchronous nesting, which again caused more nests to be vulnerable to a second mid-July 

haying. Perlut (2007; Perlut and others, 2008) also found that savannah sparrows breeding in lessintensively managed grasslands (late-haying) incurred higher apparent survival than those breeding in 

intensively managed grasslands (early-haying, middle-haying, and grazing). Birds using late-hayed 

fields had > 25 percent higher survival than those on more intensively managed early-hayed, middlehayed, and grazed fields. Recruitment was highest on both intensively and moderately managed fields, 

perhaps as result of the lower apparent survival on these fields. For the Champlain Valley area of 

Vermont, Perlut and others (2008) recommended that advancing the initial harvest of hayfields in lateMay and delaying the second harvest for 65 days would benefit reproduction, recruitment, and survival 
126

for savannah sparrows. They also suggested increasing the preference for late-hayed fields by removing 

grass at the end of each season.

Dieni (2009) examined the effects of haying on breeding bird community and habitat structure at

the Baca NWR in 2009. Idle units were rested during 2008, while treated units were hayed during 

August and September of 2008. He found that the species composition of breeding birds and relative 

abundance of the most dominant species was structurally similar between haying and idle treatments. 

However, savannah sparrows were on average 20 percent more abundant on the idle units, although the 

effect varied with visit (Dieni, 2009). Little difference was observed during the late June visits; 

conversely, savannah sparrows were over 43 percent more abundant on the idle units during late July. 

Possible explanations for the increased abundance on idle units included (1) idle treatments were more 

productive, (2) increased fledgling activity, and (3) idle units tended to have a greater percentage of area 

that lacked standing water, thus more potential nest sites, since savannah sparrows place nests directly 

on the ground (Wheelwright and Rising, 1993). Dieni (2009) suggested that a long-term monitoring 

study would be instructive where one or more units are left idle indefinitely, with hayed controls for 

comparison. Such a monitoring study would require a well-conceived monitoring plan, and a 

commitment to measure both avian and vegetative parameters at regular intervals. This would not 

necessarily have to happen annually. 

Blank and others (2011) examined the effects of fall mowing of herbaceous buffers on wintering 

savannah sparrows on the eastern shore of Maryland. Mowing in this area is a common maintenance 

practice to control woody plants and noxious weeds in herbaceous buffers. Buffers cannot be mowed 

during the primary bird nesting season between 15 April and 15 August. Therefore, most mowing 

occurs in the late summer and early fall. Savannah sparrows were significantly more abundant in 

unmowed buffers, suggesting that unmowed buffers provide better habitat for wintering savannah 

sparrows.

Warren and Anderson (2005) studied the effects of mowing on savannah sparrows at the Canaan 

Valley National Wildlife Refuge, West Virginia. Mowing took place after the breeding season in late 

August. They found that nest success rates were similar between mowed and unmowed treatments, 

indicating that although vegetative structure was altered, mowing did not influence overall nest-failure 

rates. 

Grazing

Owens and Myres (1973), in a southern Alberta study, found that disturbance of a native fescue 

(Festuca scabrella) grassland by mowing or cattle grazing did not affect savannah sparrow populations. 

In Saskatchewan, however, Maher (1973) found that four of seven species that regularly breed in 

grasslands prefer ungrazed prairie, including the savannah sparrow. 

Kwasny (2000) examined the effects of cattle grazing on savannah sparrows in northcoastal 

California at the Humboldt Bay National Wildlife Refuge. Cattle grazing is used to create short-grass 

habitat for wintering and migrating shorebirds and waterfowl. Kwasny looked at territory size, density, 

and reproductive success of savannah sparrows in relation to vegetation cover and structure in yearround grazed, seasonally grazed, and ungrazed pastures during 1995 and 1996. He found that territory 

size did not differ among grazing regimes, but territory density and reproductive success were both 

higher in the ungrazed treatments compared to the grazed treatments. Vegetation structure was taller and 

more homogeneous in the ungrazed treatments, and savannah sparrow territory density and reproductive 

success were positively correlated with the average height of vegetation. He suggested delaying grazing 

until after 1 June or later would greatly increase the density and nesting success of savannah sparrows 
127

on the pastures. A delay in grazing would also allow young birds to become mobile and avoid being 

trampled by cattle.

Rotationally and continuously grazed pastures in riparian areas in Wisconsin held equal numbers 

of savannah sparrows, although birds avoided buffer strips; higher densities occurred on sites with 

deeper litter (Renfrew and Ribic, 2001; 2002). However, intensive grazing substantially reduced 

savannah sparrow densities on islands in the St. Lawrence River, Quebec (Blank, 1999). 

On northern mixed-grass prairies protected for conservation, active management such as grazing 

and burning is considered necessary to restore and maintain ecological integrity (Kerns and others, 

2010). At Lostwood National Wildlife Refuge in northwestern North Dakota, a prescribed burn program 

was implemented in the late 1970s to reduce exotic and woody vegetation. Because fire was not 

adequately promoting native grasses or reducing exotic smooth brome (Bromus inermis), rotational 

grazing was added in the 1990s. Kerns and others (2010) investigated the influence of rotational grazing 

on nest survival of 3 passerine species, including the savannah sparrow. Nest survival of savannah 

sparrows was found to be lower when cattle were present than when cattle were absent, and the authors 

suggested that limiting the area actively grazed during the breeding season could reduce nest mortality 

of passerines.

Cattle may actively predate savannah sparrow nests. In Wisconsin in 2000-2001, Nack and Ribic 

(2005) used video cameras to document cows removing three of four savannah sparrow eggs from one 

nest and all three nestlings from a second nest. Seven of 14 (50 percent) cattle-caused nest failures were 

inadvertent disturbances; a cow laid on one nest, one was abandoned, two were trampled, and the 

camera was knocked down at three nests, crushing the eggs. They postulated that cows may have eaten 

the egg and nestlings, or alternatively, the egg and nestlings may have been scavenged by predators or 

removed from the area by adult birds. 

Burning

In southeastern Arizona, Gordon (2000) studied the effects of spring/summer burning and cattle 

grazing on wintering sparrows. The effects of fire were studied with 1 year of pre-burn data and 1 year 

of post-burn data from one fire, plus limited sampling from a second fire. The effects of grazing were 

studied by comparing study plots at a site that had not been grazed by cattle since 1968 with a nearby 

grazed pasture. Savannah sparrows were more abundant on burned plots before the fire. In the winter 

subsequent to the fire, no difference was detected, suggesting that the savannah sparrows responded 

positively to the burn. The effects of cattle grazing on wintering savannah sparrows were less clear; 

limited data suggested that severe grazing negatively effects the abundance of sparrows.

In another study in southeastern Arizona, Bock and Bock (1992) studied the response of birds to 

wildfire in native versus exotic grasslands. A wildfire in 1987 completely burned 11 native and 11 

exotic grassland plots. The fire reduced grass and shrub cover, and increased herb cover, for 2 post-fire 

years in both grassland types. Numbers of fall birds, including savannah sparrows, increased 

dramatically on burned plots in both native and exotic plots for 2 years, probably in response to 

increased seed production and availability. 

In a study of the effects of fire on prairie grassland birds at the Texas Mid-coast National 

Wildlife Refuge Complex, savannah sparrows were most highly associated (almost exclusively) with 1-

year-since-burn conditions and where vegetation was sparse (Grace and others, 2005). Conversely, in 

north-central North Dakota, the number of savannah sparrow pairs were lowest during the first postfire 

growing season but generally increased and stabilized within 2-3 years postfire (Grant and others, 

2010). At another site on the Texas gulf coast in tallgrass prairie, wintering savannah sparrows were 

also found to prefer recently burned patches (Marx and others, 2008). Baldwin and others (2007) also 
128

found savannah sparrows associated with areas burned within one year, with sparse herbaceous 

vegetation, and with sparse shrub densities (Baldwin and others, 2007). 

Conversely, in Saskatchewan, savannah sparrows were adversely affected by a prescribed burn 

in a fescue grassland (Pylypec, 1991). By the third post-burn year, the density of savannah sparrows was 

68 percent of the value observed in the unburned area. In this particular location, only vesper sparrows 

showed a preference for the burned area in the first year after the burn. 

Other Management

Below is a selection of studies found in the literature that examined the effects of different 

management schemes on savannah sparrows:

• Thompson and others (2009) examined the relative abundance of grassland birds, including 

savannah sparrows, among Conservation Reserve Program (CRP) fields seeded with 2 

monocultures of introduced grass species and 2 mixes of native grasses in the southern High 

Plains of Texas. The introduced fields were monocultures of weeping lovegrass (Eragrostis

curvula) and Old World bluestem (Bothriochloa ischaemum). The native fields were mixes of

principally sideoats gramma (Bouteloua curtipendula), blue gramma (Bouteloua gracilis), green 

sprangletop (Leptochloa dubia), switchgrass (Panicum virgatum), and buffalograss (Buchloe

dactyloides). No breeding season differences among cover types in total avian abundance or 

species richness were observed. In contrast, during the winter season more species and a greater 

abundance of grassland birds were detected among fields with native seeding than with 

introduced seeding. Savannah sparrows were 96 percent more abundant among native CRP than 

introduced CRP fields. 

• Martin and Forsyth (2003) studied the occurrence and productivity of savannah sparrows in 

prairie farmland under conventional and minimal tillage regimes in Alberta. In conventional 

tillage, crops were frequently plowed, disced and cultivated after harvest in the fall. Fields then 

received 2-5 passes with a cultivator or disc, and were harrowed prior to planting, reducing 

surface crop residue to less than 30 percent of the soil surface. In minimum tillage, crops were 

seeded directly into standing stubble of the previous crop. They found that savannah sparrows 

preferred areas of minimum tillage and had the highest productivity in spring cereal and winter 

wheat. 

• Bock and others (1999) examined grassland bird abundance at a suburban interface in Boulder, 

Colorado. They counted nesting songbirds for three summers on sixty-two 200-meter-diameter 

plots and found that savannah sparrows were significantly more abundant on interior plots, 

independent of habitat type. Most of the grassland birds in their study avoided suburban edges. 

Bock and others (1999) suggested that if the management goal is to conserve native grassland 

birds on the western Great Plains, grassland open-space systems in the region should be 

designed to reduce edges with suburban development. 

• Graves and others (2010) studied the influence of woody vegetation on grassland birds within 

reclaimed surface mines and found that savannah sparrows were negatively associated with 

percent cover of woody vegetation within 100 m of survey locations. They suggested that 

management efforts that remove woody vegetation would enhance the value of reclaimed 

surface mines for grassland birds. Removal of encroaching woody plants would increase the 

grassland area and reduce edges and fragmentation. This practice may also decrease numbers of 

woodland predators by eliminating cover or perches and reducing movement corridors. 

• Savannah sparrows declined significantly in response to an impoundment and subsequent 

increase in water level on the shores of Lake Sainte-Anne, Quebec (Lariviere and Lepage, 2000). 
129

Although rising water levels benefited certain avian species such as waterbirds, low-nesting 

species were negatively affected.

• Daily nest survival probabilities for savannah sparrows were found to be higher in 

pastures/hayfields than in fallow fields in western New York (Norment and others, 2010). 

• The effects of an invasive plant, leafy spurge (Euphorbia esula), infestation on grassland birds 

was examined in North Dakota on the Sheyenne National Grassland (Scheiman and others, 

2003). Savannah sparrow density was lower on high-spurge points than on low and mediumspurge points. However, spurge cover did not appear to be an important factor in nest-site 

selection. 

• Both mowing and burning were used to restrict shrubland encroachment and maintain grassland 

habitat on Nantucket Island, Massachusetts (Zuckerberg and Vickery, 2006). Management 

included prescribed burning on 142 ha of scrub oak shrubland and >26 ha of 

grassland/heathland, and repeated mowing and brush cutting on 205 ha of shrubland. The 

frequency of the management activities differed among study sites: shrubland areas were burned 

no more than once, and mowing frequently ranged from 0 (control areas) to 1-3 cuts annually. 

Bird surveys were conducted during the breeding season. Savannah sparrow abundance did not 

differ among grasslands that had been burned, mowed, or left unmanaged. 

Research Priorities

Dieni (2009) suggested that a long-term monitoring study of savannah sparrows would be 

instructive in which one or more units are left idle indefinitely, with hayed controls for comparison. 

Such a monitoring study would require a well-conceived monitoring plan, and a commitment to 

measure both avian and vegetative parameters at regular intervals. This would not necessarily have to 

happen annually. 

Vickery and Herkert (2001) suggested that research directed at determining the causes of 

grassland-bird population declines needs to consider the possibility that population problems may be 

occurring on winter grounds. The wintering ecology of many grassland species is poorly known. 

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137

IX. Northern Leopard Frog 

[Lithobates (=Rana) pipiens] 

Introduction 

The northern leopard frog [Lithobates 

(=Rana) pipiens] is a widespread species that has 

experienced significant declines across most of its 

range, while remaining abundant in some areas 

(Smith and Keinath 2007). [For the remainder of 

this section, I will use the abbreviation NLF for 

the northern leopard frog.] In the Rocky 

Mountain region, which includes U.S. Fish and 

Wildlife Service Region 6, Colorado and 

Montana legally protect the species and the 

USDA Forest Service (USFS) Rocky Mountain Region (Region 2) has listed the NLF as a sensitive 

species. In 2009, the USFWS announced a 90-day finding on a petition to list the western population of 

the NLF as federally threatened under the Endangered Species Act of 1973 (U.S. Fish and Wildlife 

Service 2009). The petition provided substantial information suggesting that the western population is 

genetically distinct from the eastern NLF (Hoffman and Blouin, 2004). The NLF is experiencing threats 

from habitat loss, disease, nonnative species, pollution and climate change that individually and 

cumulatively have resulted in population declines, local extinctions, and disappearance from vast areas 

of its historical range in the western United States and Canada (U.S. Fish and Wildlife Service 2009). 

The 90-day finding indicated that listing of the western population of the NLF may be warranted. 

There is a large amount of literature available on the NLF and it is one of the best known frogs 

in North America. A simple Google search on the Internet for “Lithobates and Rana pipiens” revealed 

257,000 results and a Google Scholar search provided 16,200 results (accessed 6 June 2011). “Northern 

leopard frog” revealed 237,000 results through Google and 2,180 results through Google Scholar 

(accessed 6 June 2011). 

Natural History

Morphological Characteristics

The NLF is a medium-sized ranid with a maximum snout-vent length (SVL) of 11 cm 

(Hammerson, 1999). It has a green or brown dorsum with large rounded or oval spots, the skin is 

smooth, the eardrum usually lacks a distinct light spot, the hind legs are powerful and it has extensively 

webbed hind feet that are well adapted for jumping and swimming (Pace, 1974). It is also characterized 

by conspicuous dorsolateral ridges bordering the spots at the edge of the dorsum (Smith and Keinath, 

2007). Males have swollen thumbs on their forefeet during the breeding season and paired vocal 

pouches at the sides behind the head, which are visible when vocalizing. Males are also typically 

smaller than females (Smith and Keinath, 2007). Tadpoles of most frogs are difficult to identify without 

a dissecting microscope and expert advice. The tadpoles of this species are brown, olive, or gray with 

the snout pointed in the dorsal view (Hammerson, 1999). The maximum tail length for tadpoles in 

Colorado is 10 cm (Hammerson, 1999). The eggs of NLFs are black from above, whitish from below, 

with a diameter averaging 1.3-2.3 mm. Eggs are typically enclosed in 2-3 jelly envelopes and deposited 

Northern Leopard Frog, courtesy of Laurie Averill-Murray, U.S. Fish 

and Wildlife Service, Pacific Southwest Region. 
138

in large, somewhat flattened globular masses 5-15 cm in maximum diameter that include up to several 

thousand eggs (Hammerson, 1999). 

Distribution and Abundance

NLFs are widely distributed in North America, from southern Canada and the northern United 

States south to Maryland, West Virginia, Kentucky, northern Illinois, extreme northwestern Missouri, 

Nebraska, New Mexico, Arizona, and eastern California (Conant and Collins, 1998; Hammerson, 1999; 

Stebbins, 2003). NLFs occur throughout Colorado, excluding most of the southeastern and east-central 

portions of the state (Hammerson, 1999). Their elevational range extends from 1,065 m in northeastern 

Colorado to above 3,355 m in southern Colorado (Hammerson, 1999). 

The NLF was formerly abundant across its entire range, but has since suffered wide-ranging 

population declines (Smith and Keinath, 2007). Particularly in the western United States, the species has 

undergone major declines and suffered extirpations in some areas (Rorabaugh, 2005). Generally, it is 

thought that eastern populations are in somewhat better condition than western populations (Smith and 

Keinath, 2007). In Colorado, NLFs have become scarce in many areas and some populations have 

disappeared due partly to changes in habitat (Hammerson, 1999). Suitability of breeding sites can 

change in response to climatic variation and flooding, though some sites may be created by these events. 

In Larimer County, Colorado, from 1973 to 1982, Corn and Fogelman (1984) documented the 

extirpation of nine populations at elevations from 2,365 to 2,520 m. Six of these nine populations 

disappeared due to drought-related drying of breeding ponds and the other three may have been 

extirpated due to random events that can befall small populations. Local extirpations and recolonizations 

are most likely common features in the ecology of the NLF in the drought-prone climate of Colorado 

(Hammerson, 1999); however, some disappearances in Colorado, especially along the Front Range, do 

not have an adequate explanation (Corn and others, 1989; 1997; Hammerson, 1989; 1992; Livo, 1995). 

Some reduced or extirpated NLF populations in lowland areas of Colorado are associated with 

the presence of the increasingly abundant bullfrog (Rana catesbeiana) (Hammerson, 1982). Purposeful 

introductions of bullfrogs to Colorado began in the 1900s into the South Platte River drainage; since 

that time bullfrogs have been reported from 30 of Colorado’s 63 counties (Livo and others, 1998). NLFs 

are now greatly outnumbered by bullfrogs in many streams, ponds, reservoirs, and wetland areas, 

especially in eastern Colorado. In some areas, human alterations of wetlands have increased water depth 

and permanence, resulting in the replacement of NLF by bullfrogs and fishes, especially near large 

rivers and other permanent bodies of water (Hammerson, 1999). Sites in eastern Colorado where NLFs 

occur in greatest abundance are small, relatively isolated reservoirs that have not been colonized by 

bullfrogs or predatory fish. Lannoo and others (1994) conducted a survey of native amphibian species in 

Iowa and concluded that bullfrogs were the most immediate threat to existing populations of native 

amphibians. In Ontario, Hecnar and M’Closkey (1997b) found significant changes in ranid relative 

abundance after a bullfrog extinction and concluded that bullfrogs are important in structuring 

amphibian communities through direct predation and competition for resources. 

Hecnar and M’Closkey (1997a) also examined the effects of predatory fish on amphibian 

communities in Ontario. They found that amphibian species richness was significantly lower at ponds 

having predatory fish present than at nonpredatory, or fish-free, ponds. Numbers of NLFs also occurred 

in proportionately more nonpredatory fish ponds because NLF larvae are palatable to fish, unlike 

bullfrog larvae, which are unpalatable to fish (Kats and others, 1988). Nonnative predatory fish were 

also implicated in the decline of the Cascades frog (Rana cascadae) in Lassen Volcanic National Park, 

California. The introduction of nonnative fish has eliminated local amphibian populations in many parts 
139

of North America (Hayes and Jennings, 1986; Sexton and Phillips, 1986; Bradford, 1989; Liss and 

Larson, 1991). 

Declines of NLFs in lowland areas of Colorado are not always attributable to predation by 

bullfrogs or fish. Flood control measures and diversion of water for irrigation probably have reduced the 

availability of breeding habitat along floodplains in the lowland segments of some streams 

(Hammerson, 1999). Additionally, although NLFs have disappeared from many natural breeding sites in 

Colorado, they still occur in many human-created habitats, both in the lowlands and in the mountains. It 

is unknown how natural habitat loss has been offset by the addition of artificial habitat for this species. 

Habitat

Because NLFs have a complicated life history, they require a broad range of habitats in close 

proximity. Three categories of major habitat type were described by Merrell and Rodell (1968): winter 

habitat (overwintering in lakes, streams, and ponds); summer habitat (feeding by adults in upland areas); 

and, tadpole habitat (up to three months spent as tadpoles in shallow breeding ponds). Their complicated 

movement patterns during the year should also be considered (Smith and Keinath, 2007). Relevant 

literature found describing NLF habitat comes from populations scattered across North America. 

Therefore, it is likely that there are differences in habitat use regionally.

Typical habitats of NLFs described in Colorado include wet meadows and the banks and 

shallows of marshes, ponds, glacial kettle ponds, beaver ponds, lakes, reservoirs, streams, and irrigation 

ditches (Hammerson, 1999). Adult NLFs are usually found at the water’s edge, but they may roam far 

from permanent water during mild wet weather. Winter months are spent at the bottom of water bodies 

such as ponds and streams. Cunjak (1986) observed NLFs hibernating in a Canadian river in winter 

beneath the rubble, generally where the water depth was greater than 85 cm. The frogs were torpid, but 

capable of swimming. Emery and others (1972) observed NLFs hibernating in ice-covered ponds in 

Ontario, Canada. The frogs were resting in small excavated pits, lightly covered with silt, and capable of 

some movement including swimming.

Activity and Movement Patterns

In the plains region of Colorado, NLFs typically emerge from their winter retreats in March 

(Hammerson, 1999). The frogs are then active until October or November, when the colder temperature 

forces them into torpor, or dormancy. However, frogs have been found to be active in January and 

February in the San Luis Valley (SLV) at pools formed by discharge of warm artesian wells (Hahn, 

1968). NLFs can move long distances. Gillis (1975) marked a frog in a pond in Cheyenne County, 

Colorado, and found the same individual 3 km away at another pond the following year. 

Dole (1965) studied movement patterns of adult NLFs in Michigan by trailing individuals with a 

spool of thread. This study found that the adult frogs are rather sedentary and generally confined their 

activities to relatively limited areas (within 5-10 m), but on rainy nights, individuals embarked on 

excursions that would involve distances of more than 100 m across habitats not usually inhabited by 

frogs. Conversely, newly metamorphosed NLFs in Michigan would disperse much further, commonly 

traveling up to 800 m in a few days (Dole, 1971). Climatic variables such as precipitation or minimum

air temperatures are generally good predictors of amphibian activity (Mazerolle, 2001). 

Mazerolle and Desrochers (2005) examined landscape resistance to NLF movements in New 

Brunswick and found that 72 percent of frogs avoided disturbed surfaces and lost the most water on 

substrates associated with disturbance. They concluded that anthropogenically disturbed areas devoid of 

cover, such as agricultural fields, disrupt the ability of frogs to reach habitat patches and are a likely 

explanation for the reduced abundance of frogs in agricultural areas.
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Reproduction and Growth

NLFs breed in shallow, quiet areas of permanent bodies of water, in beaver ponds, and in 

seasonally flooded areas adjacent to or contiguous with permanent pools or streams (Hammerson, 

1999). Breeding areas contain vegetation, mats of algae, and clear water. At lower elevations in 

Colorado, males begin calling on warm, sunny days in March or April. Males will call while floating on 

the surface of the water and will frequently wrestle with other males between calling bouts. 

Females begin laying eggs a few days after calling begins (Hammerson, 1999). In the plains 

region of Boulder County, Colorado, most females lay their eggs by mid-April, sometimes by the end of 

March if warm weather arrives early. Along the Arkansas River in Pueblo County, Post (1972) found 

that the breeding season can be variable from year to year (April 15 to May 20, April 24 to May 7, and 

April 12 to May 16 in consecutive years). Females produce only one clutch per season (Hammerson, 

1999). 

Egg masses are attached to vegetation just below the surface of warm shallows about 7 to 25 cm 

deep. At elevations of 1,570 to 2,520 m in Colorado and southern Wyoming, egg masses average about 

3,000 eggs (range 645-6,272) (Corn and Livo, 1989). There does not appear to be a relationship between 

clutch size and elevation. In the plains region of Boulder County, eggs hatch in 4-15 days (Livo, 1981). 

Egg masses deposited in late March are often covered with tiny hatchlings by mid-April (Hammerson, 

1999). Early clutches may hatch before later clutches are deposited (Livo, 1981). Newly 

metamorphosed individuals have been found in late June and early July (Hammerson, 1999). Frogs do 

not begin to breed until their second spring. In the SLV, calling may continue until late June 

(Hammerson, 1999). Hahn (1968) observed newly metamorphosed individuals in the SLV on July 24. 

Food Habits

There is little information on the food habits of the NLF in Colorado; however, invertebrates 

most likely dominate the diet of adult frogs (Hammerson, 1999). Gehlback (1965) reported that NLFs 

ate tenebrionid beetles in northwestern New Mexico. Drake (1914) studied the food habits of NLFs in 

Ohio and found that the diet consisted of molluscs, crustaceans, spiders, and insects. Hendricks (1973) 

investigated the intestinal contents of NLF tadpoles in Mexico and found 10 genera of algae that could 

be positively identified and concluded that tadpoles are filter feeders on free-floating algae suspended in 

the aquatic environment. 

Predators

Predators of NLFs vary depending on the stage of their life history. Known predators on larvae 

in Colorado include pied-billed grebes (Podilymbus podiceps) (Hammerson, 1999) and tiger 

salamanders (Ambystoma tigrinum) (Corn, 1981). Garter snakes (Thamnophis radix) were reported to 

predate larvae in western Nebraska (Ballinger and others, 1979). Predators on metamorphosed 

individuals in Colorado include the great blue heron (Ardea herodias) (Hammerson, 1999), burrowing 

owl (Athene cunicularia) (Hamilton, 1941), northern water snake (Nerodia sipedon), and western 

terrestrial garter snake (Thamnophis elegans) (Cockerell, 1910; Hammerson, 1999). Adults and larvae 

are also vulnerable to predation by various game and nonnative predatory fishes, as previously noted 

(Bagdonas, 1968). Eggs and small larvae are most likely vulnerable to predation by bullfrog larvae 

based on experiments with the plains leopard frog (Rana blairi) (Ehrlich, 1979).
141

Threats

There are extrinsic and intrinsic threats to NLF populations in the Rocky Mountain region. 

Extrinsic threats can fall into three main categories: 1) landscape-scale processes that threaten the 

viability of populations; (2) direct threats of mortality from various nonindigenous biotic agents; and, 

(3) water quality issues (Smith and Keinath, 2007). Intrinsic factors considered a threat to NLF

populations are those factors related to their unique habitat use requirements, reproductive biology, and 

physiology.

Extrinsic Threats 

Landscape-level Processes

Landscape-scale changes such as loss and fragmentation of breeding habitat, disruption of 

migratory pathways, and loss or alteration of over-wintering sites, may have caused range-wide declines 

in NLFs (Smith and Keinath, 2007). Habitat loss and fragmentation in the Rocky Mountain region are 

caused by a variety of land uses, including urbanization and growth of suburbs, logging, recreation, and 

building of roads and trails (Smith and Keinath, 2007). In the mid-1950s there were an estimated 44 

million hectares of wetlands in the lower 48 states (Frayer and others, 1983; Tiner, 1984). In just 20 

years, these wetlands were reduced to 40 million hectares (Tiner, 1984), and this trend has continued 

into this century despite a national policy of no net loss of wetlands through mitigation measures 

(Turner and others, 2001). 

Human development of wildlands has shown to negatively affect amphibian populations at a 

regional scale (for example, Hecnar and McCloskey, 1996). In Minnesota glacial ponds, amphibian 

species richness at ponds declined with distance to other ponds and with increased road density, both 

common results of urbanization (Lehtinen and others, 1999). Trombulak and Frissell (2000) conducted a 

review of the effects of roads on terrestrial and aquatic communities and found that roads of all kinds 

have seven general effects: mortality from road construction, mortality from collision with vehicles, 

modification of animal behavior, alteration of the physical environment, alteration of the chemical 

environment, spread of exotics, and increased use of areas by humans. Roads affect amphibian 

populations most obviously by eliminating or altering migration and dispersal corridors, but also 

directly through road mortality or simply by damaging habitat (Smith and Keinath, 2007). 

Herpetologists are well aware of the negative affect of roads on amphibians because they frequently use 

roads to survey their target organisms (Shaffer and Juterbock, 1994). Roads located too close to 

wetlands have been known to reduce populations of salamanders (Semlitsch, 1998). It has been known 

for some time that roads cause extensive mortality of juvenile NLFs (Bovbjerg, 1965; Bovbjerg and 

Bovbjerg, 1964; Merrell, 1977). Ashley and Robinson (1996) found that young-of-the-year NLFs were 

disproportionately represented among dead herpetofauna on roads at their study site in Ontario. Road 

runoff can also be a source of environmental pollution, and can significantly threaten nearby aquatic 

habitats (Welsh and Oliver, 1998; Trombulak and Frissell, 2000; Dorchin and Shanas, 2010). NLFs are 

more vagile than other species of frogs and this could be a disadvantage in landscapes with roads 

because they are more likely to encounter roads and incur traffic mortality (Carr and Fahrig, 2001). In 

Ontario, NLF population density was found be negatively affected by traffic density within a radius of 

1.5 km whereas compared with a less vagile frog, the green frog (Rana clamitans), there was no 

evidence that the presence of vehicular traffic affected their populations (Carr and Fahrig, 2001). 

Climate factors, such as drought, also can have a major influence on NLF populations (Merrell, 

1977; Corn and Fogleman, 1984; Koch and others, 1996). In drought years, habitat for frogs will decline 

and local populations may be extirpated. Drought conditions were associated with a decline in 
142

amphibian populations in southern Alberta in the 1930s (Fowler, 1935). Drought conditions were also 

prevalent in southern Alberta in the late 1970s and 1980s, suggesting a possible link between climatic 

conditions and the abrupt decline of this species in Alberta beginning in the late 1970s (Alberta 

Northern Leopard Frog Recovery Team, 2011). There is clear evidence that the number of Growing 

Degree Days (GDD) have increased significantly in the SLV, indicating climate is changing 

disproportionately during the growing season (Mix and others, 2010). This could potentially lead to 

more drought conditions in the SLV in the future, which would negatively affect NLF populations. 

Increasing ultraviolet (UV) radiation from a thinning ozone layer has also been linked with 

declines in amphibian populations and with increased incidence of limb malformations (Crump and 

others, 2000; Ankley and others, 2002; Blaustein and others, 2003). Embryo mortality has been 

attributed to UV radiation in the genus Rana (Blaustein and others, 2003). NLFs lay their eggs close to 

the water surface, which could potentially increase their susceptibility to UV radiation. UV radiation 

may also reduce resistance to disease and pathogens. UV radiation and contaminants may interact with 

one another synergistically and enhance the detrimental effects to amphibians (Blaustein and others, 

2003). Bridges and others (2004) found that the presence of the contaminants atrazine and carbaryl, and 

the interaction with UV radiation increased the incidence of hatchling deformities. Ankley and others 

(1998) examined the effects of UV light and the insecticide methoprene on the survival and 

development of NLFs. They were able to induce relatively specific hindlimb malformations in frogs by 

treatment with UV light, but that methoprene treatments both in the absence and presence of UV, did 

not result in limb abnormalities. Croteau and others (2008) also looked at exposure of NLF tadpoles to 

UV-B radiation and an estrogenic chemical 4-tert-octylphenol (OP), both alone and in combination. 

UV-B radiation alone was sufficient to produce mortality of tadpoles, and exposure to UV-B radiation 

or OP alone induced significant increase in malformations and abnormalities of tadpoles (Croteau and 

others, 2008). 

Oil and gas development threatens the NLF and its habitat (Center for Native Ecosystems, 

2006). Federal land management agencies, including the Bureau of Land Management (BLM) and the 

U.S. Forest Service (USFS), have determined that oil and gas development in areas of Colorado, 

Montana, New Mexico, and Wyoming, including the drilling and maintenance of wells, related 

construction of roads, and disposal of wastes, will negatively affect the NLF (Center for Native 

Ecosystems, 2006). The USFS concluded that coal-bed methane development in Wyoming will 

negatively affect the NLF and its habitat in several areas of Thunder Basin National Grassland (USFS 

2004a; 2004b; 2004c). Impacts associated with coal-bed methane development include road-related 

mortality, discharge of contaminated water into breeding ponds, the loss of spring flows due to massive 

groundwater removal, discharge of excessively cold water into breeding ponds, and discharge of water 

that may facilitate the introduction of predatory fish (Powder River Basin Resource Council, 2002). In 

general, coal-bed methane development, in which massive amounts of groundwater are removed and 

later discharged at the surface, is detrimental to aquatic systems that may support the NLF and has 

negatively affected aquatic systems supporting the western United States population of NLF (Powder 

River Basin Resource Council, 2002). Other areas where oil and gas development is reported to be 

negatively affecting or threatening to negatively affect the western population of NLF and its habitat 

include La Plata and Archuleta Counties in southwestern Colorado, Big Horn, Blaime, Carbon, Custer, 

Gallatin, Musselshell, Park, Powder River, Rosebud, Stillwater, Sweetgrass, Treasure, and Yellowstone 

Counties in Montana, and Carbon County, Wyoming (Center for Native Ecosystems, 2006). Oil and gas 

development in these areas will adversely affect the NLF and its habitat by directly disturbing the 

species and its habitat, indirectly altering local hydrology, by introducing contaminants into waters, and 

by reducing the availability of water for the species. Oil field development may also be contaminating 
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NLF habitats in Wyoming (Ramirez, 2002). A majority of oil field discharge samples were found to 

exceed the state of Wyoming water quality standards for oil and grease. In Colorado, oil and gas 

development also threatens NLF populations in the Roan Plateau region (Center for Native Ecosystems, 

2006).

Agriculture strongly dominates the landscapes of some regions of the United States, including 

the SLV (Mix and others, 2010). Agricultural practices have a potentially large negative influence on 

amphibian populations because of the attendant habitat loss, isolation, and chemical and nutrient 

contamination (Bishop and others, 1999; Kolozsvary and Swihart, 1999; Zampella and Bunnell, 2000; 

Joly and others, 2001). Chemical and nutrient contamination is a serious threat to NLFs and is discussed 

in more detail under the “Water Quality” section below. However, in regions where natural wetlands are 

scarce, constructed agricultural ponds may represent important alternative breeding habitats for 

amphibians (Baker and Halliday, 1999). If properly managed, agricultural ponds may effectively 

increase the total amount of breeding habitat in a region and help to sustain populations (Meyer-Aurich 

and others, 1998; Pechmann and others, 2001). 

Nonindigenous Biotic Agents

The effects of introduced organisms on amphibians has been extensively documented. The 

potential negative effects of introduced predaceous fish and bullfrogs on NLF populations have already 

been discussed in a previous section. The impact of diseases to NLFs is also relevant to this review. 

Chytridiomycosis, caused by the bacterium Batrachochytrium dendrobatidis (Bd), has been recognized 

as a threat to the persistence of amphibian populations and is causing population declines in many frog 

species (Berger and others, 1998; Rollins-Smith and others, 2002; Longcore and others, 2007). Chytrid 

fungi are ubiquitous and found in aquatic and moist habitats where they degrade cellulose, chitin, and 

keratin. In amphibians, the chytrid fungi degrades the keratin component of the skin, leading to 

excessive keratin production and sloughing of the skin (Berger and others, 1999), which can impair 

cutaneous respiration and osmoregulation and result in death (Berger and others, 1998). There appears 

to be considerable variation among and within species in their susceptibility to the disease (Blaustein 

and others, 2005; Daszak and others, 2004). Voordouw and others (2010) studied the prevalence of Bd

in an endangered population of NLFs in British Columbia and concluded that prevalence appeared to 

have stabilized in the last three years of the study, young of the year emerging from breeding ponds 

were rarely infected with Bd, and some individuals cleared their Bd infection. These findings indicated 

that NLFs may have evolved a resistance to chytrid-related mortality in this particular location. 

Generally, ranid frogs should have a natural deterrent to chytrid infection in natural antimicrobial 

peptide defenses (Rollins-Smith and others, 2002). However, environmental factors, such as cold 

temperatures, UV-B or toxic chemicals may inhibit peptide synthesis and release, thereby allowing the 

chytrid fungus to survive and eventually kill amphibians (Rollins-Smith and others, 2002). 

Infections by certain bacteria (most commonly Aeromonas hydrophilia) results in a condition 

referred to as “red-leg” that is nearly always fatal. This infection affects the kidneys and other organs of 

frogs and appears to be triggered by factors such as stress, injury, malnourishment or environmental 

conditions such as poor water quality (Nace and others 1996; Kendell, 2002a). Mortality and apparent 

population declines in NLFs have previously been associated with “red-leg” in North America (Gibbs 

and others, 1971; Hine and others, 1981; Koonz, 1992). 

Viral infections such as Ranavirus and Herpesvirus have also been noted in NLFs and may 

contribute to mortality of adults or embryos (McKinnell, 1973; Hunter and others, 1989; Green and 

others, 2002; Echaubard and others, 2010). In a review of 20 amphibian morbidity events from 1996-

2001, Green and others (2002) found that the most common cause of amphibian mortality was infection 
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by Ranaviruses. Mortality events from Ranaviruses affected predominately larval amphibians, were 

observed primarily in late spring and summer, had a sudden onset, and casualties numbered in the scores 

to thousands of amphibians daily (Green and others, 2002). Echaubard and others (2010) found density 

was a primary factor determining the direction of the interaction between NLF tadpoles and Ranavirus. 

They found higher mortality, earlier death rates, longer development rates, and growth rates 

significantly lower in high-density situations when exposed to Ranaviruses. 

Declines in amphibians may be due to a variety of environmental factors, not just the disease 

pathogen. For example, Brodkin and others (1992) found that a naturally occurring bacteria, 

Pseudomonas aeruginosa, was pathogenic only for NLFs under suboptimal environmental conditions. 

Their results suggested that there was an interaction between crowding of individuals, temperature, and 

bacteria on mortality. It is also hypothesized that multiple environmental stressors (for example, 

temperature extremes, increase UV radiation, exposure to heavy metals and toxins) may be 

synergistically compromising the response of amphibian immune systems (Carey and others, 1999). 

Water Quality

Amphibians are particularly susceptible to ecotoxicological agents because of their complex life 

cycle and the permeability of their skin (Cooke, 1981; Duellman and Trueb, 1986; Bishop, 1992; Hall 

and Henry, 1992). There has been extensive work on the effects of water quality on NLFs. A simple 

Google Scholar literature search revealed more than 1,700 citations with the keywords NLF and water 

quality. I focus the summary of this literature on those hazards that appear to be of most concern in the 

Rocky Mountain region and the SLV complex of Refuges. I list below the 10 main factors contributing 

to poor water quality in wetland habitats frequented by NLFs (Smith and Keinath, 2007): 

(1) Pesticides

Pesticides are the most studied classes of ecotoxicological agents that can affect amphibians, and 

specifically NLFs (Sparling and others, 2010). Various pesticides are used across the Rocky Mountain 

region including 2,4-D Amine, Escort®, Plateau®, and Roundup® to name a few. Many of the most 

common pesticides used currently are supposed to be nonpersistent and have short half-lives, usually 

from one week to 30 days (Wauchope and others, 1992), unlike more persistent pesticides used in the 

past (for example, organochlorines). However, year-round detections of low pesticide concentrations in 

streams draining from agricultural and urban areas in Colorado could indicate that these compounds 

persist in the shallow alluvial aquifer year-round (Kimbrough and Litke, 1996). Pesticides have been 

shown to cause direct mortality, hermaphrodism, limb malformations, lower survival rates, alter 

behavior of tadpoles, disrupt hormone production and impair reproductive potential, and suppress 

immune systems. Below I discuss a selection of studies that have examined pesticides used and studied 

in relation to NLFs in various parts of North America. 

• Berrill and others (1993; 1994; 1997) studied various pesticides commonly used in agriculture 

and silviculture in Canada. They found that the pyrethroid insecticides permethrin and 

fenvalerate did not cause significant mortality in NLF eggs or tadpoles exposed to low levels of 

these insecticides (Berrill and others, 1993). However, tadpole growth was delayed following 

exposure, and tadpole larvae twisted abnormally instead of darting away when being prodded by 

investigators. Both of these results indicated a greater vulnerability to predation. Berrill and

others (1994) also studied the effects of pesticides used to manage coniferous forests in Canada 

on NLF embryos and tadpoles. They exposed embryos and tadpoles to low levels of the 

insecticide fenitrothion and the herbicides triclopyr and hexazinone. None of these pesticides 

affected hatching success or subsequent behavior in tadpoles that were exposed. However, newly 
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hatched tadpoles were very sensitive to the concentrations used in the study (2.4 to 4.8 ppm 

triclopyr and 4.0 to 8.0 ppm fenitrothion), exhibiting paralysis or death (Berrill and others, 

1994). NLFs were exposed to low levels of insecticides permethrin, fenvalerate, and fenitrothion 

and the herbicides hexazinone, triclopyr, glyphosate, bromoxynil, triallate, and trifluralin, all of 

which are used in the management of coniferous forests in Canada (Berrill and others, 1997). 

Embryos were unaffected by the exposures, but newly hatched tadpoles were paralyzed or killed. 

Of those listed herbicides, glyphosate is the active ingredient in Roundup® and is the most 

widely used herbicide in the United States (Grube and others, 2011). Relyea (2005) also 

examined the impact of Roundup® on the survival of tadpoles and found that when using 

realistic exposure times (longer than a few days) combined with the frequently occurring stress 

of predators found in natural communities, Roundup® had the potential to kill many species of 

amphibians, including the NLF. 

• Harris and others (1998) examined the effects of a combination of six pesticides commonly 

applied in apple orchards to premetamorphic NLFs. They found that Basudin® and Dithane® at 

environmentally relevant concentrations (<0.01 mg/L) caused mortalities, deformities, or growth 

inhibition during early development, whereas Imidan®, Guthion®, and Nova® required higher 

concentrations to produce survival or growth effects (5-10 mg/L). 

• Ouellet and others (1997) sampled NLFs in farm ponds and nonfarm ponds in the St. Lawrence 

River Valley of Quebec, Canada, and found a large number of hind limb deformities in frogs 

from farmland ponds. They blamed pesticides, but no causal link was established. 

• Corn and soybean agricultural sites in southwestern Ontario rely heavily on pesticide and 

nutrient inputs to maximize production (McDaniel and others, 2008). In these agricultural areas, 

up to 75 percent of male NLFs were found to have some degree of testicular abnormalities 

(McDaniel and others, 2008). In a few males the severity was such that no differentiated 

testicular tissue was present, in which case reproductive potential of these individuals was likely 

impaired. A variety of pesticides, particularly herbicides, and nutrients were detected in the 

water of these agricultural sites, including atrazine and metolachlor. 

• Atrazine disrupts endocrine function and, even at concentrations as low as 0.1 ppb, causes 

retarded gonadal development, hermaphroditism, and oocyte growth in male NLFs (Hayes and 

others, 2002a; 2002b; 2003). Atrazine contamination is reported to be negatively affecting NLFs 

from Utah to Iowa (Hayes and others, 2002a; 2002b). Hayes and others (2002b) specifically 

noted that atrazine contamination has negatively affected NLFs in the Platte River drainage of 

southeastern Wyoming. Most water sources in the United States, including rain, contain more 

atrazine than the effective dose determined in laboratory studies (Hayes and others, 2002b). 

• The pesticide endosulfan was found to be highly toxic to nine species of amphibians, including 

NLFs in laboratory experiments (Jones and others, 2009). 

• Malathion can threaten NLFs throughout the western United States (Fordham, 1999). Malathion 

is one of the most commonly used organophosphate insecticides, and is registered for residential 

and agricultural purposes (US EPA, 1998). Fordham (1999) reported malathion use in the SLV, 

Colorado, could negatively affect the NLF by altering avoidance behavior, decreasing overall 

survival, and negatively affect tadpole growth and development. 

• Organochloride pesticides (for example, DDE, DDT, and Dieldrin) were banned in the 1970s 

and 1980s, but are environmentally persistent compounds (Diana and Beasley, 1998). They 

continue to negatively affect the NLF in the western United States (Smith and Keinath, 2007). 

• Pesticides have been shown to compromise the immune system of NLFs (Gilbertson and others, 

2003; Christian and others, 2004; Gendron and others, 2003). Christin and others (2004) 
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exposed NLFs to a mixture of pesticides known to be present in surface waters of many 

tributaries of the St. Lawrence River (Quebec, Canada). The mixture included atrazine, 

metribuzine, endosulfan, lindane, aldicarb, and dieldrin. They found that mixture of pesticides 

altered the lymphocyte proliferation of NLFs, thereby altering their immune response. They 

concluded that their altered immune response could be a contributing factor to their global 

decline by rendering them more susceptible to certain infections (Christin and others, 2004). 

Gilbertson and others (2003) conducted an injection study and a field study to investigate NLFs 

immune function response to sublethal doses of DDT, malathion, and dieldrin. They found that 

antibody response was dramatically suppressed in all of the pesticide-injected groups. The field 

component revealed significant differences in the immune function between NLF populations in 

pesticide-exposed and pesticide-free locations, with suppressed immunity in pesticide-exposed 

sites (Gilbertson and others, 2003). Exposure of NLFs to a pesticide mixture has been shown to 

affect the infection dynamics of a common parasite of ranid frogs, the lungworm Rhabdias

ranae (Gendron and others, 2003). Gendron and others (?) found that migration of the lungworm 

was significantly accelerated in hosts exposed to the highest concentrations of pesticides, leading 

to the establishment of twice as many adult worms in the lungs of frogs 21 days post-infection. 

They concluded that certain components of the frog immune response were significantly 

suppressed after exposure to the pesticide mixture. 

(2) Fertilizers

Fertilizers have been found to be toxic to amphibians. Rouse and others (1999) found that nitrate 

levels in many agricultural ecosystems in North America exceeded 1 mg/L (for example, concentrations 

that are toxic to amphibians and other aquatic organisms), and concluded that it was highly probable 

that nitrate concentrations in surface waters in North America are adversely affecting amphibian 

survival. Hecnar (1995) examined the effects of fertilizers on NLFs (Hecnar, 1995). In this study, NLF

tadpoles were exposed to chronic and acute doses of ammonium nitrate fertilizer. In acute doses, 

tadpoles suffered severe weight loss. In chronic doses, tadpoles not only lost weight, but also died, even 

at fairly low dosages. The NLF was the most severely affected of the species tested, which included 

American toads (Bufo americanus), chorus frogs (Pseudacris triseriata), and green frogs (Rana

clamitans). Toxic effects of ammonium nitrate occurred in all four species at concentrations that are 

commonly exceeded in agricultural areas globally. Hecnar (1995) emphasized that the differential 

mortality of the species tested would likely cause shifts in species composition in free-living 

communities of amphibians. Runoff from croplands could potentially affect Refuges in the SLV. 

Because application of the herbicide atrazine coincides both spatially and temporally with the 

application of fertilizer, Allran and Karasov (2000) examined the effects of atrazine and nitrate on NLF

larvae in the laboratory. They found that atrazine, nitrate, and their interaction had no significant effect 

on development rate, percent metamorphosis, time to metamorphosis, percent survival, mass at 

metamorphosis, or hematocrit. They did find that nitrate slowed the growth of larvae. 

(3) Mining & Heavy Metals

Mining has been practiced in the Rocky Mountain area for at least 150 years (Smith and 

Keinath, 2007). Although mining has become less important and less destructive in recent years, residue 

from abandoned mines continues to adversely affect many drainages. Mining causes acidification of 

water and leads to metals-laden effluent from mines and smelter sites that is discharged into receiving 

waters. Porter and Hakanson (1976) found that an entire drainage in Colorado was devoid of amphibian 

life due to runoff from mining residue almost 80 years after mining had ceased. Heavy metals such as 
147

aluminum, lead, zinc, cadmium, mercury, silver, copper, arsenic, manganese, molybdenum, and 

antimony have a number of effects on amphibians (Blaustein and others, 2003). These metals can be 

directly lethal or induce sublethal effects such as slowing growth and development and altering 

behavior. 

Arsenic is frequently released to surface waters from mining, smelting, coal-burning power 

plants, industries, urban runoff, application of biosolids, fertilizers and pesticides, and atmospheric 

deposition (ATSDR, 2007; Irving and others, 2008). Arsenic was used in fertilizers in the early 1900s 

and is a byproduct of the production of sulfuric acid (Smith and Keinath, 2007). Production of arsenic in 

most nations, including the United States, ceased around the turn of the 20th century, but arsenic can still 

be a significant pollutant in some parts of the United States (Aurilio and others, 1995). Birge and Just 

(cited in Linder and Grillitsch, 1973) found that high levels of sodium arsenite were lethal to NLF

tadpoles. Chen and others (2009) exposed NLF tadpoles to environmentally relevant concentrations of 

arsenic and found no significant effect of arsenic exposure on tadpole survival, growth, and 

metamorphosis. However, they found that tadpoles had decreased swimming speeds in arsenic-treated 

groups as opposed to control groups. 

Lead has the potential to enter waterways via discharge from sewage treatment plants, industries, 

mining, and highway runoff (Chen and others, 2006). Chen and others (2006) exposed NLFs to

sublethal and environmentally relevant concentrations of lead from embryonic stage to metamorphosis. 

Tadpole growth was significantly slower in the early stages for all doses, and more than 90 percent of 

the tadpoles in the highest dosage tested in the experiments developed lateral spinal curvature. Spinal 

deformity was associated with a slower swimming speed. These same authors also studied the effects of 

copper and cadmium on NLFs and found that both heavy metals affected development, growth, 

survival, and behavior in larval NLFs at ecologically relevant concentrations (Chen and others, 2007; 

Gross and others, 2007). Both copper and cadmium occur naturally in soil and is ubiquitous in the 

environment; however, elevated levels of copper in surface water can occur from mining operations, 

industrial development, waste water treatments, and agricultural runoff (Chen and others, 2007). 

(4) pH/Acidification

Acidification may contribute to local extinctions of amphibians (Blaustein and others, 2003). For 

example, Harte and Hoffman (1989) suggested that episodic acidification might have contributed to 

declines in tiger salamander (Ambystoma tigrinum) populations in the Rocky Mountains, Colorado. The 

effects of heavy metals on amphibian survival are linked closely to acidification because heavy metals 

may leach from soil in contact with acidic water (Blaustein and others, 2003). For example, aluminum 

solubility increases as water pH drops and aluminum often acts synergistically with pH to cause embryo 

mortality in amphibians (Clark and Hall, 1985; Clark and LaZerte, 1985; Freda and McDonald, 1990; 

Freda and others, 1990). 

The effects of acidification of water on frogs has been studied intensively due to the potential for 

acid rain to cause amphibian declines (Smith and Keinath, 2007). NLF eggs cannot develop normally at 

pH 5.8 or lower (Schlichter, 1981), and tadpoles were amongst the most sensitive to low pH of all 

species tested so far (Freda and Taylor, 1992). NLF sperm also showed decreased motility at lower pH, 

with 50 percent of normal motility at pH 5.5 and maximum motility at pH above 6.5 (Schlichter, 1981). 

NLFs also show chronic effects of decreased pH, with increasing mortality over time when exposed to 

low pH waters (Freda and Dunson, 1985; Vatnick and others, 1999). Low pH can also act 

synergistically with higher levels of UV-B radiation that might result from loss of the ozone layer, 

resulting in higher mortality in amphibians (Long and others, 1995). Exposure to chronically low levels 
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of pH (5.5) also caused a decrease in white blood cell number and viability in NLFs, which led to 

systemic bacterial infections, and increased mortality (Simon and others, 2002). 

Acidification of water in the Rocky Mountain region can result from acid rain (from air 

pollution) or mining. Porter and Hakanson (1976) showed that acidification of water locally due to point 

sources, such as old mines, can cause amphibian mortality. However, Corn and Vertucci (1992) studied 

the effects of acidic deposition on six species of amphibians, including the NLF, in the mountains of 

Colorado and Wyoming. They concluded that acid deposition poses little risk regionally to amphibians 

in the Rocky Mountains. In a follow-up study, Vertucci and Corn (1996) concluded that NLF declines 

in Colorado were not likely to have been caused by episodic acidification because NLFs generally 

occupy lakes that have a high acid-neutralizing capability.

(5) Motorized T ransportation

Roads can cause substantial direct mortality as previously mentioned as a landscape-level 

process. They can also have indirect impacts on NLFs as a consequence of sedimentation and toxic 

runoff during rain and snowmelt (Smith and Keinath, 2007). Motor oil washes off roads during rain and 

snowmelt runoff, and when suspended in water can kill some amphibians (Sparling and others, 2000). 

Crankcase oil in runoff from roads may inhibit amphibian growth (Diana and Beasley, 1998). 

(6) Cattle Grazing

Domestic livestock grazing is one of the most widespread land uses in North America, with 

more than 70 percent of the western United States reported to be grazed (Maxell, 2000). Livestock, and 

in particular cattle, produce considerable amounts of waste products that are incorporated in runoff to 

waterways (Smith, 2003). Directly, livestock grazing has caused mortality to NLFs in Montana (Maxell, 

2000). Trampling of amphibians by domestic livestock can result in severe population-level impacts 

(Bartelt, 1998; Maxell, 2000; Smith, 2003). Trampling of overwintering habitats by domestic livestock 

may directly or indirectly lead to mortality of vulnerable overwintering NLFs (Maxell, 2000; Smith, 

2003). Indirectly, domestic livestock grazing can cause changes in water quality that significantly 

negatively affect the NLF and its habitat (Maxell, 2000; Smith, 2003). Livestock defecation can increase 

fecal coliform counts in runoff, in turn increasing phosphorous and nitrogen concentrations and 

contributing to eutrophication of waters, which can deplete oxygen needed for tadpoles or overwintering 

adults (Wetzel, 1983; Maxell, 2000). Bacterial pollution caused by defecating livestock can also lead to 

increases in Planorbid snails, which are carriers of trematode parasites (Ribeiroia spp.), thereby 

increasing the risk of parasitic infection that can lead to limb malformations and mortality (Johnson and 

others, 1999; Johnson and Lunde, 2005). Knutson and others (2004) found that ponds used for watering 

livestock had elevated concentrations of phosphorus, higher turbidity, and a trend toward reduced 

amphibian reproductive success. 

Grazing can adversely alter soils and vegetation and indirectly affect the NLF and its habitat. By 

trampling banks and vegetation, livestock can eliminate emergent vegetation alongside breeding ponds 

(Kauffman and Krueger, 1984; Fleischner, 1994; Belskey and others, 1999). Livestock can also reduce 

or eliminate riparian habitats that are used as dispersal corridors for adults and subadult frogs 

(Fleischner, 1994; Maxell, 2000; Smith, 2003; Mazerolle and Desrochers, 2005). The elimination of 

vegetation from grazing can lead to increases in water temperatures (Belskey and others, 1999), which 

can negatively affect the NLF by making aquatic habitats more hospitable to nonnative bullfrogs 

(Smith, 2003). 
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(7) Sedimentation

Sediment can run into waterways due to erosion caused by a variety of sources, including roads, 

domestic livestock grazing, and vegetation removal (Smith, 2003). Sediment can cover NLF eggs in 

water and probably inhibits gas exchange by eggs. Large fires can also contribute to sedimentation of 

waterways by producing a lot of wood ash and dead vegetation, thereby increasing the potential for 

erosion. 

(8) Rotenone

Rotenone is commonly used in fisheries management and can have negative effects on NLFs 

(Hamilton, 1941; Diana and Beasley, 1998, Patla, 2005). NLF tadpoles cannot survive levels of 

rotenone typically used to sample fish (Hamilton, 1941). 

(9) Polychlorinated biphenyls (PCBs)

A large amount of literature on PCBs and their effects on amphibians is summarized by Sparling 

(2000). Although the manufacture and use of PCBs were banned decades ago, these chemicals are still a 

concern in many areas because of their slow degradation rates and demonstrated ability to undergo 

food-web biomagnification (Zaranko and others, 1997). These compounds continue to threaten 

amphibians, including the NLF, throughout the United States (Diana and Beasley, 1998). Phaneuf and 

others (1995) studied PCB contamination in NLFs in Canada following a fire at a PCB warehouse. They 

found elevated concentrations in frogs in the area, but were not able to assess mortality in frogs by 

PCBs released by the fire. In Wisconsin, high concentrations of PCB 126 caused significant mortality 

on NLF tadpoles in the laboratory (Jofre and others, 2000). However, these authors found no negative 

effect at concentrations of PCB 126 found in the wild. Karasov and others (2005) studied exposed frog 

embryos and tadpoles of NLFs along a pollution gradient in the Fox River/Green Bay ecosystem in 

Wisconsin. They found that hatchability of frog embryo batches and species richness at sites were 

negatively correlated with contamination levels, but tadpole growth and survival were not. Adult NLFs 

were found to have low elimination rates for PCBs (Leney and others, 2006). Adults in contaminated 

areas would likely continue to bioaccumulate PCBs, and therefore could potentially be used as 

biomonitors for these persistent organic chemicals. 

(10) Other Contamination and L imb Malformation

Organic compounds such as caffeine, acetaminophen, and triclosan in wastewater can have both 

behavioral and physiological effects on NLF tadpoles (Fraker and Smith, 2004). Both triclosan and 

caffeine separately, as well as the significant interaction between caffeine and acetaminophen, affected 

tadpole activity levels, even at the very low concentrations at which the compounds have been detected 

in the environment. The effects were nonlethal; however, the traits studied in this particular case may 

have significant impacts on tadpole performance and success. Wastewater effluents are complex 

mixtures containing a variety of anthropogenic compounds, many of which are known endocrine 

disruptors. Sowers and others (2009) exposed NLFs to a range of concentrations of municipal 

wastewater effluent from the egg stage through metamorphosis. At treatments of 50 percent and 100 

percent wastewater, individuals took significantly longer to reach metamorphosis than individuals in 0 

percent and 10 percent wastewater. Male frogs also had an increased incidence of testicular oocytes in 

the 50 percent and 100 percent treatments. Study results suggested that municipal wastewater effluent 

can alter the timing of the metamorphic process and impact male sexual development in NLFs (Sowers 

and others, 2009). 
150

Fluoxetine (Prozac) is a selective serotonin reuptake inhibitor (SSRI) commonly found in 

aquatic ecosystems and organic wastewater. NLF tadpoles exposed to fluoxetine at both high and low 

concentrations exhibited delays in development and gained less weight than control tadpoles (Foster and 

others, 2010). Delayed development could put tadpoles at greater risk of desiccation and overcrowding 

as temporal ponds dry, as well as size-specific predation.

Perfluorooctanesulfonates (PFOS) represent a class of chemicals produced in large amounts over 

the past two decades for uses such as protective coatings for fabrics, greases and lubricants, adhesives, 

firefighting foams, and insecticides (Ankley and others, 2004). Monitoring studies found that there were 

several PFOS extensively distributed in both humans and the environment (Giesy and Kannan, 2001; 

Ankley and others, 2004). PFOS was lethal to NLFs at concentrations of 10 mg/L and caused 

developmental delays in tadpoles at lower concentrations (Ankley and others, 2004). 

Limb malformations in amphibians are known from forty states and four Canadian provinces 

(Northern Prairie Wildlife Research Center, 1997). The most famous recent reports were those from 

Minnesota, which virtually all concerned metamorphic NLFs (Helgen and others, 1998; 

Vandenlangenberg and others, 2003). Malformed frogs usually do not survive long because locomotion 

and basic behavior are seriously compromised by these gross malformations (Merrell, 1969). Currently, 

three major environmental factors are identified as the causes of malformations: contaminants, parasites, 

and injuries from predators. The trematode parasite Ribeiroia ondatrae has been responsible for limb 

malformations and mortality among NLF populations throughout the western United States, including 

Minnesota (Schotthoefer and others, 2003; Johnson and Lunde, 2005; Sutherland, 2005). This trematode 

can cause significant levels of mortality among NLF frog tadpoles (Schotthoefer and others, 2003). 

Parasite-induced limb malformations are reported to be on the rise in the western United States, and has 

been attributed to extensive human alterations to aquatic systems (Johnson and Lunde, 2005). 

Researchers suggest multiple causes are likely to blame for worldwide reports of amphibian 

malformations, and factors leading to malformations at one site may differ from causes at another site 

(Meteyer and others, 2000. 

Intrinsic Threats

The NLF is intrinsically vulnerable from disturbance factors for a number of reasons. NLFs use 

small (usually <5 ha) ponds in which to breed (Merrell, 1968; 1977, Collins and Wilbur, 1978; Corn and 

Livo, 1989; Hammerson, 1999). Ponds of this size are often not protected by law (Semlitsch, 2000b) 

and typically are not addressed in management plans. Smith and Keinath (2007) suggested that it is 

imperative that management plans take into consideration the distribution of small ponds across the 

landscape and protection of these ponds to foster healthy populations of NLFs. NLFs also selectively 

breed in small, fishless ponds, and the introduction of predatory fish to these ponds is a well-known 

threat to this species. Federal and state agencies frequently introduce such fish throughout the range of 

the NLF. NLFs also use moist upland habitats surrounding their breeding ponds for summertime 

foraging. Semlitsch (1998) determined that an upland habitat of up to 164 m surrounding each breeding 

pond might be sufficient to protect 95 percent of the adult population of salamanders of various species 

that use a pond for reproduction in the spring. However, this type of research has not been done for 

NLFs and it is a critical information need. Amphibians have highly permeable skin, which allows for 

toxins to be readily absorbed through the skin and incorporated into the body of amphibians, including 

NLFs. The accumulation of toxins in body tissues of NLFs can have repercussions throughout the food 

web. Smith and Keinath (2007) suggested that NLFs might serve as a key indicator species in 

ecosystems in which they occur and that it was possible that NLFs might indicate overall ecosystem 

health. NLFs are also susceptible to a variety of diseases as mentioned in the previous section. 
151

Ranavirus has probably always been a threat to NLF populations (Jancovich and others, 1997; Carey 

and others, 1999, Daszak and other, 1999). Finally, an additional intrinsic threat to NLF populations is 

their need to move across the landscape for many reasons, including dispersal of metamorphs (Dole, 

1971; Merrell, 1977), summer movements associated with feeding (Dole, 1965b; 1967; Merrell, 1977), 

and migrations to and from overwintering sites (Dole, 1967; Merrell, 1977). The routes followed 

probably include wet meadows, tall grass, and riparian corridors. 

Management 

Semlitsch (2000a) suggested three critical factors to consider in any management plan for 

aquatic-breeding amphibians: (1) the number or density of the individuals dispersing from individual 

wetlands (local population dynamics); (2) the diversity of wetlands with regard to hydroperiod 

(community dynamics); and, (3) the probability of dispersal among adjacent wetlands or the rescue and 

recolonization of local populations (metapopulation dynamics). Semlitsch (2000a) suggested that it is 

critical to maintain natural terrestrial habitats peripheral to wetlands to enable amphibians to carry out 

all life-history functions. Management plans should also consider restoration projects where habitat is 

degraded or where natural processes have been lost or averted. Small wetland restoration projects were 

found to be beneficial to NLF populations on Prince Edward Island, Canada (Stevens and others, 2002). 

Activities near breeding areas that should be avoided include intensive timber harvest, agriculture 

(including the periodic planting of food plots for game species or mowing), and construction of 

buildings. If these activities are necessary, buffer zones should be maintained around wetlands or along 

streams to prevent sedimentation and soil compaction, and removal of coarse woody debris should be 

minimized (de-Maynadier and Hunter, 1995). Minimizing construction of roads or motorized vehicle 

traffic near breeding sites was also suggested by Semlitsch (2000a). The use of agricultural and other 

chemicals near aquatic habitats should be restricted; for example, herbicide use along roads for weed 

control, salt on roads to eliminate ice, fire retardants, growth retardants on vegetation along roads or 

under powerlines, and pesticide use to control insects, especially near pools, ditches, streams, and ponds 

where runoff would find its way into wetlands. Sedimentation in aquatic habitats from permanent roads 

or temporary logging roads should be minimized to prevent harmful effects on eggs, embryos, and 

gilled larvae. Wetland management practices that drawdown marshes (also lowering the water table for 

ephemeral pools) at appropriate times to enhance waterfowl production may decrease the probability of 

metamorphosis of amphibian species. Flooding marshes (also raising the watertable) for waterfowl may 

facilitate the colonization of ephemeral pools by predatory fish. 

Managers should also consider responses to natural processes such as floods, fires, or succession 

(Semlitsch, 2000a). Any of these processes can be beneficial or harmful depending on the species, the 

region, and how the natural disturbance regime interacts with anthropogenic changes to the landscape. 

Management plans should also eliminate avenues of colonization by predatory fish such as drainage 

ditches that connect fish-inhabited areas (streams or rivers) with amphibian breeding ponds. Elimination 

of other invasive species such as exotic fish and bullfrog larvae would serve to promote successful 

metamorphosis of native species. Plans that maintain natural vegetation buffer zones around ponds 

(recommended 164 m from the edge of wetlands; Semlitsch, 1998) and along streams (recommended 

width 30-100 m but adjusted for stream width, slope, and site use; Rudolph and Dickson, 1990; 

McComb and others, 1993; deMaynadier and Hunter, 1995) will increase the probability of species 

persistence. 

A selection of citations that provided specific management recommendations for conservation of 

NLF populations is described below: 
152

• Hammerson (1999) made several suggestions in relation to management specific to NLFs. He 

suggested that management for NLFs is probably best accomplished through the creation and 

protection of habitats that favor leopard frog reproduction but limit or prohibit the successful 

breeding by bullfrogs and predatory fishes (Hammerson, 1999). Specifically, periodic 

dewatering or large water-level reductions in late summer after NLFs have metamorphosed 

might be beneficial. Establishment of semipermanent wetlands among permanent wetlands may 

allow the persistence of NLFs in areas where bullfrogs and exotic predatory fishes cannot be 

eliminated without unacceptable damage to other wetland resources. Refuges for NLFs could be 

established in isolated bodies of water, such as foothill ponds, far from permanent deepwater 

habitats inhabited by bullfrogs and fishes (Hammerson, 1999). 

• In Washington, site occupancy models at a 1-km2 scale indicated that occupied areas for NLFs 

had greater average pond depths, fewer ponds occupied by bullfrogs and carp, and more 

herbaceous vegetation surrounding the ponds (Germain and Hays, 2009). 

• Knutson and others (2005) suggested limiting livestock access to ponds to improve water 

quality, reducing nitrogen input, and avoiding the introduction of fish. 

• There have been several studies on NLFs in managed forest habitats (Bull and Wales, 2001; 

Blomquist and Hunter, 2009). Blomquist and Hunter (2009) studied NLF habitat selection in a 

managed forest in Maine. They radiotracked 40 adult NLFs and found that they selected activity 

centers in the late spring and summer that focused on open-canopy areas with ample moisture 

and standing water. The frogs also used unharvested forest for longer distance movements. Bull 

and Wales (2001) reviewed the possible effects of disturbance from various forest management 

strategies on several species of amphibians, including NLFs, in eastern Oregon and Washington. 

The management strategies considered included thinning of vegetation, prescribed burns, road 

removal, and spraying with pesticides and biological microbial agents. They concluded that there 

should be no negative effects on NLFs from management strategies associated with forest health 

unless spraying with pesticides decreased the abundance of their potential prey, there was road 

building associated with thinning, or water level or quality was compromised during the 

embryonic and larval periods (Bull and Wales, 2001). 

• McLeod and Gates (1998) concluded that in general, forest cutting and burning could locally 

decrease herpetofaunal diversity, but diversity on a much larger scale could potentially increase. 

Prescribed burning on experimental forest plots has been shown to increase diversity and 

abundance of amphibians and reptiles over control plots (Mushinsky, 1985). However, there is 

no consistent response of amphibians to prescribed fire; it appears to vary by ecoregion, habitat, 

and species. Some studies suggested burning negatively affects amphibians (for example, Jones 

and others, 2000; Pilliod and others, 2003; Schurbon and Fauth, 2003), whereas others found 

either positive effects, failed to detect significant effects, or found mixed effects whereby some 

species benefit and others decline (for example, Means and Campbell, 1982; Greenberg, 1994; 

Ford and others, 1999; Moseley and others, 2003). In a review of the effects of prescribed fire on 

herpetofauna, Russell and others (1999) found that prescribed fire can be an important tool to 

maintain aquatic habitats of some herpetofauna. In the current literature survey, however, no 

studies were found that specifically addressed the effects of prescribed fire on NLFs, and 

therefore this is definitely a research gap. 

• Several studies have suggested that connectivity among breeding, summer, and overwintering 

habitat for NLFs may be equally as important as the simple abundance of these habitats (Pope 

and others, 2000; Gibbs and others, 2005; Eigenbrod and others, 2008). Mazerolle and 

Desrochers (2005) studied the landscape movements of NLFs and found that when presented 
153

with a choice, 72 percent of the individuals avoided anthropically disturbed surfaces devoid of 

cover, such as mined peatlands and agricultural fields. The authors concluded that these 

disturbed areas disrupted the ability of frogs to reach new habitat patches. 

• One of the most important management activities could be simply initiating surveys for NLFs 

(McAllister and others, 1999). It is important for the conservation of NLFs in the SLV to find 

additional occupied habitat, if it occurs. McAllister and others (1999) suggested that once ponds 

occupied by NLFs are found, they should be mapped, monitored, and evaluated. Annual or 

biennial egg mass counts could also be performed. Transects could be established for the late 

summer or early fall to determine numbers, sex, and ages of leopard frogs. Dieni (2009) also 

strongly recommended that an inventory of amphibians be conducted on the wet meadows of 

Baca NWR, with an emphasis on anurans, such as the NLF. 

• Smith and Keinath (2007) suggested that the following conservation and management factors 

should be considered when trying to conserve NLFs in the Rocky Mountain region: 

• Protection of known and potential breeding sites 

Semlitch and Bodie (1998) and Semlitch (2000b) noted that breeding ponds that 

produced the highest density and biodiversity of amphibians are not protected by current 

Federal law. While streams, springs, and large lakes are protected, NLFs need smaller 

(less than 5 ha) seasonal and semi-permanent ponds (Semlitch and Bodie, 1998; 

Semlitch, 2000b). 

Upland areas surrounding these ponds are used throughout the summer as 

foraging habitat and must also be protected (Semlitsch, 1998; 2000a). Dole (1965a; 

1965b) showed that NLFs typically used a home range of about 68 to 503 m2

, but Dole’s 

work does not provide specific guidance for how much upland core area should be 

protected to conserve an entire population because some of that population will have 

home ranges farther from the breeding pond than other members (Smith and Keinath, 

2007). Adult NLFs typically avoid upland areas that have grass more than one meter tall, 

wooded areas, open areas lacking vegetation, and heavily grazed or mowed areas 

(Merrell, 1977; Mazerolle and Desrochers, 2005). 

The pattern of spacing of suitable breeding sites across the landscape and upland 

movements made by NLFs are probably both important in colonization of new ponds or 

recolonization of ponds in which breeding populations have gone extinct, thus 

maintaining a healthy metapopulation of the species in any area. Frog movements and 

pond spacing are two of the most important factors to consider in management of NLFs 

(Smith, 2003). 

• Control of introduced predaceous fish 

Introduced nonnative predaceous fish have been clearly implicated in the decline 

of some frogs (Bovbjerg, 1965; Hecnar and M’Closkey, 1997b), and ongoing fisheries 

management by various agencies maintain populations of these fish throughout the 

Rocky Mountain region. If fish are found in ponds with NLFs, they might be safely 

removed by electroshocking or rotenone, but only when NLFs are not in the pond (Smith 

and Keinath, 2007). 

• Protection of overwintering sites 

It is important to identify and protect overwintering sites for NLFs since 

overwintering mortality of ranid frogs can be high at times (Bradford, 1983). NLFs use 

lakes, larger ponds, and streams in which to overwinter (Merrell, 1977; Cunjak, 1986), 

but identification of specific sites requires underwater work during the winter, and no 
154

work has been done in the Rocky Mountain region to my knowledge. More research is 

needed on winter ecology for NLFs.

• Control of introduced infectious diseases 

Modes of infectious disease transmission are now well known for NLFs (Smith 

and Keinath, 2007). Carey and others (1999) noted that introduced exotics might spread 

disease, and tiger salamanders are known to harbor iridoviruses (Daszak and others, 

1999). Presumably, diseases can travel on any animal agent, including humans, from 

pond to pond. 

• Water quality

Water quality has major effects on amphibians, as demonstrated previously. 

Ponds, especially those known to have breeding and overwintering populations of frogs, 

should be monitored for the quality of their water, particularly those close to sources of 

pollutants, those where herbicides and pesticides may be used to control exotic plants 

and other noxious organisms, those near construction sites or logging activities, those 

near oil and gas development, and those near roadways and ponds near recreation sites 

(Smith and Keinath, 2007). 

• Protection of migratory or dispersal pathways 

Dole (1965b; 1967; 1971) and Merrell (1977) found that NLFs use many routes 

and habitats for migration and dispersal, including wet meadows, tall grass, and riparian 

corridors. Habitat destruction and road construction interrupt these pathways, and it is 

likely that grazing and logging do so as well (Smith and Keinath, 2007). Smith and 

Keinath (2007) suggested that these movement corridors be afforded protection, but no 

detailed studies were found that examine the management of such areas. It would be 

necessary to understand the landscape (for example, where ponds are placed, where 

riparian corridors are located and how these relate to ponds, and where wet meadows and 

tall grass are located) to manage migratory pathways (Smith and Keinath, 2007). Also 

needed is more research on how frogs move among suitable breeding ponds. 

• Road-related mortality

Limiting and controlling the amount of road-related mortality was suggested as

an important management consideration by Smith and Keinath (2007). As already noted, 

in addition to the fragmentation of the landscape, roads can result in direct mortality to 

NLFs. 

Other factors can have unexpected effects on NLFs, such as immobility response 

due to generalized noise (Nash and others, 1970) and temporary blinding by bright lights 

(Buchanan, 1993). Sometimes controlling for these factors could be impractical, but they 

should be considered in any land management and work planning (Smith and Keinath, 

2007). For instance, the effects of bright lights should be considered where important 

breeding ponds are located near lighted facilities such as offices, work areas, and 

campgrounds. 

Research Priorities

The following research priorities were found in the literature review: 

• Smith and Keinath (2007) describe the following broad categories of information needs 

for NLFs in Region 2 of the USFS (Rocky Mountain Region): (1) Survey and 

monitoring; (2) Mapping of habitat; (3) Characterization of habitat; (4) Characterizing 
155

overwintering habitat using radiotelemetry; (5) Studies of movements both within the 

breeding season and during migration; (6) Disease, pollution, and limb malformations; 

(7) Effects of introduced predaceous fish; (8) Effects of cattle grazing; (9) Effects of 

forest management practices, specifically timber removal; (10) Effects of roadways; (11) 

Genetic studies; and, (12) Effects of prescribed fire and fire suppression. 

• The NLF Recovery Plan for Alberta (Alberta Northern Leopard Frog Recovery Team, 

2010) suggested that research to facilitate the conservation and recovery of the species 

should focus on monitoring of populations and habitat, and evaluating the effectiveness 

of recovery initiatives. Research should also continue on developing increasingly 

effective protocols for reintroduction and establishment of NLF populations. The 

Recovery Team further suggest that “genetic information collected from Alberta 

populations and populations from other jurisdictions should be used to help guide the 

reintroduction program (for example, information on genetic diversity, local 

adaptations).” Research tools such as geographic information systems (GIS) technology 

should be used to identify potential sites for translocation of NLFs and to identify 

previously unknown sites that may support the species. Finally, all recovery activities 

will require the surveillance for, and possible study of, diseases that may affect 

populations and reintroduction efforts (Alberta Northern Leopard Frog Recovery Plan, 

2010). 

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174

X. Tadpole Shrimp 

(Triops longicaudatus) 

Introduction 

The tadpole shrimp (Triops

longicaudatus), also called the 

longtail tadpole shrimp, American 

tadpole shrimp or rice tadpole 

shrimp, is a freshwater crustacean 

of the order Notostraca. It 

resembles a miniature horseshoe 

crab. It is characterized by an elongated, segmented body, flattened shield-like brownish carapace 

covering two thirds of the thorax and two long filaments on the abdomen. They are called tadpole 

shrimp because of their superficial resemblance to frog larvae (Golzari and others, 2009). Triops refers 

to its three eyes and longicaudatus refers to the elongated tail structures. The tadpole shrimp is found in 

freshwater ponds and pools, often in places where few higher forms of life can exist. Like its relative 

Triops cancriformis (horseshoe shrimp, endangered in Europe), the tadpole shrimp is considered a 

living fossil because its basic prehistoric morphology has changed little in the last 70 million years. The

tadpole shrimp is widely distributed throughout western United States and its range is expanding 

eastward; it has recently been found in rice fields in the bootheel of Missouri (Tindall and others, 2009; 

Ridings and others, 2010). 

A simple Google search on the Internet for “Triops longicaudatus” revealed 1,050,000 results 

and a Google Scholar search provided 579 results (accessed 11 July 2011). “Tadpole shrimp” revealed 

107,000 results through Google and 983 results through Google Scholar (accessed 11 July 2011). 

Natural History

The tadpole shrimp is in the crustacean class Branchiopoda, a morphologically diverse group of 

ecologically important, largely freshwater organisms that has a fossil record extending back to the 

Upper Cambrian (Walossek, 1993; Brendonck and others, 2008). Most large branchiopod species rely 

on seasonal wetlands, salt flats, and alkali pans, while other species occur in permanent playas, fishless 

alkali lakes, and salt lakes (Dumont and Megrea, 2002). These habitats are, for the most part, seasonal 

pools that are dry for a significant portion of the year, or several years (Dumont and Megrea, 2002). The 

tadpole shrimp are predominantly benthic and omnivorous, feeding on detritus and living or dead 

organisms (Martin, 1992). They are also known to feed on mosquito larvae (Maffi, 1962), amphibian 

tadpoles (Knoepffler, 1979), mayflies (Callibaetis californicus; Walton and others, 1991), and rice 

seedlings (Crossland, 1964). 

Branchiopod crustaceans rely on banks of resting eggs (or “cysts”) to bridge periods of drought 

or frost (Brendonck, 1996) and buffer against the effects of environmental variability. The eggs lay 

dormant in the substrate until the pool dries and refills during subsequent rains. In the temporally 

fluctuating environment of temporary aquatic habitats, usually only a portion of the dormant eggs, if 

any, hatch during each flooding period (Hildrew, 1985). This process can lead to the generation of an 

egg bank that can remain viable for decades or possibly centuries without rehydration (Belk, 1998), and 

can harbor potentially great genetic and species diversity (Ellner and Hairston, 1994). Many species of 

branchiopoda are distributed throughout their range as eggs, blown about as dust or transported by birds 

or insects (Pennak, 1989). In fact, on the Colorado Plateau, wind is suspected to be the primary dispersal 

Tadpole Shrimp, courtesy of Dominik Tomaszewski. 
175

mechanism for large branchiopod cysts (Graham and Wirth, 2008). Other modes of cyst dispersal 

include transport in or upon larger animals (for example, ducks; Proctor, 1964) or attached to dry pieces 

of wind-blown vegetation. 

The eggs of tadpole shrimps are small, light in weight because of their dry alveolar shells, and 

sticky when laid (Longhurst, 1955). A release from diapause is triggered by changes in temperature, 

daylight length, dissolved oxygen, or salinity (Scott and Grigarick, 1979; Pearse and others, 1987; Su 

and Mulla, 2002a). Previous studies on the reproductive and sex ratios of the tadpole shrimp have 

revealed the occurrence of hermaphrodites along with gonochoric individuals (individual containing 

only male or only female reproductive organs). Grigarick and others (1961) reported that tadpole shrimp 

in the rice fields of California exhibited gonochoric and hermaphroditic reproduction. Sex ratios in 

many populations of tadpole shrimp are skewed towards females (Garcia-Valzco and others, 2009). 

Tadpole shrimp do not feed until they develop into the juvenile stage when they initially feed on detritus

and later become omnivorous. Individuals reach sexual and trophic maturity in as little as six days 

(Harings, 2008). 

Sassaman and others (1997) used electrophoresis to analyze 31 populations of tadpole shrimp at 

sites ranging from California to Kansas and found that the species is actually a mixture of at least two 

reproductively isolated species. In the central United States, including the SLV in Colorado, the 

predominant species is Triops longicaudatus, which is found typically in ephemeral prairie pools. In the 

southwestern United States, the predominant species is Triops newberryi (Packard), which 

characteristically inhabits large playa pools. The two species coexist occasionally and in sympatric 

situations they are reproductively isolated from each other. The two species are genetically distinct at a 

level greater than is typical of conspecific populations. 

It is not uncommon for tadpole shrimp to co-occur in the same ephemeral pond with members of 

several other branchiopod orders such as Anostraca and Conchostraca (Graham, 1995). Branchiopod 

communities in ephemeral pools with at least one species in each of the three orders (Notostraca, 

Anostraca, and Conchostraca) have been reported in a number of studies (for example, Horne, 1971; 

O’Brian and DeNoyelles, 1972; McKay and others, 1990). Tadpole shrimp have also been shown to 

structure macroinvertebrate communities in playa lake microcosms (Yee and others, 2005). Yee and 

others (2005) found that removing tadpole shrimp shortly after hatching reduced the abundances of 

many other taxa, and decreased subsequent taxonomic richness and diversity. Direct effects of tadpole 

shrimp removal include the reduction of prey species abundance, which in turn may alter biotic 

interactions among other taxa. Indirect effects include physical modification of the environment during 

foraging through surface sediments. Walton (2001) also found that the successional pattern and size 

structure of aquatic insect communities differed among ponds in the California desert (Coachella 

Valley) depending on the presence of the tadpole shrimp, Triops newberryi. The densities of aquatic 

insects in Coachella Valley ponds were nearly an order of magnitude lower when Triops was present.

Both developing and adult tadpole shrimp can survive long periods in hypoxic situations that 

alter Hemoglobin (Hb) concentration (Scholnick and Snyder, 1996; Guadagnoli and others, 2005), 

which affords the animal an increase in oxygen-carrying capacity. Tadpole shrimp also have been found 

to reduce oxygen demand to match oxygen supply (Harper and Reiber, 2006). 

A major selective force on the biota of ephemeral environments is the limited time individuals 

have to grow to reproductive age before the environment becomes inhospitable (Davis and Madison, 

2000). The ability of the tadpole shrimp to develop and grow quickly is crucial to its success. Much of 

the research on tadpole shrimp has focused on the desiccation-resistant eggs, whereas almost all 

research on the behavior of the species has focused on the adult form (Davis and Madison, 2000). 

Young tadpole shrimp typically inhabit the surface areas of ephemeral pools, whereas older individuals 
176

are usually benthic. Davis and Madison (2000) postulated several reasons why it is adaptive for young 

tadpole shrimp to be surface water inhabitants. First, the higher oxygen concentration in surface water 

would promote rapid growth and earlier maturation of the young. Scholnick (1995) found that 

individuals raised in hyperoxic conditions grew faster and produced eggs four days earlier than those 

raised in hypoxic conditions. Second, surface water is warmer, and therefore could accelerate 

development of young tadpole shrimp. Oxygen consumption rates are three times more sensitive to 

temperature in immature tadpole shrimp than in older individuals, whose oxygen consumption rates 

remain relatively constant with changes in temperature (Hillyard and Vinegar, 1972). As a result, young 

tadpole shrimp (<5 days) raised at higher temperatures grow faster than those raised at lower 

temperatures, a pattern that is not seen in individuals older than five days. Individuals raised at higher 

temperatures produced eggs a day earlier than those raised at lower temperatures (Scholnick, 1995). 

Additionally, there is some evidence that older individuals grow faster in colder temperatures (Scott and 

Grigarick, 1978). Diet differences between young and adult tadpole shrimp could also lead to vertical 

habitat segregation. Adult tadpole shrimp are primarily benthic feeders (Kaestner, 1970), but young 

tadpole shrimp do not have the appendages needed for efficient benthic feeding. These differences in 

body morphology make it possible for older individuals to spend more time lower in the water column, 

which eases selective pressures on younger individuals. Predation pressure on young and old tadpole 

shrimp may also favor segregation of habitat. Larger and older tadpole shrimp are depredated by birds 

(Takahashi, 1977), which are likely to attack individuals in surface waters. In contrast, young tadpole 

shrimp, are too small for avian predators and can benefit from the higher oxygen and warmer 

temperatures at the surface with less risk from avian predation. Young tadpole shrimp are also 

extremely vulnerable to cannibalism by adults (Scott and Grigarick, 1978; Murataw, 1995) making it 

particularly advantageous for young tadpole shrimp to swim higher in the water column, away from 

benthic adults. 

Factors affecting survival and growth of tadpole shrimp are water temperature (Fry and Mulla, 

1996; Hillyard and Vinegar, 1972) and breeding density (Takahashi and Gohda, 1981; Weeks, 1990; 

Weeks and Sassaman, 1990). Unfavorable temperatures can suppress hatching of eggs (Hall, 1961; 

Moore, 1967; Horne, 1971). For instance, in Wyoming the seasonal appearance of seven different 

branchiopods was thermally dependent, whereas survival of adult shrimp was unaffected by wide 

fluctuations in temperature (Horne, 1967). A similar pattern was also found in Texas ephemeral ponds 

(Horne, 1971). 

Succession of macroinvertebrates in playas of the Southern High Plains of western Texas was 

studied by Moorhead and others (1998). They found that the playas were all dry within 90 days. Species 

richness and diversity increased significantly over time; some taxa increased in abundance (especially 

insects), whereas others decreased (most crustaceans). Trophic structure also changed over time, with a 

significant reduction in detritivores and filter feeders as time progressed coincident with an increase in 

the abundance of predators (Moorhead and others, 1998). These results suggested a rapid development 

of macroinvertebrate assemblage in playas, beginning with an early dominance of crustacean 

detritivores and filter-feeders (for example, tadpole shrimp), followed by later dominance of 

herbivorous and predaceous insects. 

Threats

Worldwide changes in land use, brought about mainly by agriculture and urbanization, have led 

to a global loss and deterioration of temporary pools. Large branchiopods are used in the United States

for assessment of ephemeral wetland habitat functions and values (Rogers, 1998), and five species in the 

United States are currently listed as threatened or endangered in California and Oregon. 
177

The application of insecticides can be a threat to tadpole shrimp, and various synthetic 

pyrethroids and organophosporous pesticides have been used to control tadpole shrimp, especially in 

rice fields of southern California (Walton and others, 1990b). Copper sulfate has also been used to 

control tadpole shrimp populations in rice fields (Grigarick and others, 1985). 

Salinity has a negative effect on the tadpole shrimp (Horne, 1967). Horne (1967) conducted field 

and laboratory experiments on tadpole shrimp by exposing them to different concentrations of salinity. 

Definite decreases occurred in percentage hatch of eggs subjected to osmotic concentrations greater than 

22 mmoles/liter (999 ppm), and eggs failed to hatch during the 2-week incubation period in higher 

salinities (66 mmole/liter). After having been exposed for 2 weeks to the latter solution, tadpole shrimp 

eggs hatched if the environment was subsequently diluted. Tadpole shrimp are also less resistant to 

increasing salinities if the magnesium concentration is high (Horne, 1968). 

The introduction of exotic fishes to naturally fishless waters is often associated with declines of 

species native to those waters. Mosquitofish (Gambusia affinis) have been introduced to waters 

worldwide for mosquito control. Taxa negatively affected by mosquitofish potentially include tadpole 

shrimp. Leyse and others (2003) tested the effects of mosquitofish on a fairy shrimp (Linderiella

occidentalis) in experimental ponds that included a community of native aquatic invertebrates. Survival 

of fairy shrimp was significantly reduced in ponds with mosquitofish, with only 16.9 percent survival in 

ponds with mosquitofish and 64 percent survival in control ponds. Fish presence also reduced the 

abundance and biomass of other invertebrates. They concluded that introduction of mosquitofish into 

naturally fishless wetlands may lead to lower diversity of highly specialized aquatic fauna.

Climate change has the potential to threaten tadpole shrimp populations. The species most 

directly impacted by changes in climate may be the obligate aquatic organisms with life histories that 

are tightly coupled to hydrologic conditions. Vernal pool branchiopods, such as the tadpole shrimp, fit 

this definition (Pyke, 2005). Pyke (2005) assessed the potential impacts of climate change on vernal 

pool ecosystems and endemic branchiopods in the Central Valley of California. He examined the 

potential changes in hydrological regimes for vernal pools from various climate change scenarios. 

Current climate change predictions for terrestrial areas in the Northern Hemisphere indicate higher 

maximum and minimum near-surface air temperatures, decreasing diurnal temperature ranges, more 

intense precipitation events, increased summer continental drying, and an increased risk of drought 

(Houghton and others, 2001; Stott and others, 2001; Johns and others, 2003). Pyke (2005) predicted that 

the warmer, higher-precipitation conditions would result in a net shift toward longer, more frequent 

periods of inundations for vernal pools in the Central Valley. An increased duration of inundation has 

the potential to yield increases in both the diversity and abundance of aquatic predators (Schneider, 

1997). Several authors have noted negative correlations between the abundance of predators and 

branchiopods (Sublette and Sublette, 1967; Dodson, 1987; Loring and others, 1988). Increased 

predation could be mitigated by changes in the relative frequency of reproductively suitable, but more 

ephemeral, vernal pools (Pyke, 2005). 

Management

Literature found related to management and tadpole shrimp mostly focused on its role as a pest 

in ricefields and its capabilities as a mosquito-control agent. No studies that examined the effects of 

mowing, haying, or burning were found. The following is a selection of citations found related to 

management of tadpole shrimp: 

• The tadpole shrimp is a pest in rice fields in at least seven countries on four continents 

(Grigarick and others, 1961; Grigarick, 1984). Alternatively, these shrimp have been used to 

control weeds in rice fields in Japan (Takahashi, 1977; 1994), and also have been proposed as a 
178

biological control agent of mosquitoes (Tietze and Mulla, 1991; Fry and others, 1994). In the 

United States, tadpole shrimp are sporadic pests in rice fields, mainly in California (Grigarick 

and others, 1961; Rosenberg, 1947). Recently, tadpole shrimp were found in ricefields in 

Missouri, representing a range expansion into the northern Mississippi River alluvial plains 

(Tindall and others, 2009). Damage occurs when the shrimp feed upon and uproot young 

seedlings and the characteristic increase in suspended silt as a result of their burrowing action 

decreases the photosynthetic capacity of the rice plant (Rosenberg, 1947). Tadpole shrimp can 

be controlled after hatch by chemical applications (Grigarick and others, 1961). In contrast, 

tadpole shrimp are considered beneficial to rice fields in Japan, where cultural practices differ 

from those in the United States (Fry and Mulla, 1992). In Japan, rice seedlings transported from 

nurseries to fields have reached a size at which shrimp are no long able to dislodge or harm 

them. This rice culture practice promotes the use of tadpole shrimp for weed control because 

they feed upon and uproot young weed plants (Takahashi, 1977). 

• Tadpole shrimp have been noted to devour mosquito larvae (Mail, 1934; Maffi, 1962; Scott and 

Grigarick, 1979; Tietze and Mulla, 1989), which led to the supposition that they could be used in 

mosquito control efforts. In recent studies, Tietze and Mulla (1989; 1990; 1991) have 

demonstrated reduction in mosquito populations by tadpole shrimp through active predation. 

Tietze and Mulla (1991) also reported that the surface swimming behavior of tadpole shrimp 

interferes with the oviposition activity of gravid mosquitoes on the water surface. Furthermore, 

tadpole shrimp were demonstrated to have the potential to enhance the efficacy of microbial 

control agents such as Bacillus thuringiensis spp. israelensis (Bti) for mosquito control as their 

digging activity and vertical foraging process in the water column facilitate the availability of Bti 

toxin particles to mosquito larval feeding (Fry and others, 1994; Fry-O’Brien and Mulla, 1996; 

Kumar and Hwant, 2006). However, when another version of larvicidal oil was used, Bacillus

sphaericus Neide and Golden Bear-1111®, it caused 100 percent mortality of tadpole shrimp 

within 48 hours after treatment in the laboratory even at very low dosages (Su and Mulla, 2005). 

• The effectiveness of tadpole shrimp for controlling mosquitoes depends on the species of 

mosquito and local habitat features. For example, Harings (2008) conducted experiments with 

tadpole shrimp and the mosquito, Culex pipiens, and concluded that tadpole shrimp would 

probably be ineffective control agents of this particular species’ larvae, especially when algae 

are highly abundant. 

• To develop and promote tadpole shrimp as predators for controlling mosquitoes, eggs and 

mature adult shrimp were introduced in 2000 in a date garden in the Coachella Valley, southern 

California that was devoid of preexisting tadpole shrimp (Su and Mulla, 2002b). One year after 

the introduction of the shrimp eggs and adults, it was found that the plots with relatively high 

numbers of tadpole shrimp had reduced numbers of mosquitoes (Psorophora columbiae) by 73 

to 99 percent. 

• Pond or field flooding regimens may directly affect hatch rates of tadpole shrimp eggs by 

providing soil conditions either unsuitable for or amenable to hatching (Fry and Mulla, 1992). 

As in other floodwater organisms, desiccation of eggs is an integral part of tadpole shrimp 

biology in natural situations. Qualitative observations suggest that their eggs do not readily hatch 

on flooding when the soil or substrate containing eggs is moist (Igarishi, 1970). Therefore, in 

intermittent flooding systems with controlled irrigation, soil moisture content may be managed 

by manipulation of flood intervals. Potentially, control of soil moisture content may be used

alternatively either to maximize shrimp numbers or to eliminate their hatch altogether (Fry and 

Mulla, 1992). Fry and Mulla (1992) examined the relationship between drying (moisture 
179

content) of soil after flooding and tadpole shrimp hatches at two locations in southern California. 

They found that soil moisture content is indeed a very important factor influencing the hatching 

rates of tadpole shrimp. Decreasing soil moisture may provide a physiological cue to tadpole 

shrimp eggs that allows hatching to occur at the next flooding, provided embryonic development 

is complete. In the field, no hatch was apparent from soils measuring >24.3 percent moisture. 

They suggested that if large populations of tadpole shrimp are desired as either mosquito or 

weed control agents, flooding of the habitat containing tadpole shrimp eggs can begin as soon as 

the soil surface is dry. The only other apparent requirement is that the water temperature exceed 

a lower thermal limit. For tadpole shrimp collected from the Central Valley of California, this 

lower limit is 14oC (Scott and Grigarick, 1979). 

• Irrigation practices may have a large effect on the population of tadpole shrimp where they exist 

in agricultural systems (Fry and Mulla, 1992). In flood-irrigated fields or other areas where 

water supplies can be manipulated, tadpole shrimp populations can be manipulated as well. Fry 

and Mulla (1992) suggested that this type of water management could ultimately be used to 

increase shrimp populations where they are beneficial organisms playing a role as mosquito or 

weed control agents. Similarly, the control of flooding regiments could also be used to inhibit 

hatching of tadpole shrimp where they are considered pests (for example, rice fields in 

California). Water management would reduce the use of pesticides necessary to control the 

shrimp. 

• Hamasaki (1999) showed that tadpole shrimp did not occur in rice paddy fields employing 

organic farming methods, although they occurred in conventionally farmed fields. Organic fields 

had lower water pH (high acidification) than conventional fields and he suggested that water pH 

could be responsible for this difference (Hamasaki and Ohbayashi, 2000). Hamasaki and 

Ohbayashi (2000) examined survival of tadpole shrimp under different pH conditions and the 

effect of rectifying the acidity of organic farming paddy soil using Ca(OH)2. Treatments with pH 

of 4.4 to 5.3 killed all larvae within 12 hours. At treatments with pH 5.9 to 6.6, the survival rate 

of the larvae until 24 hours was more than 50 percent, and the rate increased with increasing pH. 

Treatment of soil with Ca(OH)2 increased the pH to 6.7 and increased survival to 60 percent. 

They concluded that the boundary of suitable and unsuitable water pH for larval survival was 

between 5.7 and 5.9, and the larvae can grow to ovipositional stage by treating soil acidity 

(Hamasaki and Ohbayashi, 2000). 

• Selected rice fields on the Sacramento National Wildlife Refuge Complex were aerially sprayed 

one time during May or June 1982 with either ethyl (0.11 kg AI/ha) or methyl (0.84 kg AI/ha) 

parathion for control of tadpole shrimp (Custer and others, 2008). No sick or dead vertebrate 

wildlife were found in or adjacent to the treated rice fields after spraying. Specimens of the 

following birds and mammals were assayed for brain cholinesterase (CheE) activity to determine 

exposure to either form of parathion: house mouse (Mus musculus); black-tailed jackrabbit

(Lepus californicus); mallard (Anas platyrhynchos); ring-necked pheasant (Phasianus

colchicus); American coot (Fulica americana); and red-winged blackbird (Agelaius phoeniceus). 

Both mice and pheasants from methyl parathion-treated fields had overall mean ChE activities 

that were significantly (P < .05) inhibited compared with controls, and 7, 40, 54, and 57 percent

of individual blackbirds, pheasants, mice, and coots, respectively, had inhibited brain ChE 

activities (for example, less than -2 SD of control mean)(Custer and others, 1985). 

• Colonization and succession of mosquitoes and macroinvertebrate predators such as tadpole 

shrimp were studied in 30-m2

 ponds (mesocosms) during summer and fall 1987 (Walton and 

others, 1990a). Larval abundance of Cx. tarsalis Coquillett was lower during the hot, summer 
180

months than during the fall. In all studies, larval populations declined markedly 2-3 wk after 

habitat flooding. Although predator abundances differed in these studies, sometimes by an order 

of magnitude, the common predators colonized mesocosms in the following order: tadpole 

shrimp, hydrophilid beetle larvae, dytiscid beetle larvae, mesoveliids, dragonfly and damselfly 

naiads, and notonectids. The similarity of the colonization phenologies probably resulted from 

the vagility of the adult insects and species-specific developmental rates. Stepwise multiple 

regression was used to identify factors potentially affecting larval mosquito populations. For 

most studies, coleopteran larvae were related inversely to per capita change in the entire larval 

population and the third- and fourth-instar subpopulation (for example, large coleopteran larval 

populations were associated with large declines in the Cx. tarsalis larval population). Maximum 

water temperatures and pond age (days after flooding) also were identified as significant factors 

affecting larval abundance and per capita change of mosquitoes. Potentially lethal water 

temperatures (≥35°C) occurred during the summer; however, the declines in larval abundance of 

Cx. tarsalis were not restricted to (or obviously associated with) periods of high water 

temperature. Study results indicated that predation by coleopteran larvae and factor(s) associated 

with pond age, such as mosquito ovipositional preferences, significantly affected Cx. tarsalis

larval populations (Walton and others, 1990a). 

• Methyl farnesoate (MF), shown to inhibit adult metamorphosis in several crustaceans, was 

shown to be a native, juvenilizing factor that delays adult metamorphosis of tadpole shrimp 

(Nelson, 2006). Juvenile hormone III, possessing a structure similar to that of MF, has been 

established as a hormone that regulates metamorphosis in insects. Methyl farnesoate may have 

an analogous role in crustaceans. The effect of MF on oocyte production, performed exclusively 

by adults, and MF metabolism were assayed. Treatment with MF reduced oocyte production in 

individuals treated during the larval/juvenile stages, suggesting retention of the juvenile 

morphotype and inhibition of gonad development. When administered to adults, no reduction in 

oocyte production was observed. MF-synthesis by the enzyme farnesoic acid O-methyl 

transferase was found in the several tissues. Nelson (2000) concluded that MF is synthesized by 

the tadpole shrimp and regulates adult metamorphosis; thus it is a native, juvenilizing agent in 

tadpole shrimp. 

Research Priorities

No suggested list of research priorities or information needs for tadpole shrimp were found 

during this literature review.

Bibliography

Angeler, D.G., Viedma, O. and Sanchez-Carrillo, and Alvarez-Cobelas, M., 2008, Conservation issues 

of temporary wetland Branchiopoda (Anostraca, Notostraca: Crustacea) in a semiarid agricultural 

landscape: what spatial scales are relevant?: Biological Conservation, v. 141, p. 1224-1234. 

Belk, D., 1998, Global status and trends in ephemeral pool invertebrate conservation: implications for 

California fairy shrimp: Pp. 147-150 in Whitham, C.W., Bauder, E.T., Belk, D., Ferren, W.R., Jr., 

Ferren, W.R., Jr., and Ornduff, R., eds., Ecology, Conservation, and Management of Vernal Pool 

Ecosystems – Proceedings from a 1996 Conference, California Native Plant Society, Sacramento, 

California.

Brendonck, L., 1996, Diapause, quiescence, hatching requirements: what we can learn from large 

freshwater branchiopods (Crustacea: Branchiopoda: Anostraca, Notostraca, Conchostraca): 

Hydrobiologia, v. 320, p. 85-97. 
181

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Summary 

Table 1. A summary of general and specific (when reported in the literature) habitat requirements for the 10 focal species in the Alamosa / Monte 

Vista / Baca National Wildlife Refuge Complex, Colorado. This table includes the species common name, the size of wetland or area, general 

habitats, nesting/breeding habitats, and foraging habitats. Text in this table can be found in the individual species accounts with appropriate citations.

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

American 

Avocet

• 6-19.3 ha;

• Daub (1993) 

sampled 

wetlands ranging 

from <1 to 19.3 

ha, and 

concluded that 

the avocet is an 

area-dependent 

species. 

• >8 ha (Prescott 

and others, 

1995). 

• Seasonal wetlands and 

habitats dominated by Baltic 

rush (Juncus balticus), 

sedges (Carex spp.), and 

grasses <40 cm tall over 

semipermanent wetlands, 

habitats dominated by cattail 

(Typha spp.) and softstem 

bulrush (Scirpus validus) 

>40 cm tall, saltgrass 

habitats, or upland shrub 

habitats;

• Use exposed, sparsely 

vegetated salt flats, sandbars, 

peninsulas, mudflats, or 

islands adjacent to shallow 

(<1 m deep) water, 

conditions that occur in 

wetlands, lakes, fallow and 

flooded fields, or 

impoundments.

• Found in tilled, alkali, ephemeral, 

temporary, seasonal, 

semipermanent, and permanent 

wetlands, and in impoundments, 

sewage lagoons, and evaporation 

ponds;

• Unvegetated ground or in areas 

with short, sparse vegetation that 

provides unobstructed views from 

nests;

• Distance of avocet nests to water 

range from 0 to 300 m, although 

most nests are located within 60 m 

of water;

• Prefer islands for breeding and this 

may be due to greater protection 

from nest predators.

• Main foods taken by avocets are 

aquatic invertebrates of the water 

column and in sediment; also eat 

terrestrial invertebrates, small fish, 

and seeds;

• Ceratopogonidae (biting midges) 

and Chironomidae (nonbiting 

midges) were the dominant 

families of invertebrates eaten by 

avocets in migratory stopover sites 

in the Southern Great Plains;

• Avocets forage at various water 

depths, depending on age and bill 

length: Young (0-3 week-old) 

chicks forage at water depths of 0-

90 mm, but mainly at about 8 mm. 

Older (3-6 week-old) chicks 

forage at depths from 0 to 100 

mm, but mostly at about 53 mm. 

Adult females forage at >80 mm, 

and adult males forage at about 

100 mm.

 
188

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

Wilson’s 

Phalarope

• There is some 

evidence for area 

sensitivity;

• the species was rare 

in patches of CRP 

grassland that were 

<100 ha.

• Seasonal wetlands and 

habitats dominated by baltic 

rush (Juncus balticus), 

sedges (Carex spp.), and 

grasses <40 cm tall over 

semipermanent wetlands, 

habitats dominated by cattail 

and softstem bulrush 

(Schoenoplectus 

tabernaemontani) >40 cm 

tall, saltgrass habitats, or 

upland shrub habitats. 

• Three characteristics: open water, 

emergent vegetation, and open 

shoreline. Prefer to nest in moist 

sedge and rush meadows 

characterized by low plant height;

• Adjacent open water is also a 

habitat requirement;

• Nest site selection varies seasonally 

for this species. They nest in 

upland vegetation early in the 

breeding season and wet-meadow 

vegetation later in the season; 

• They usually nest <100 m from 

shoreline; 

• In Nebraska, habitat measurements 

around nests were 26-32 cm 

average height, 46-55 cm 

maximum vegetation height, 18-23 

percent bare area, 99.9 percent 

grass cover, 3876-4385 stems/m2

stem density, 0.74 cm litter depth, 

4.2-4.6 m from water, 1.5 mm stem 

diameter, 829 g/m2 above-ground 

biomass, 1.7 cm interstem distance, 

and 32-44 ha lake surface area.

• Main foods eaten by Wilson’s 

phalaropes are small aquatic 

invertebrates (dipterans, 

crustaceans, particularly brine 

flies and brine shrimp) in 

freshwater or hypersaline 

environments;

• They also feed on some terrestrial 

invertebrates and occasionally on 

seeds of aquatic plants;

• Forage in open-water, flooded 

meadows, and less frequently in 

upland habitats and long beaches;

• Forages principally while 

swimming. 

Sora

• Area-independent in 

their selection of 

wetlands.

• Across their breeding range, 

soras primarily use wetlands 

with shallow and 

intermediate water depths, 

dominated by robust or fineleaved emergent vegetation, 

especially cattails (Typha

spp.), sedges (Carex spp., 

Cyperus spp.), burreeds 

(Sparganium spp.), and 

bulrushes (Scirpus spp.);

• In Iowa, mean water depth in 71 

breeding territories was 38.4 cm +

16.1 standard deviation (SD)(range 

0-92) and was not significantly 

different from the Virginia rail 

(Rallus limicola); 

• Mean vegetation height was 128.0 

cm + 42.5 SD, and mean number of 

stems/m2 in territories was 121.9 

cm + 80.9 SD; 

• Soras did not use areas where 

heavy snow, ice, and high water 

• The main foods eaten by soras are 

the seeds of wetland plants and 

aquatic invertebrates; 

• Proportion of invertebrates in their 

diet increases in the spring; 

• Common plant foods include 

seeds of wild or cultivated rice, 

smartweeds, sedges, bulrushes, 

and grasses; 

• Principal animal foods include 

adults, larvae, and pupae of a 

variety of aquatic insects (Orders 
189

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

• The Breeding Bird Atlas 

Colorado (1998) reported 

soras breeding in a variety of 

wetlands in Colorado, 

including cattail marshes, 

grass or sedge marshes, wet 

meadows, and irrigated 

hayfields especially in 

mountain parks;

• Soras prefer cattails with 

shallow water for breeding 

and escape, and 80 percent 

of Atlas reports came from 

marshes, with cattails 

usually the dominant 

vegetation.

had flattened emergent vegetation 

until early May, new sedge growth 

was 20-30 cm in height and 80-100 

stems/m2 provided some cover. 

Coleoptera, Diptera, Hemiptera, 

Odonata) and snails (Gastropoda);

• Soras generally feed from a 

standing position. In autumn, 

soras strip seed heads of maturing 

smartweeds and annual grasses; 

• They occasionally feed by 

swimming in open water like a 

coot; 

• Stands of robust emergent 

vegetation interspersed with 

shorter, seed-producing

vegetation, and vegetative debris 

that provides good substrate for 

invertebrates near the water 

surface;

• May be attracted to shallower 

portions of wetlands during broodrearing and premigration periods 

where seed-producing plants such 

as sedges, bulrushes, smartweeds, 

beggarticks (Bidens spp.), and 

grasses provide food; 

• In the late summer, soras may 

leave wetlands for short periods to 

feed in upland fields and row 

crops. 

White-faced 

Ibis

• Not specified • Prefer, almost exclusively, 

areas with emergent 

vegetation of some type;

• Usually nest in emergent vegetation 

or low trees and shrubs over 

shallow water, but can also nest on 

the ground on small islands; 

• Main foods taken are aquatic and 

moist-soil insects, crustaceans, 

and earthworms;

• Forages in flooded hay meadows, 

agricultural fields, and wetlands 

(for example, pond and reservoir 

margins, mudflats, and marshes) 

with short, emergent vegetation;

Sora

(continued)
190

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

• They inhabit primarily 

freshwater wetlands and 

marshes, especially those 

containing cattail (Typha

spp.) and bulrush (Scirpus

spp.);

• The landscape mosaic must 

consist of wetlands that have 

new vegetation growth and 

that contain suitable water 

levels (approximately 1 m in 

the SLV, Colorado) from 

late April through the end of 

July to sustain ibises through 

the breeding season.

• Ibises need some degree of 

isolation by marshes or small 

islands for both nesting colonies 

and social night roosts;

• At two locations in Utah all nests 

were in hardstem bulrush or alkali 

bulrush; 

• A nest’s height, above the water or 

ground, may vary depending upon 

substrate type or other conditions; 

• In one Utah colony, nests ranged 

from 99.0-202.2 cm above water 

that was 61 cm deep; 

• The availability of feeding areas 

within 15-25 km of breeding 

colonies is a very important 

landscape feature for the whitefaced ibis; 

• The importance of foraging areas 

was demonstrated the summer of 

2003 at the Monte Vista NWR in 

Colorado where nesting colonies 

were deserted in the middle of the 

breeding season. It is believed that 

the primary reason for 

abandonment and low reproduction 

on the Refuge was due to the 

minimal or lack of shallow-water 

wet meadows where ibises forage. 

Most ibises nesting in Colorado 

favor tall emergent vegetation such 

as bulrushes and cattails growing as 

“islands” surrounded by water 

more than 45 cm deep.

• Dominant plant species in 

foraging areas generally are 

sedges (Carex spp.), spikerushes 

(Eleocharis spp.), salt-tolerant 

grassworts (Salicornia spp.), 

saltgrass (Distichlus stricta), and 

greasewood (Sarcobatus

vermiculatus);

• In Nevada, Colorado, Utah, Idaho, 

California, and Oregon irrigated 

crops of alfalfa, barley, and hay 

are important feeding sites for 

ibises;

• In Nevada, white-faced ibises feed 

in recently flooded agricultural 

fields where vegetation is <5 to 90 

cm high, and they show a strong 

preference for alfalfa fields.

• Strong preferences were also 

found for large (>30 ha), level (<5 

percent slope) fields with clay or 

clay-loam soils and pools of 

standing water;

• Laubhan and Gammonley (2000) 

speculated that differences in food 

abundance, vegetation structure 

and composition, and behavioral 

strategies during the breeding 

season, all influenced the ibis’ 

selection of foraging habitats in 

the SLV of Colorado.

White-faced

Ibis (continued)
191

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

Black Tern

• Preferred marshes 

>20 ha and only 

rarely used smaller 

marshes (5 ha 

minimum) unless part 

of larger wetland 

complexes

• In eastern South 

Dakota, found that 

terns nested in larger 

basins and wetland 

complexes;

• The minimum area 

requirement for the 

black tern as a 

semipermanent 

wetland basin of 12.4 

ha.

• Freshwater lakes, rivers, 

other interior wetlands, 

plowed fields, and coastal 

wetlands;

• More likely to occur in 

wetlands with surrounding 

grasslands that were <50 

percent tilled for agriculture.

• Use a wide variety of vegetation 

types in nesting areas; 

• Nests are usually built on floating 

substrates of matted or decaying 

marsh vegetation and often mixed 

with mud;

• Floating substrates can also be 

detached root masses, logs and 

boards, muskrat (Ondatra

zibethicus) feeding platforms or 

clippings, algae or peat mats, lily 

pads, dried cowpies, and old nests 

of grebes, American Coots (Fulica

americana), and Forster’s terns 

(Sterna forsteri). These substrates 

are usually anchored to or lodged in 

emergent vegetation or dense beds 

of submerged rooted aquatics;

• Nonfloating substrates include 

muskrat lodges, raised mud 

patches, marshy hummocks, rooted 

flattened vegetation, and upturned 

tree roots with attached vegetation; 

• Nests can be particularly prone to 

flooding;

• In Colorado, black terns breed in 

large cattail marshes next to open 

water; 

• Rockwell (1911) described a nest at 

Barr Lake as built on a dense carpet 

of cattails blown over by the wind 

about 3 m from shore in a sparsely 

covered spot amid dense cattail 

growth and over about 15 cm of 

water, another on a large round top 

of a duck blind floating in almost 

waist deep water at the edge of 

cattails in a small “rush-bound” 

pond, and a third on a mass of dead 

• Breeding black terns are mainly 

insectivorous, but fish make up a 

large part of the diet in some 

habitats and regions;

• Forage over freshwater lakes, 

rivers, and other interior wetlands;

• They also forage over plowed 

fields and coastal wetlands. 
192

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

floating cattails;

• Nest semi-colonially, placing their 

nests in clusters in favorable areas 

of marshes; clusters are typically 

about 11-50 nests but can range 

from two to hundreds; 

• Nests are usually 5 to 20 m apart, 

but sometimes as close as 1 m; 

• Up to 25 percent to 30 percent of 

birds nest “solitarily,” or from 20 to 

30 m up to 600 m from other nests; 

• Generally select nest sites with an 

approximate 50:50 vegetation 

cover to open water ratio that is 

well interspersed with water. 

Short-eared 

Owl

• No specific area 

requirements found 

although they 

“typically occur in 

large expanses of 

prairie or coastal 

grasslands, 

heathlands, shrubsteppe, and tundra.” 

• Require large 

(minimum 100 ha) 

tracts of land for 

successful nesting.

• Most often in grasslands, 

prairies, and wetlands (68

percent of habitat 

observations), less 

commonly in shrublands, 

and rarely in scrub oak or 

pinyon-juniper woodlands.

• On the Great Plains, short-eared 

owls prefer to nest within large 

patches of relatively tall (30 to 60 

cm), dense, ungrazed grassland;

• Herkert and others (1999) in 

Illinois found that they required 

relatively short grass (<50 cm tall) 

and most grasslands in the region 

needed to be mowed or burned to 

maintain this condition.

• Small mammals, particularly 

Microtus, are the main foods taken 

by short-eared owls;

• Less frequently, this species will 

hunt birds;

• Short-eared owls hunt primarily 

on the wing, approximately 0.3 to 

3.0 m above ground;

• They also hover 2-30 m above 

ground or, less frequently, perched 

on poles or hills.

Brewer’s 

Sparrow

• Density of this 

species was 

significantly 

correlated with total 

shrub cover, and 

capture rates were 

significantly 

• Sagebrush-steppe obligate 

species;

• Throughout most of its range 

in the United States, it is 

most closely associated with 

landscapes dominated by big 

• Breeds in shrublands and the 

average canopy height is usually 

<1.5 m;

• Compared with surrounding 

habitat, nests are located in 

significantly taller, denser shrubs 

• The main foods taken by Brewer’s 

sparrows are small insects, mainly 

gleaned from foliage and bark of 

shrubs and dwarf trees;

• Brewer’s sparrows forage mostly 

in shrubs (>75 percent of over 600 

Black Tern

(continued)
193

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

correlated with shrub 

cover and dead wood 

cover; 

• They were positively 

influenced by factors 

indicating more 

structurally complex 

habitat;

• Appear to be areasensitive, and isolated 

stands of sagebrush 

smaller than 2 ha are 

not likely to be used 

as nesting habitat.

sagebrush (Artemisia

tridentata), where it is 

usually the most abundant 

bird species;

• Also prefer small openings 

(for example., <1 ha) of 

short vegetation surrounded 

by sagebrush.

(primarily big sagebrush) with 

reduced bare ground and 

herbaceous cover;

• Available cover may be the 

primary factor in nest-site 

selection;

• Nests are located toward the edge 

of the densest portion of sagebrush 

relative to nests of the sage sparrow 

and the sage thrasher;

• Brewer’s sparrows prefer nest 

shrubs to be entirely alive or mostly 

alive

• They have also been found nesting 

in willow in Colorado.

observation periods) and less on

open ground between shrubs or at 

the base of bunchgrasses;

• Shrubs selected by foraging 

sparrows differ significantly from 

those randomly available; they are 

larger, more vigorous, and more 

likely to consist of sagebrush 

rather than green (Chrysothamnus

viscidiflora) or gray (C. 

nauseosus) rabbitbrush.

Savannah 

Sparrow

• Area-incidence 

functions indicated 

that savannah 

sparrows require open 

areas of 10-15 ha in 

grasslands and 

blueberry barrens; 

• In South Dakota, 

savannah sparrows 

were found to be 

area-sensitive and 

used only large 

grassland patches.

• In Colorado, croplands, 

montane grassland, and 

emergent marshes accounted 

for 58 percent of habitats 

reported. Salt meadows, 

mountain sagebrush, 

montane carr, and willow 

carr made up another 22 

percent of sightings;

• Irrigated hay meadows are 

an important component of 

savannah sparrow habitats in 

mountain parks and western 

valleys in Colorado.

• Breeding range habitat is open 

country including grassy meadows, 

cultivated fields (especially alfalfa), 

lightly grazed pastures, roadsides, 

coastal grasslands, sedge bogs, 

edge of salt marshes, and tundra;

• They avoid areas with extensive 

tree cover, generally occurring in 

fields with some herbaceous plants 

or weeds;

• In arid parts of their range, they are 

generally restricted to irrigated 

areas or to the grassy margins of 

ponds;

• Savannah sparrows prefer dense 

ground vegetation, especially 

grasses, and moist microhabitats 

are favored;

• In Colorado, moist, grassy, 

mountain meadows are favored 

breeding habitats, but irrigated hay 

meadows and alfalfa are also 

important throughout Colorado.

• During the breeding season, 

primarily forage on adult and 

larval insects, spiders, seeds and 

fruits;

• Occasionally, they will eat insect 

eggs, millipedes, isopods, 

amphipods, decapods, mites, and 

small mollusks;

• Mainly forages on the ground in 

short vegetation in pastures, sedge 

bogs, tundra, beach wrack, mud 

flats (near cover), and dunegrass;

• In winter, they forage primarily on 

small seeds of grasses and other 

herbs, small-seeded fruits, insects, 

and spiders.

Brewer’s

Sparrow 

(continued)
194

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

Northern 

Leopard Frog

• No specific area 

requirements found, 

but because NLFs 

have a complicated 

life history, they 

require a broad range 

of habitats in close 

proximity; 

• Three categories of 

major habitat type 

were described by 

Merrell and Rodell 

(1968): winter habitat 

(overwintering in 

lakes, streams, and 

ponds); summer 

habitat (feeding by 

adults in upland 

areas); and, tadpole 

habitat (up to three 

months spent as 

tadpoles in shallow 

breeding ponds);

• Their complicated 

movement patterns 

during the year 

should also be

considered;

• Prefer smaller (less 

than 5 ha) seasonal 

and semi-permanent 

ponds.

• Typical habitats described 

in Colorado include wet 

meadows and the banks and 

shallows of marshes, ponds, 

glacial kettle ponds, beaver 

ponds, lakes, reservoirs, 

streams, and irrigation 

ditches;

• Adult NLFs are usually 

found at the water’s edge, 

but they may roam far from 

permanent water during 

mild wet weather;

• Winter months are spent at 

the bottom of water bodies 

such as ponds and streams;

• Adult NLFs typically avoid 

upland areas that have grass 

cover one meter tall, 

wooded areas, open areas 

lacking vegetation, and in 

heavily grazed or mowed 

areas.

• NLFs breed in shallow, quiet areas 

of permanent bodies of water, in 

beaver ponds, and in seasonally 

flooded areas adjacent to or 

contiguous with permanent pools or 

streams; 

• Breeding areas contain vegetation, 

mats of algae, and clear water; 

• They use small (usually <5 ha) 

ponds in which to breed;

• Egg masses are attached to 

vegetation just below the surface of 

warm shallows about 7 to 25 cm 

deep; 

• At elevations of 1,570 to 2,520 m 

in Colorado and southern 

Wyoming, egg masses average 

about 3,000 eggs (range 645-

6,272). 

• There is little information on the 

food habits of the NLF in 

Colorado; however, invertebrates 

most likely dominate the diet of 

adult frogs;

• Gehlback (1965) reported that 

NLFs ate tenebrionid beetles in 

northwestern New Mexico;

• Drake (1914) studied the food 

habits of NLFs in Ohio and found 

that the diet consisted of molluscs, 

crustaceans, spiders, and insects;

• They use moist upland habitats 

surrounding their breeding ponds 

for summertime foraging
195

Species 

(common name) Size of wetland or area General habitats Nesting/Breeding habitats Foraging habitats

Tadpole 

Shrimp

• No area requirements 

found in the 

literature.

• Rely on seasonal wetlands, 

salt flats, and alkali pans, 

permanent playas, fishless 

alkali lakes, and salt lakes;

• These habitats are, for the 

most part, seasonal pools 

that are dry for a significant 

portion of the year, or 

several years;

• Young tadpole shrimp 

typically inhabit the surface 

areas of ephemeral pools, 

whereas older individuals 

are usually benthic.

• Rely on banks of resting eggs 

(or “cysts”) to bridge periods 

of drought or frost and buffer 

against the effects of 

environmental variability;

• The eggs lay dormant in the 

substrate until the pool dries 

and refills during subsequent 

rains;

• In the temporally fluctuating 

environment of temporary 

aquatic habitats, usually only a 

portion of the dormant eggs, if 

any, hatch during each 

flooding period;

• This process can lead to the 

generation of an egg bank that 

can remain viable for decades 

or possibly centuries without 

rehydration.

• Predominantly benthic and 

omnivorous, feeding on detritus 

and living or dead organisms; 

• They are also known to feed on 

mosquito larvae, amphibian 

tadpoles, mayflies (Callibaetis

californicus), and rice seedlings. 

Publishing support provided by: 

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For more information concerning this publication, contact:

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2150 Centre Ave., Bldg. C

Fort Collins, CO 80526-8118

(970) 226-9398

Or visit the Fort Collins Science Center Web site at:

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