United States 

Department of Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical Report 

RMRS-GTR-135-vol. 2.

September 2005

Assessment of Grassland 

Ecosystem Conditions in the 

Southwestern United States: 

Wildlife and Fish

Volume 2

Editor

Deborah M. Finch
Abstract__________________________________________________

Finch, Deborah M., Editor. 2005. Assessment of grassland ecosystem conditions in the Southwestern United States: 

wildlife and fish—volume 2. Gen. Tech. Rep. RMRS-GTR-135-vol. 2. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 168 p.

This report is volume 2 of a two-volume ecological assessment of grassland ecosystems in the Southwestern United 

States. Broad-scale assessments are syntheses of current scientific knowledge, including a description of uncertainties 

and assumptions, to provide a characterization and comprehensive description of ecological, social, and economic 

components within an assessment area. Volume 1 of this assessment focused on the ecology, types, conditions, and management practices of Southwestern grasslands. Volume 2 (this volume) describes wildlife and fish species, their habitat 

requirements, and species-specific management concerns, in Southwestern grasslands. This assessment is regional in 

scale and pertains primarily to lands administered by the Southwestern Region of the USDA Forest Service (Arizona, New 

Mexico, western Texas, and western Oklahoma). A primary purpose of volume 1 is to provide information to employees 

of the National Forest System for managing grassland ecosystems and landscapes, both at the Forest Plan level for Plan 

amendments and revisions, and at the project level to place site-specific activities within the larger framework. This 

volume should also be useful to State, municipal, and other Federal agencies, and to private landowners that manage 

grasslands in the Southwestern United States.

Key words: grasslands, ecological assessment, Southwestern United States, ecosystem conditions, wildlife, fish, birds, 

mammals, Arizona, New Mexico

Acknowledgments_________________________________________

This project is a collaborative effort between the Rocky Mountain Research Station and the Southwestern Region of 

the U.S. Department of Agriculture, Forest Service. In early 2000, a Southwestern Grassland Ecosystem Sustainability 

Team was formed to determine the assessment approach and general content. By 2001, the team had evolved into a 

core group that finalized the assessment topics and authored the report. This current work is volume 2 of a two-volume 

report. Volume 1 was published in September 2004.

The core team members are:

Robert Calamusso, Southwestern Region, Tonto National Forest

Cathy Dahms, Southwestern Region, Ecosystem Analysis and Planning (retired)

Deborah Finch, Rocky Mountain Research Station

Reggie Fletcher, Southwestern Region, Ecosystem Analysis and Planning (retired)

Paulette Ford, Rocky Mountain Research Station

Gerald Gottfried, Rocky Mountain Research Station

Burt Pendleton, Rocky Mountain Research Station

Rosemary Pendleton, Rocky Mountain Research Station

Deborah Potter, Southwestern Region, Ecosystem Analysis and Planning

Carol Raish, Rocky Mountain Research Station

Bryce Rickel, Southwestern Region, Wildlife, Fish and Rare Plants

Wayne Robbie, Southwestern Region, Ecosystem Analysis and Planning

Michele Merola-Zwartjes, Rocky Mountain Research Station

We acknowledge Art Briggs (Southwestern Region), Jim Saveland (Rocky Mountain Research Station), Dave Stewart 

(Southwestern Region), and George Martinez (Southwestern Region) for their early role in the team’s development, and 

Alison Hill (Rocky Mountain Research Station), Bob Davis (Southwestern Region), Amy Unthank (Southwestern Region), 

and Don DeLorenzo (Southwestern Region) for assisting in seeing both volumes through.

We thank Louise Kingsbury and her staff for assistance in editorial work and layout of chapters in preparation for 

publication. We thank Nora Altamirano for help in formatting the document. The team also recognizes the valuable input 

provided by other team members and by numerous manuscript reviewers. We thank Carl and Jane Bock, Rob Beltars, 

John Rinne, Janet Ruth, Fred Samson, John Sidle, and Amy Unthank for reviewing manuscripts for volume 2. We also 

acknowledge helpful input from Fritz Knopf, Christopher Rustay, Troy Corman, Bill Howe, and Sartor O. Williams on 

one or more chapters.

Support for this project was provided by USDA Forest Service’s Southwestern Region, and by the Forest Service’s 

Rocky Mountain Research Station (Ecology, Recovery, and Sustainability of Grassland and Riparian Ecosystems in the 

Southwest, Cultural Heritage Research, and Ecosystem Management Southwestern Borderlands).

Editor’s Note______________________________________________

Although this report may be used as input in processes initiated under the National Environmental Policy Act (NEPA), 

National Forest Management Act (NFMA), and other applicable laws, it is not a decision document, does not allocate 

resources on public lands, and does not make recommendations to that effect. The information in this report is general 

in nature rather than site-specific. The opinions expressed by the authors do not necessarily represent the policy or position of the U.S. Department of Agriculture and the Forest Service.
Assessment of Grassland 

Ecosystem Conditions in the 

Southwestern United States: 

Wildlife and Fish

Volume 2

Editor

Deborah M. Finch

USDA Forest Service

Rocky Mountain Research Station

333 Broadway SE, Suite 115

Albuquerque, New Mexico 87102

Cover photos clockwise from upper left: Black-tailed prairie dog (photo by Jeff Vanuga, USDA Natural Resources Conservation Service). Pronghorn (photo by Gary Kramer, USDA Natural Resources Conservation Service). Spikedace (photo 

by James E. Johnson, U.S. Fish and Wildlife Service). Lesser prairie chicken (photo by Gary Kramer).
The Authors_______________________________________________

Bob Calamusso has a B.S. degree in fish and wildlife management from Montana State University, Bozeman, 1983 and 

an M.S. degree in fisheries science from New Mexico State University, Las Cruces, 1996. Bob has worked in the fisheries management field with the Montana Department of Fish, Wildlife, and Parks throughout Montana. He has worked 

in the fisheries management area in Alaska, New Mexico, and Arizona with the Forest Service and has also conducted 

research in affiliation with the Rocky Mountain Research Station in New Mexico and Arizona. The majority of Bob’s 

work in Region 3 has been focused on the conservation of native fishes and their ecosystems on National Forest properties and adjoining lands.

Deborah M. Finch is a Research Wildlife Biologist and Project Leader with the USDA Forest Service’s Rocky Mountain 

Research Station in Albuquerque, NM. She is also Research Associate Professor in the Department of Biology at University 

of New Mexico. She received her B.S. degree in wildlife management from Humboldt State University, California, her 

M.S. degree in zoology from Arizona State University, and her Ph.D. degree in zoology from the University of Wyoming. 

Deborah’s research interests include neotropical migratory bird ecology and reproduction, bird migration, endangered 

species, riparian and grassland ecosystem ecology and conservation, and relationships between commodity use and 

natural resources. Deborah has published more than 125 articles and journal papers and is an active member of several 

professional societies. Since 1998, she has served as Leader of the Technical Subgroup of the Recovery Team for the 

Southwestern willow flycatcher.

Michele Merola-Zwartjes is a Fish and Wildlife Biologist and is working for the U.S. Fish and Wildlife Service, Division 

of Endangered Species Recovery, Region 1, Portland, OR. She was a postdoctoral Research Wildlife Biologist with the 

Rocky Mountain Research Station in Albuquerque, NM. She holds a B.A. degree in communications from the University 

of California, San Diego, and a Ph.D. degree in biology from the University of New Mexico. Amongst her varied work 

experiences, Michele has served as a conservation planner for The Nature Conservancy and taught in the Wildlife Department of Humboldt State University. Michele’s interests include ornithology, conservation of threatened, endangered, 

and sensitive vertebrate species, animal behavior, wildlife management, and natural resource management issues.

Bryce W. Rickel received a B.S. degree in general biology and an M.S. degree in fishery biology from the University of 

Arizona. He has worked for the USDA Forest Service for 23 years as a Fishery and Wildlife Biologist in Arizona, Alaska, 

Utah, and New Mexico. Currently, he is the Regional Wildlife and Fish Habitat Coordinator for the Southwestern Region. 

His interests are in developing GIS applications and in using artificial intelligence for species habitat quality models and 

for habitat relationships programs.

ii
Preface___________________________________________________

This report is volume 2 of a two-volume ecological assessment of grassland ecosystems in the Southwestern United 

States, and it is part of a series of planned publications addressing major ecosystems of the Southwest. Volume 1, An 

Assessment of Grassland Ecosystem Conditions in the Southwest (Finch, editor, 2004), focused on the ecology, types, 

conditions, and management practices of Southwestern grasslands. The second volume (herein) describes wildlife and 

fish species, their habitat requirements, and species-specific management concerns, in Southwestern grasslands.

The first Southwestern ecological assessment, General Technical Report RM-GTR-295, emphasized forested ecosystems and was titled, An Assessment of Forest Ecosystem Health in the Southwest (by Dahms and Geils, editors, 1997). 

Given the complexities of grassland ecology and the increasing number of challenges facing grassland managers, the 

USDA Forest Service Southwestern Region, in partnership with the agency’s Rocky Mountain Research Station, focused 

on grasslands in its second assessment. The assessment is regional in scale and pertains primarily to lands administered 

by the Southwestern Region (Arizona, New Mexico, Texas, and Oklahoma) of the U.S. Department of Agriculture, Forest Service.

Broad-scale assessments are syntheses of current scientific knowledge, including a description of uncertainties and assumptions, to provide a characterization and comprehensive description of ecological, social, and economic components 

within an assessment area (USDA Forest Service 1999b). A primary purpose of volume 2 of the grassland assessment is 

to provide information to employees of the National Forest System for managing habitats and lands for wildlife and fish 

populations, both at the Forest Plan level for Plan amendments and revisions, and at the project level to place site-specific 

activities within the larger framework. This volume should also be useful to State, municipal, other Federal agencies, 

and to private landowners that manage or regulate wildlife and fish populations and their habitats in the Southwestern 

United States. The assessment is not a decision document because it identifies issues and risks to grassland ecosystems 

that provide the foundation for future changes to Forest Plans or project activities, but it does not make any site-specific 

decisions or recommendations.

To conduct the entire assessment, we assembled a team of authors from the Southwestern Region and the Rocky 

Mountain Research Station whose expertise focused on or included grassland ecosystems. An outline of chapter titles 

and chapter contents was prepared using a group consensus process. Authors volunteered to write specific chapters that 

were then reviewed by the team. Following team review, each individual chapter was sent to a minimum of two peer 

reviewers for critique. This volume then went through an editorial process by myself, and by the Rocky Mountain Research Station’s Publishing Services Office.

I thank the authors of volume 2 for writing and rewriting their chapters, and I repeat my thanks to authors of volume 

1 for their contributions. Authors and I thank reviewers of volume 2 chapters for their helpful and constructive comments 

and advice. These reviewers were John Sidle and Mark Rumble (USDA Forest Service, Rocky Mountain Research Station) 

and Fred Samson (USDA Forest Service, Northern Region) for chapters 1, 2, and 3; Carl and Jane Bock (University of Colorado) and Janet Ruth (U.S. Geological Survey) for chapter 4; and Rob Bettaso (Arizona Game and Fish Department), John 

Rinne (USDA Forest Service, Rocky Mountain Research Station, Flagstaff, AZ), and Amy Unthank (USDA Forest Service, 

Southwest Region) for chapter 5. We also thank all members of the grassland assessment team for their contributions to 

the team planning effort. On behalf of the team, I express thanks to Don DeLorenzo, Art Briggs, Cathy Dahms, and Bob 

Davis of the Southwest Regional Office, and to Alison Hill of the Forest Service’s Rocky Mountain Research Station for 

supporting this project. In addition, I am pleased to acknowledge Louise Kingsbury and her staff for publication editing 

and layout, and to Nora Altamirano for assistance in assembling and formatting the report.

Deborah Finch, Editor

USDA Forest Service

Rocky Mountain Research Station

Albuquerque, NM

iii
Contents_________________________________________________

Chapter 1: Wildlife .............................................................................................................. 1

Bryce Rickel

Chapter 2: Large Native Ungulates .................................................................................... 13

Bryce Rickel

Chapter 3: Small Mammals, Reptiles, and Amphibians ...................................................... 35

Bryce Rickel

Chapter 4: Birds of Southwestern Grasslands: Status, 

 Conservation, and Management ....................................................................................... 71

Michele Merola-Zwartjes

Chapter 5: Fishes of Southwestern Grasslands: Ecology, 

 Conservation, and Management ..................................................................................... 141

Bob Calamusso

iv
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 1

Bryce Rickel

Chapter 1:

Wildlife

Introduction_______________________

This volume addresses the wildlife and fish of the 

grasslands in the Southwestern Region of the USDA 

Forest Service. Our intent is to provide information that 

will help resource specialists and decisionmakers manage wildlife populations within grassland ecosystems 

in the Southwestern United States. The information 

and analysis presented is at a Regional scale.

Many of the references and examples in this volume 

are from grassland types and/or States outside the 

Region but are applicable to our discussions. Some species are addressed at length, either because information 

is available, the species are threatened, endangered, 

being considered for listing, are sensitive, have special 

management requirements, or play or have played a 

special ecological role.

An underlying theme in managing wildlife residing in grassland ecosystems is that wildlife and their 

environments have evolved together over millennia. 

Each wildlife species has an important role and function to fill in its grassland environment. The challenge 

of managing grassland-adapted wildlife species is to 

sustain or restore, where possible, the processes, attributes, and habitat structures that have codeveloped 

with populations over time and that may be required 

by individual wildlife species.

Many human activities have significant effects on 

wildlife of grasslands, but two in particular are notable 

and frequently identified in the literature: habitat loss 

in relation to private farming, and habitat alteration 

from grazing by domestic livestock. Much of the literature addressing grassland wildlife can be classified 

under range management topics, especially as it relates 

to grazing effects and management of livestock.

Historic Conditions________________

As pioneers moved west in the United States 

century and a half ago, they found large expanses of 

prairie comprising grasses suitable for livestock forage, and level ground with soils amenable to growing 

domestic crops. The development of ranches, farms, 

towns, railroads, and roads began to fragment and 

change native grasslands, and this trend continues 

today. This advancement of civilization has placed an 

increasing demand on natural resources as fundamental as space, water, air, minerals, water, plants, and 

animals. Humans cleared native grasses and trees 

and planted new species. They combated and then 

controlled a natural process that was considered to 

be destructive—wildfire. Fire suppression changed 

various system processes that were important for the 

maintenance of grasslands. Because early pioneers 

found water to be a limiting factor in the Southwest, 

they drilled wells, built dams, and redirected water 

to meet their needs. As this development of the West 

progressed, the face of the landscape changed. The 

combination of fire suppression, water control, and 
2 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

habitat conversion and fragmentation significantly 

altered native grasslands and respective ecosystem 

processes that wildlife depended on.

The topic of grassland-associated wildlife in the 

Southwestern Region of the Forest Service is complex. 

The grasslands in the Region are subdivided into three 

distinct areas: the Great Plains, the Great Basin, and 

the desert grassland. While there are some similarities 

among wildlife species across the grassland types, each 

grassland has unique species with unique requirements 

and roles in its respective ecosystems.

The Great Plains grasslands include the shortgrass 

steppes of southeastern Colorado and northeastern 

New Mexico, northwestern Texas, western Oklahoma, 

and western Kansas. The desert grasslands are in 

southeastern Arizona, southwestern New Mexico, and 

Mexico. The Great Basin, a cold desert, is bounded by 

the Sierra Nevada on the west and Colorado Rockies 

on the east, and by the Idaho batholith of central Idaho 

on the north and a vegetation line defined by creosote 

bush (Larrea divaricata)—a warm desert species—and 

Great Basin sagebrush (Artemisia tridentata) in the 

south. The Great Basin, which is similar to other desert 

grasslands in the Southwest, was once dominated by 

perennial grasses but is currently dominated by desert 

shrubs (Hastings and Turner 1965, Hanley and Page 

1982), including creosote (Larrea tridentata), tarbush 

(Flourensia cernua), mesquite (Prosopsis juliflora), 

Mormon tea (Ephedra trifurca), and acacia (Acacia

spp.).

The Great Plains, Great Basin, and Desert Grassland 

remarkably differ in climate. The desert grasslands 

and the Great Basins are semiarid, whereas the Great 

Plains is wetter and colder. The grassland ecosystems 

of Western North America are climatically diverse with 

differences in productivity, seasonality, and disturbance 

regimes. These environmental dissimilarities have 

resulted in different fauna. If our goal is to sustain 

grasslands, we must understand the climatic and 

evolutionary constraints operating in each ecosystem 

and the resulting differences in resilience of these 

ecosystems and the wildlife species therein.

The grassland ecosystems discussed here are usually described as homogenous types of grass, tallgrass, 

shortgrass, mixed grass, interspersed with many 

smaller ecosystems, including wetlands, playas, riparian areas, shrublands, rocky outcrops, and in some 

locations, scattered trees and shrubs. Each system has 

a unique fauna, management problems, and needs. 

Consequently, only the major groups of wildlife and a 

few of the more important species can be discussed in 

any detail. By addressing species groups and some key 

species most of the management needs of other species 

will, we hope, be met. Most of the species groups and 

single species will be addressed in relation to functions 

within their respective grassland ecosystems and with 

respect to various anthropogenic activities that occur 

on the grasslands.

Many issues need to be considered when managing 

wildlife on grasslands, such as:

• Habitat fragmentation caused by urban development, agriculture (farming), changes in plant 

structure and composition, fencing, and roads.

• The use of chemicals such as pesticides and urban 

contaminates that have affected wildlife, directly 

or indirectly.

• Changes in the abundance and types of water 

sources. For example, playas have been drained 

to create farmlands, streams have been dammed, 

diverted, impacted by grazing or have dried up 

due to lowering water tables, and destruction 

and changes in riparian zones.

• Population declines of several wildlife species 

that have played important ecological functions 

(keystone species) such as bison (Bison bison), 

and prairie dogs (Cynomys spp.). Some grassland 

wildlife species are to the point of being listed or 

being considered for listing.

• Competition between domestic and wild species. Competition will be of sufficient interest 

throughout the following discussions that more 

explanation is merited. Some general definitions of 

competition as they apply in this report follow.

Factors Influencing Wildlife 

Populations_______________________

Competition

Competition can occur in one of two ways. The first 

way involves direct use of a common limited resource 

by the same or different species (Miller 1967, Nelson 

1982). On the grasslands, the common resource is 

vegetation. Many factors influence competition for 

edible plants between wildlife and domestic herbivores 

(Vavra and others 1989). Nelson (1982) lists eight ways 

species could compete for vegetation:

• consumption equivalence

• dietary overlap

• forage quantity and quality

• forage use

• timing of use

• height of foraging reach

• density or stocking rate of animals

• spatial and temporal distribution of animals

These factors are not independent; they can interact 

in multiple ways to heighten or minimize the actual 

degree of competition.

The second way competition can occur involves active defense of a territory or other spatial resources by 

one animal against another Nelson (1982). This can 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 3

also take a more passive form known as disturbance 

competition, when the mere presence of an animal 

intimidates or annoys another animal into leaving 

the area (Nelson 1982).

Figure 1-1 illustrates the competitive interactions 

among such native ungulates as elk, deer, and such 

exotic ungulates as cattle and sheep. Although, figure 

1-1 depicts general competitive interactions between 

ungulates, both native and exotic, it is also a good 

representation of competition between many wildlife 

and domestic species.

Predators

In all of the grassland ecosystems, predators play 

an important function of controlling prey populations 

that may include other predators. Many of the large 

top carnivorous predators—bears, mountain lions, 

wolves—have been extirpated from the grasslands.

Important predators in contemporary desert grasslands include coyotes (Canis latrans), hawks, owls, bats, 

the insectivorous desert shrew (Notiosorex crawfordi), 

the grasshopper mouse, snakes, lizards, and toads. 

Local distributions and numbers of predators can often 

be linked to prey abundance. For example, studies of 

coyote and jackrabbit populations in the Great Basin 

desert showed that jackrabbit abundance rises and falls 

on a decadal cycle and that coyote populations track the 

jackrabbit cycle with a 1 to 2 year lag time (Stoddart 

1987a,b). Predators are also known to regulate prey 

populations. Range caterpillar populations have been 

reduced by as much as 50 percent by rodent predators, 

as have been various darkling (Tenebrionidae) and 

ground beetles (Carabidae) (Parmenter and MacMahon 

1988a,b). Insectivorous birds can have a similar impact 

on grasshoppers as rodent predators (Joern 1986).

Habitat Modification and Fragmentation

Grasslands have been subjected to various kinds 

of modifications, fire events, climate changes, wildlife uses, and other influences, for millions of years 

(Colbert and Morales 1992). In the grasslands of North 

America, native wildlife habitats have been modified, 

removed completely, or replaced with other kinds of 

vegetation by humans. As a result, native habitats 

in some areas have become more disjointed and the 

vegetation composition and structure have changed. 

Volume 1 of this assessment describes the causes and 

consequences of grassland alterations and landscape 

fragmentation in relation to plant communities, and 

the following chapters in this volume 2 document many 

of the effects of habitat modification and fragmentation on wildlife.

Studies of native mammals in disturbed grassland 

situations reveal that some animals have adapted 

remarkably well to these changes; for example, coyotes 

(Canis latrans) now use urban areas. However, many 

larger herbivores, such as bison, elk, antelope, and 

their respective predators, have not been able to adapt 

as well to grassland alterations as smaller herbivores. 

An explanation may be that large herbivores cannot 

readily adapt to habitat fragmentation because they 

depend on large geographic ranges and expanses of 

continuous open grasslands, whereas smaller animals 

may be able to adjust to fragmented habitats because 

they have smaller ranges and need smaller amounts 

of continuous habitat.

Figure 1-1. Potential interactions among variables that affect ungulate competition for food. Interactions 

can heighten or reduce the potential for competition in a myriad of ways that are not predicted (Wisdom 

and Thomas 1996).

Spatial and Temporal 

Distribution of

Animals

Density of Animals

(Stocking Rate)

Height of

Foraging Reach

Timing of

Utilization

Dietary

Overlap

Forage 

Utilization

Consumption

Equivalence

(Animal Equivalence)

Forage Quantity

and Quality

Interactions
4 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Influence of Large Mammals

Effects of large herbivores (ungulates) on plant 

community composition and structure have been 

studied in many grasslands worldwide (Frank and 

McNaughton 1992). Ungulate activities affect many 

aspects of grassland structure and function, including 

the physical structure of the environment and the rates 

of certain ecosystem-level processes, for example, fire 

frequency and erosion probability (Frank and Evans 

1997). The grazing habits of large herds of American 

bison, for example, resulted in a mosaic pattern of 

grazed and ungrazed areas. This pattern in turn 

influenced the occupancy patterns of other mammals 

in the areas.

Role of Riparian Areas in Grassland 

Environments

Riparian vegetation, trees, grasses, sedges, and 

rushes provide shade, cover, and food for many 

wildlife species. Streamside vegetation filters and 

traps sediments, which improves water quality and 

helps to maintain water temperatures and oxygen 

levels. Leaves, bark, and rotting wood provide homes 

and food for many insects and many vertebrate 

species. Insects are prey for birds, small mammals, 

reptiles, amphibians, and fish. Riparian habitats are 

important migratory corridors for birds and bats and 

travel corridors for large mammals such as elk and 

deer. Fragmented riparian habitats can also lead to 

isolated populations of animal species preventing 

both population expansion and gene flow (Ohmart 

and Zisner 1993).

The management of riparian ecosystems within the 

Southwestern Region is a special concern because of 

their importance to so many wildlife species (table 

1-1). About three-fourths of the vertebrates in the 

Southwest, for example, depend on riparian areas for 

some portion of their life history (Johnson 1989). In 

the Southwestern Region, 57 percent of the terrestrial 

vertebrate species on National Forest lands occur 

in riparian systems (estimated from the Southwest 

Wildlife Information System 2001). The significance of 

this habitat type to wildlife becomes even more important when considering that riparian habitats comprise 

just less than 2 percent of the National Forests in the 

Southwestern Region (estimated from the Southwest 

Wildlife Information System 2001).

Because livestock spend much of their time in riparian zones, where ecological stakes are high, many of 

the adverse impacts of grazing are magnified in these 

habitats (Fleischner 1994). Destruction of vegetation 

adjacent to riparian streams has resulted in siltation 

of streambed gravels, warming of the water, and elimination of aboveground flows during summer and fall. 

The United States General Accounting Office (1988) 

reported that degradation of riparian and aquatic 

areas was the most serious threat to biodiversity in 

the West.

Role and Effects of Fire

One of the most important ecological processes of 

the grasslands has been fire. Fire has occurred either 

by lighting naturally, or by humans. Anthropologic 

fires were set by Native Americans and then by 

pioneers for various reasons; for example, to clear a 

piece of land for farming, to renew land for farming, 

to herd large mammals for hunting, and for warfare. 

Once the grasslands were settled, fires were viewed 

as more of a setback than a tool, and fire suppression 

commenced. Fire suppression altered the frequency of 

wildfire events, causing many grassland ecosystems 

to shift in plant species composition from areas composed primarily of grasses to areas having increasing 

numbers of shrubs and trees.

Birds—Some birds react to fire directly; raptors, 

for example, are attracted to fire and smoke. This 

response appears to be related to vulnerability and 

ease of capturing prey that are forced to escape a 

fire. Raptors are also attracted to burned areas where 

insects and small mammals are often plentiful and 

are easy prey. Other bird species are attracted to 

recently burned grasslands (Clark 1935, Handley 

1969, Komarek 1969, Kramps and others 1983, Lyon 

and Marzluff 1984, Tombak 1986). The attraction is 

the increased forb composition. Seed production after 

fire is beneficial to avian herbivores and granivores 

(Bock and Bock 1990, Bock and others 1976, Lawrence 

1966, Wirtz 1977).

Other effects of fire on birds include increased 

habitat heterogeneity. In shrub-grass complexes, bird 

diversity and abundance are enhanced if shrub cover 

and nesting sites are interspersed with open grassy 

areas maintained by fire (Baldwin 1968, Kramp and 

others 1983). Fire can have a cleansing effect on bird 

populations by lowering the numbers of parasites that 

affect health and vigor of individuals (Kramp and 

others 1983, Lyon and Marzluff 1984).

Table 1-1. The total number of birds, mammals, amphibians, 

and reptiles and their respective number in riparian habitats, 

and the percent of the totals on the 11 National Forests in the 

Southwestern Region. (Data compiled from the Southwest 

Wildlife Information System, Rickel 2002)

Species group Total Riparian Percent

Birds 458 252 55

Mammals 203 153 75

Amphibians 34 15 44

Reptiles 114 41 36

Totals 809 461 57
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 5

Species that use shrubby habitat for nesting and 

perching are likely to be negatively impacted by fire if 

shrubs are destroyed. The shrub-associated common 

yellowthroat (Geothlypis trichas), for example, was 

found to reach highest abundance on unburned areas 

in mixed-grass prairie (Madden 1995).

Some bird species exhibit both negative and positive 

response to fires. This dual response is a consequence 

of requiring different habitats for different purposes. 

They may be attracted to recently burned grasslands 

for feeding if productivity of grasses and forbs has been 

improved. However, they may require shrub-dominated 

habitats for cover and nesting (Brown 1978, Komarek 

1969, Kramp and others 1983, Renwald and others 

1978, Wolfe 1973).

Fire or the lack of fire may affect the distribution of bird species. Populations of Baird’s 

sparrow (Ammodramus bairdii), grasshopper sparrow (Ammodramus bairdii), LeConte’s sparrow 

(Ammospiza leconteii), and Sprague’s pipit (Anthus 

spragueii) were common to abundant in mixed-grass 

prairie, where fire has been used as a habitat management tool since the 1970s, but are virtually absent from 

unburned prairie (Madden 1995). Fire suppression in 

grasslands can lead to habitat type replacement, with 

plant community complexes comprising grass species 

replaced by shrub communities over time. This habitat 

conversion is likely to have beneficial effects on some 

bird (and other wildlife) species but detrimental effects 

on other species. Table 1-2 lists a few grassland bird 

species and their responses to fire.

Mammals—How mammals respond in the short 

term to fire is a function of mammal size and mobility. 

Large mobile mammals such as deer and elk can easily avoid injury during fire (Boeker and others 1972, 

McCulloch 1969, Dills 1970, Hallisey and Wood 1976). 

Young ungulates that are less mobile are frequently 

killed by large fires (Daubenmire 1968, Kramp and 

others 1983). Most small mammals escape fires by 

hiding in burrows or rock crevices (Howards and others 1959, Heinselman 1973). Bendell (1974) indicates 

that soil provides insulation from fire for burrowing 

animals (Kramp and others 1983).

The most common fire-related deaths for small 

mammals are from a combination of heat effects and 

asphyxiation, physiological stress caused by overexertion during escape, trampling by stampeding large 

mammals, and predation during and after escape from 

fire (Kaufman and others 1990).

Fires that temporarily remove food and cover may 

be detrimental to small rodents immediately after 

fire (Daubenmire 1968, Kaufman and others 1990). 

However, repopulation of burned areas by small rodents 

is reported to be nearly complete within 6 months (Cook 

1959). Small mammal populations often increase after 

fire. This response is due to increased availability of 

forb seeds and insects (Lyon and others 1978). Burned 

areas often support more diverse animal populations 

than comparable unburned sites due to increased 

habitat diversity (Beck and Vogl 1972, Wirtz 1977). 

Omnivores and carnivores are attracted to burns owing to increased plant diversity and associated small 

prey populations (Gruell 1980).

Kaufman and others (1990) suggests that most of 

the effects of fire on small mammals in grasslands 

are not neutral but instead are either positive or 

negative. Fire-negative mammals (table 1-3) include 

species that forage on invertebrates in the litter layer, 

species that hide in dense vegetation and eat foliage, 

and species that use, at least partially, aboveground 

nests of plant debris. Fire-positive mammals (table 

1-3) include species that feed on seeds and/or insects 

(Kaufman and others 1990). They exhibit an increase 

in populations and habitat use after fire because of an 

increased availability of forb seeds, insects, new green 

vegetation, the creation of open areas in otherwise 

dense habitat, and eventually an increase in forb cover. 

Increases can be immediate or can occur gradually as 

the areas begin to revegetate and habitat diversity 

increases. Carnivores of the southern Great Plains 

increase in population and habitat use in response to 

fire-enhanced rodent populations (prey) (Gruell 1980, 

Kramp and others 1983, Wirtz 1977). Most native ungulates, bison, white-tailed deer, elk, and pronghorn, 

increase in populations and habitat use after fire. 

These population responses are due to an increase in 

forage quality and quantity in newly burned areas.

Ecological Perspective______________

In this section, the ecosystem management issues of 

biodiversity, scale, fragmentation, population viability, 

and keystone species are discussed.

Resource managers must remember that biodiversity, scale, fragmentation, viability, and keystone 

species are not exclusively separate but are all closely 

interrelated and interconnected facets of ecosystems, 

and consequently it is difficult to understand and 

manage for one issue without understanding how they 

interact with others.

Biodiversity

Probably one of the most widely cited definitions 

of biodiversity comes from Technologies to Maintain 

Biological Diversity (Office of Technology Assessment 

1987: 9): “Biological diversity refers to the variety and 

variability among living organisms and the ecological 

complexity in which they occur.” Biodiversity has 

become one of the focal points of much discussion 

about ecosystem management. One of today’s leading 

scientists, E.O. Wilson, predicted that, worse than 
6 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

destruction caused by war, is the loss of genetic and 

species diversity in destruction of natural habitats 

because such loss will take millions of years to repair 

(Wilson 1984). An eloquent quote from Aldo Leopold 

(1949: 190) captures the same thought:

 The last word in ignorance is the man who says 

of an animal or plant: “What good is it?” If the land 

mechanism as a whole is good, then every part is 

good, whether we understand it or not. If the biota, 

in the course of eons, has built something we like but 

do not understand, then who but a fool would discard 

seemingly useless parts? To keep every cog and wheel 

is the first precaution of intelligent tinkering.

Biological diversity is of particular interest since our 

activities affect both vegetative and wildlife diversity. 

People alter the biodiversity of plant and animal communities both directly and indirectly.

Table 1-2. Some grassland birds their response to fire and description of their habitat (Kaufman and others 1990).

Species Response to fire Habitat description

Common yellowthroat Negative Shrubs are required for nesting and perching, but nests are 

Geothlypis trichas destroyed by fire. Yellowthroats were found to reach highest 

abundance on unburned areas in mixed-grass prairie, where

prescribed fire has been used as a management.

Northern bobwhite Both negative and positive. Is attracted to recently-burned grasslands and is most 

Colinus virginianus productive in grass-forb habitat. However, it requires scattered

woody plants for cover, and populations decrease in shrubdominated habitats.

Eastern meadowlark Both negative and positive. May be attracted to recent burns, but fires that destroy all shrub 

Sturnella magna cover may be detrimental.

White-crowned sparrow Both negative and positive. Depends on shrub cover, and may decrease habitat use on some 

Zonotrichia leucophrys burns. But it also aggregates in large groups to feed in open 

burns.

Lark sparrow Both negative and positive. Reportedly benefits from litter removal in grasslands and reducChondestes grammacus tion but not complete removal of shrubs.

Baird’s sparrow Positive These birds were the most common, and abundant birds, overall,

Ammodramus bairdii in mixed-grass prairie, where fire has been used as a habitat 

LeConte’s sparrow management tool since the 1970s, but were all completely 

Ammospiza leconteii absent from unburned prairie. Baird’s sparrow was found to reach 

Sprague’s pipit high densities in areas that had been frequently burned. The 

Anthus spragueii areas were characterized by low litter and high cover variability of 

Western meadowlark forbs and bunchgrass.

Sturnella neglecta 

Montezuma quail Positive The decline of Montezuma quail has been linked with widespread 

Cyrtonyx montezumae replacement of grassland with shrubland in the last 150 years. It 

may benefit from fires that decrease shrub cover.

Burrowing owl Positive Populations of the burrowing owls have reportedly declined on 

Anthene cumicularia grasslands with increases in litter cover. This suggests that the 

use of fire to reduce litter cover may be beneficial to this species

Lesser prairie-chicken Positive The lesser prairie-chicken is also a grassland species reported to 

Tympanuchus pallidicinctus be declining in the Southwest because of decreased grassland 

habitat due to suppression of grassland.

Sandhill crane Positive Regrowth of grasses, reduced litter, and decreased shrub cover in

Grus Canadensis grasslands following fire is beneficial for the crane.

Managing Wildlife Diversity—To maintain suitable grassland conditions for the variety of wildlife 

species adapted to southwestern grasslands, managers 

would benefit by developing a comprehensive approach 

for conserving and enhancing biological diversity, taking into account both species and ecosystem functions 

and processes (Probst and Crow 1991). In developing 

such an approach, three interrelated elements of ecosystem diversity must be analyzed and incorporated 

into management plans and desired future conditions 

(Crow and others 1993):

• Compositional diversity—The diversity of various 

components of an ecosystem, namely species, communities, and ecosystems within larger systems. 

The effective protection of these components usually depends on the next two components.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 7

• Structural diversity—The diversity of habitats 

and how they are arranged relative to each other 

spatially and temporally. Structural diversity includes the size, shape, and distribution of species 

and habitat types, communities across landscapes, 

and patterns of successional change.

• Functional diversity—This refers to the diversity 

of ecological processes that maintain and are 

dependent on the other components of diversity. 

Functional diversity includes, for example, competition, predation, parasitism, and other biological 

interactions, as well as other processes, such as, 

nutrient retention and cycling. The ecological 

processes represented by functional diversity 

provide the “ecological services” necessary to 

support all organisms, including humans.

As resource managers develop approaches for applying the concepts of biodiversity and communities 

to designing desired future conditions, they should 

avoid oversimplifying the complexity of a landscape 

and wildlife habitat relationships. Communities with 

similar resources may still vary in the number of 

species because of different degrees of niche overlap. 

Niche overlap is the degree two species exploit the same 

resource; for example, cattle and deer. Communities 

differ in species requirements for a variety of reasons: 

differences in climatic stability and predictability, 

completion, predation, and disturbances (Solbrig 1991). 

Consequently, as management plans are developed, 

good field inventories will be needed.

Table 1-3. Examples of mammals that have positive or negative response to fire 

(Kaufman and others 1990).

Hispid cotton rat (Sigmodon hispidus berlandieri) Negative

Bailey’s pocket mouse (Chaetodipus baileyi baileyi) Negative

Pinyon mouse, (Peromyscus truei truei) Negative

White-tailed antelope squirrel (Ammospermophilus leucurus) Negative

Southern red-backed vole (Clethrionomys gapperi) Negative

Western harvest mouse (Reithrodontomys megalotis) Negative

Meadow vole (Microtus pennsylvanicus modestus) Negative

Deer mouse (Peromyscus maniculatus) Positive

White-footed mouse (Peromyscus lleucopus) Positive

Merriam’s kangaroo rat (Dipodomys merriami) Positive

Southern grasshopper mouse (Onychomys torridus) Positive

Nuttall’s cottontail (Sylvilagus nuttallii) Positive

Thirteen-lined ground squirrel (Spermophilus tridecemliniatus) Positive

Hispid pocket mouse (Chaetodipus hispidus) Positive

Bison (Bison bison) Positive

Elk (Cervus elaphus) Positive

White-tailed deer (Odocoileus virginianus) Positive

Pronghorn (Antilocapra americana) Positive

Scale

One of the most significant changes and challenges 

in moving from traditional management to ecosystem 

management is to analyze and manage systems at 

multiple spatial and temporal scales, simultaneously 

(Fischer and others 2004). In order to maintain and/or 

restore grassland ecosystem conditions for wildlife, 

such management approaches need to be developed 

and implemented with scale in mind (Crist and others 

2003, Su and others 2004).

A hierarchical (multiple-scaled) perspective emphasizes three strategic concerns in an analysis of 

landscape patterns:

• To detect patterns and define their spatial and 

temporal scales; in other words, define functional 

patches at a specific scale for a species (Crist and 

others 2003, Fischer and others 2004, Kotliar 

and Wiens 1990).

• To infer which factors generate these patterns.

• To relate these patterns to adjacent levels” (Urban 

and others 1987, Fischer and others 2004).

Spatial Scale—The scale at which nature is viewed 

determines the patterns and processes detected. 

Fischer and others (2004) explain that organisms are 

affected by ecological processes operating at multiple 

scales, different habitat variables vary over different 

spatial scales, and some species reflect habitat at 

one, but not all, spatial scales. For any management 

needs or problem, spatial and temporal scales must 
8 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

both be carefully defined. There is no “best” scale at 

which to manage for. The appropriate scale depends on 

management needs (Noss 1992) and the requirements 

of the species. When developing plans for managing 

landscapes, it is essential for managers to address 

scaling in relation to specific questions and needs.

Temporal Scale—Problems in time correspond to 

our limited temporal scale of concern in conservation 

strategies. In many cases these problems involve disturbance. In general, most human impacts on biodiversity 

represent a change in the environment and/or an increased rate of change. Changes in disturbance regimes 

threaten biodiversity when they introduce stresses or 

events either qualitatively or quantitatively different 

from the disturbance or stresses to which organisms 

have adapted over evolutionary time (Noss 1992).

Over long time spans, centuries or millennia, grasslands of any size never remain the same. Changes in 

species distributions most typically result from variation in disturbance regimes and changes in climate. 

Because species migrate at different rates, grassland 

communities are ephemeral and unlikely to attain 

equilibrium (Davis 1981).

Managing Wildlife at Different Scales—Only by 

examining wildlife populations along habitat gradients 

encompassing environmental variations within species’ 

ranges can wildlife be effectively managed. Species 

exhibiting ecological differences in habitat use among 

areas will likely require different types of management, 

depending on location. Species that exhibit specific 

habitat requirements having little variation across 

their range may require only one set of management 

strategies (Block and Morrison 1991).

Biodiversity should be considered at a regional scale 

where attention is paid to species composition as well 

as diversity. Even if management concerns are strictly 

focused on a local area, regional ecological processes 

as well as local processes must be taken into account. 

Failure to consider regional processes that control local biodiversity may result in the disruption of these 

processes, as when habitat fragmentation eliminates 

opportunities for species to migrate in response to 

changing climate or human activities. Conservation 

strategies, even for single species, will be most effective 

when they address ecological phenomena at multiple 

spatial scales and levels of organization (Noss 1992).

Management practices can often be effective when 

they mimic natural disturbances, and these practices 

can account for frequency, intensity, and seasonality of 

disturbances over multiple spatial scales. Managers 

are more likely to maintain native biodiversity by 

using practices that create a variety of habitats required by a range of wildlife species than when they 

create conditions unlike those occurring in “natural” 

landscapes (Noss 1992). For example, using prescribed 

fire to simulate naturally occurring wildfires may help 

to stimulate growth of suppressed plants, restore lost 

species, reduce numbers of invasive shrubs, and reduce 

amounts of senescent vegetation.

Fragmentation

The problem of fragmentation cannot be understated 

or underestimated. Rosenberg and Rapheal (1986) 

state that perhaps the most critical problem facing 

wildlife, worldwide, is the systematic shrinking and 

fragmentation of their habitat. Fragmentation is 

the process where patches of habitats are reduced 

in size and/or isolated from one another by natural 

disturbances such as fire, earthquakes, or flooding, 

or by human development (McLellan and others 

1986). Although most of the research and literature 

has focused on forested lands, the concerns and many 

of the fundamental principles and theories apply to 

grasslands. Often fragmentation of grasslands and 

its effects on wildlife and ecosystem processes and 

functions are subtle. The size and shape of a habitat 

patch, and species composition and structure of the 

patch vegetation, determine the wildlife species that 

occupy any given site. The number and kinds of edge 

and interior species are typically correlated with patch 

size and shape.

Population Viability

A fundamental problem for wildlife management is 

assuring long-term viability of populations of different 

species. Population viability refers to the probability of 

continued existence of geographically well-distributed 

population over a specified period (Marcot and Murphy 

1996). Managing viable populations is more complex 

than simply maintaining minimum population sizes 

needed for populations to survive over time with 

random fluctuations. To assess population dynamics 

accurately, resource managers must consider demographics, ability to disperse, and habitat quantity and 

quality (Probst and Crow 1991).

Central to all planning efforts that involve population viability analysis is provision of well-distributed 

and interconnected habitats (Noss and Harris 1986). 

Although some species’ distributions are naturally 

patchy, in general, the maintenance of distributions 

and connectivity is the key to sustaining genetically 

diverse and demographically healthy populations. A 

population or habitat is well-distributed if it is maintained over the long term across at least its existing 

range of geographic, environmental, and ecological 

conditions (Marcot and Murphy 1996).

Keystone Species

An ecological concept that needs explanation before 

moving on to the rest of this volume is that of keystone 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 9

species and their role within ecosystems. This understanding of keystone species will be important as various 

wildlife species and their management are discussed.

Where did the term “keystone species” come from? 

While Robert Paine, an ecologist, was studying the 

pattern of life among the intertidal rock species 

along the coast of Washington State in the 1960s, he 

found that one species of starfish preyed so skillfully 

on mussels that it effectively kept the mussels from 

monopolizing space on the rock. When he removed the 

starfish from sections of the shoreline, the mussels 

began to multiply, crowding out limpets, barnacles, 

and other marine organisms from the rock surfaces. In 

the absence of the starfish, the total number of species 

living on the rock (other than the mussels) decreased 

by half (Baskin 1997). Paine (1969) coined the term 

“keystone species” to describe species such as the 

starfish. He presented the characteristics required for 

a species to be given keystone status: (1) It provides 

top-down effects (for example, predation) on lower 

tropic levels, and thus (2) it prevents the monopolization of a critical resource in lower tropic levels. And 

(3) the synergy of this dualistic top-down (for example, 

predation) and bottom-up (for example, competition) 

interaction must stabilize community diversity. This 

narrowly defined keystone process is relatively rare 

and functionally nonredundant in ecosystems, making 

keystone species of great ecological significance (Davic 

2002). A keystone species is one whose impacts on its 

community or ecosystem are large and greater than 

would be expected from its relative abundance or total 

biomass (Keystone Species Hypothesis 1969).

Research has shown that not all species are ecologically equal in their importance to a system. Species 

that are abundant, that dominate space and resources 

or contribute to controlling processes are not necessarily the most influential in a system. Sometimes, 

keystone species are less conspicuous and can even be 

rare (Baskin 1997). Their contribution to the system 

is disproportionate to other species in the community, 

and their removal creates ripple effects that might 

both change the community and also alter ecological 

processes (Baskin 1997). These community changes 

can have drastic consequences. If a keystone species 

becomes extinct, the other species that are dependent on 

the keystone species may also become extinct. Keystone 

species help to support the ecosystem of which they 

are a part (Jain and Krishna 2002, Keystone Species 

Hypothesis 1969).

Keystone species may be top carnivores that keep 

prey in check, large herbivores that shape the habitat in 

which other species live, important plants that support 

particular insect species that are prey for birds, bats 

that disperse the seeds of plants, and many other types 

of organisms (Keystone species 2004). Keystone species 

may occur at any level of the ecosystem, from plants 

and herbivores, to carnivores and detritivores.

In many systems, “keystone groups” of species rather 

than individual species assume the role of keystone 

species (De Leo and Levin1997).

Functions of Keystone Species—Keystone 

species can occur in four ways. The first way is when 

organisms control potential dominants. In terrestrial 

systems, large and small herbivores play major roles 

in maintaining vegetation structure and species 

composition. The effect of herbivores is often to suppress potentially dominant plant species, reducing 

competitive exclusion of less competitive species, 

and promoting greater vegetation diversity (Ernest 

and Brown 2001, Payton and others 2002). A second 

way a species can be a keystone is to provide a vital 

resource or resources to a range of organisms (Payton 

and others 2002). A third way species may be keystone 

is when mutualism exists. Mutualism is where two 

species are jointly dependent, and the elimination of 

one will result in the demise of the other. In this situation they act as keystones for each other. What might 

be more in keeping with the multispecies concept of 

keystone species is “group mutualisms.” True group 

keystones arise where there is dependency of several 

species on a single mutualist, for example, pollinators 

and groups of plants (Payton and others 2002). The 

fourth way an organism can be a keystone species is 

by being an ecosystem engineer. These species modify 

the physical characteristic of the environment, for 

example, burrowing, excavating nests, and so forth, 

providing resources for other species. These groups are 

not mutually exclusive, and individual species may 

exhibit characteristics of more than one type (Payton 

and others 2002).

 Management Significations of Keystone 

Species—Natural resource managers are faced with 

complex issues and problems and an array of species, communities, and ecosystems. Faced with this 

complexity and ever-increasing demands for limited 

resources, the question is: How do I as a land manager 

adequately maintain these resources? (Payton and 

others 2002)

Protecting keystone species has become a priority for conservationists. Unfortunately, the keystone 

functions of a species may not be known until it is too 

late, that is, when it is listed or is extirpated. Where a 

keystone species has been identified, efforts to protect 

it also will help protect the other species in delicately 

balanced ecosystems (Keystone species 2004).

To determine the ecological importance of a species, 

a manager’s attention must be directed to individual 

species or groups of species whose decline or removal 

may result in dramatic changes in the structure and 

functioning of its biological community (De Leo and 

Levin 1997).

In assessing the keystone role of an organism it is 

not always possible to distinguish clearly between the 

effects it may have as a keystone. For example, a good 
10 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

case could be made for considering large grazers as both 

keystone ecosystem engineers and as herbivores. By 

suppressing invasion of woody species and maintaining 

open grassland vegetation, large grazers profoundly 

alter the physical structure of the community. Even 

the classic keystone predator, the starfish, exerts its 

influence not just as a predator but as the incidental 

creator of bare patches, which provide recruitment 

sites for other species. Most organisms probably alter 

their environment in some way that can be exploited 

by other species. Whether the engineer can be considered a keystone species is a matter of degree: how 

profound the physical effect is, how disproportionate 

it is, how many other species benefit, and how great 

their dependence on the alteration is (Payton and 

others 2002).

Monitoring is a major issue and a need the manager 

has. There have been two basic monitoring approaches 

employed: single-species (for example, indicators 

(MIS), umbrellas, or flagships) or whole ecosystems 

(for example, focals). Each tactic has its advocates 

and detractors. Single-species monitoring focuses 

on one species. In contrast, ecosystem or landscapescale approaches emphasize ecological processes and 

habitats rather than individual species (Payton and 

others 2002).

It has been suggested that the keystone species concept may allow managers to combine the best features 

of single-species and ecosystem-based management 

approaches. Monitoring keystone species retains a 

single-species focus while avoiding the need to examine 

every species, and it emphasizes the mechanisms that 

directly rather than indirectly control biodiversity. 

Where it is possible to identify a keystone species 

that is critical for the continued survival (or demise) 

of many other species in its community, management 

of that keystone may be an efficient means of managing a much wider range of biodiversity (Payton and 

others 2002).

By themselves, keystone species are unlikely to provide a panacea for managers. Not all ecosystems may 

contain keystone species, and even where keystones 

are identified they may not be easily managed as part 

of a conservation strategy (Payton and others 2002).

Summary__________________________

During the 19th and early 20th centuries, people moving westward saw the grasslands as a land that could 

fulfill their dreams of new homes, new opportunities, 

and new freedoms. The settlement of the grasslands 

brought many changes that affected wildlife. These 

changes include conversion of wild habitats to agriculture and urban areas, and the dissecting of the 

landscape with roads, fences, and rails. This land 

conversion and dissection have resulted in habitat loss 

and fragmentation.

Another drastic change to the grassland has been 

the alteration or loss of riparian habitats and wetlands, 

affecting many wildlife species that use riparian areas 

or are riparian obligates. Introduction of farm and ranch 

livestock has imposed another change by subjecting 

native species to competition pressures that they had 

not evolved with. Natural predator-prey relationships 

have been disrupted as many predators were greatly 

reduced in number or extirpated because they prey on 

livestock. Conversely, populations of some prey species 

have been reduced, at times intentionally, because they 

were and are deemed as a pest and a nuisance.

Another change that has altered grassland ecosystems has been fire suppression. Suppression of fire 

has result in changes in the structure and species 

composition of the vegetation and, consequently, has 

influenced the wildlife species of the grasslands.

In light of the many, often dramatic, changes people 

have made to the grassland ecosystems mentioned 

above and which will be discussed in more detail in 

the following chapters, it is worth repeating the words 

of Aldo Leopold (1949: 190):

The last word in ignorance is the man who says 

of an animal or plant: “What good is it?” If the land 

mechanism as a whole is good, then every part is 

good, whether we understand it or not. If the biota, 

in the course of eons, has built something we like but 

do not understand, then who but a fool would discard 

seemingly useless parts? To keep every cog and wheel 

is the first precaution of intelligent tinkering.

Over time, many of the wheels and cogs (biological 

diversity, predator-prey relationships, competition, 

wild habitats, and so forth) in the grassland ecosystem 

have been changed, broken, or replaced, resulting in a 

system that does not work as well as it had originally 

evolved to work. Consequently, it is wise for land 

managers to work with the components of the system 

that are present and strive to have them working as 

well as possible.

The information in this volume is intended to help 

manage the wildlife and fisheries components of the 

grassland ecosystems.

Acknowledgments_________________

I thank Mark A. Rumble, Fred B. Samson, and John 

Sidle for reviewing this chapter and providing insightful 

comments. I express my appreciation to Deborah M. 

Finch for editing the chapter and to Nora C. Altamirano 

for helping prepare this chapter for publication.

References________________________

Baldwin, J.J. 1968. Chaparral conversion on the Tonto National 

Forest. Tall Timbers. Fire Ecology Conference. 8: 203-208
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 11

Baskin, Y. 1997. The work of nature: how the diversity of life sustains 

us. Natural History. 106(1):48-52. URL: [online] http://www.

anapsid.org/nature.html

Beck, A.M.; Vogl, R.J. 1972. The effects of spring burning on rodent 

populations in a brush prairie savannah. Journal of Mammalogy. 

5 3(2): 336-345.

Bell, R.H.V. 1971. A grazing ecosystem in the Serengeti. Scientific 

American. 225: 86-93.

Bendell, J.F. 1974. Effects of fire on birds and mammals. In: Fire 

and Ecosystems. T.T. Kozlowski and C.E. Ahlgren, eds. New York: 

Academic Press. 545 p.

Berg, W.A. 1986. Effect of 20 years of low N rate pasture fertilization 

on soil acidity. Journal of Range Management. 39: 122-124.

Bock, C.E.; Bock, J.H. 1978. Response of birds, small mammals 

and vegetation to burning sacaton grasslands in southeastern 

Arizona. Journal of Range Management. 31: 296-300.

Bock, C.E.; Bock, J.H. 1990. Effects of fire on wildlife in southwestern 

lowland habitats. In: Effects of fire management of southwestern 

natural resources. Gen. Tech. Rep. RM-191. Fort Collins, CO: 

U.S. Department of Agriculture Forest Service Rocky Mountain 

Forest and Range Experiment Station: 50-59.

Bock, J.H.; Bock, C.E.; McKnight, J.R. 1976. The study of the effects 

of grassland fires at the Research Ranch in southeastern Arizona. 

Arizona Academy Science. 11: 49-57.

Boeker, E.L.; Scott, V.E.; Reynolds, H.G.; Donaldson, B.A. 1972. 

Seasonal food habitats of mule deer in southwestern New Mexico. 

Journal of Wildlife Management. 36(1): 56-63.

Bradley, W.G.; Mauer, R.A. 1973. Rodents of a creosote brush community in southern Nevada. Southwestern Naturalist. 17(4): 

333-344.

Brown, D.E. 1978. Grazing, grassland cover and gamebirds. 

Transactions of the North American Wildlife Conference. 34: 

477-485.

Brown, D.E.; Lowe, C.H. 1980. Biotic communities of the Southwest. 

Gen.Tech.Rep. RM-78. Fort Collins, CO: U.S. Department of 

Agriculture Forest Service Rocky Mountain Forest and Range 

Experiment Station.

Cable, D.W. 1967. Fire effects on semi-desert grasses and shrubs. 

Journal of Range Management. 20(3): 170-176.

Clark, H.W. 1935. Fire and bird populations. Condor. 37:16-18.

Colbert, E. H.; Morales, M. 1992. Evolution of the vertebrates: A 

history of the backboned animals through time. New York: John 

Wiley & Sons. 470 p.

Cook, S.F. 1959. The effects of fire on a population of small rodents. 

Ecology. 40(1): 102-108.

Crist, T.; Veech, J.A.; Gering, J.C.; Summerville, K.S. 2003. 

Partitioning species diversity across landscapes and regions: a 

hierarchical analysis of �, �, and � diversity. American Naturalist. 

162(6): 734-743.

Daubenmire, R. 1968. Ecology of fire in grasslands. In: Advances 

in Ecological Research. Vol. 5. New York: Academic Press: 209-

273.

Davic, R.D. 2002. Herbivores as keystone predators. Conservation 

Ecology 6(2): r8. [online] URL: http://www.consecol.org/vol6/

iss2/resp8

Davies, N.B. 1979. Ecological questions about territorial behaviour. 

In: J. R. Krebs and N. B. Davies, eds. Behavioural ecology, an evolutionary approach. London, UK: Blackwell Scientific: 317-350.

De Leo, G.A.; Levin, S. 1997. The multifaceted aspect of ecosystem 

integrity. Conservation Ecology 1(1):3. [online] URL: http://www.

conecol.org/vol1/iss1/art3

Dills, G.G. 1970. Effects of prescribed burning on deer browse. 

Journal of Wildlife Management. 34(3): 540-545.

Ernest S.K.M.; Brown, J.H. 2001. Delayed compensation for missing 

keystone species by colonization. Science. 291(5514):101-104

Fischer, J.; Lindenmayer, D.B.; Cowling, A. 2004. The challenge 

of managing multiple species at multiple scales: reptiles in 

an Australian grazing landscape. Journal of Applied Ecology. 

41:32-44.

Fleischner, T.L. 1994. Ecological costs of livestock grazing in western 

North America. Conservation Biology. 8: 629-644.

Ford, P.L.; McPherson, G.R. 1996. Ecology of fire in shortgrass prairie of the southern Great Plains. In: D.M. Finch, ed. Ecosystem 

disturbance and wildlife conservation in western grasslands – A 

symposium proceedings, September 22-26, 1994; Albuquerque, 

NM. Gen. Tech. Rep. RM-GTR-285. Fort Collins, CO: U.S. 

Department of Agriculture Forest Service Rocky Mountain Forest 

and Range Experiment Station. 82 p.

Frank, D.A.; McNaughton, S.J. 1992. The ecology of plants, large 

mammalian herbivores, and drought in Yellowstone National 

Park. Ecology. 73: 2043-2058.

Geier, A. R.;Best, L.B. 1980. Habitat selection by small mammals of 

riparian communities: Evaluating effects of habitat alterations. 

Journal of Wildlife Management. 44(1): 16-24.

Gruell, G.E. 1980. Fire’s influence on wildlife habitat on the BridgerTeton National Forest, Vol. II. Gen. Tech. Rep. INT-252. Ogden, 

UT: U.S. Department of Agriculture Forest Service Intermountain 

Forest and Range Experiment Station. 35 p.

Hallisey, D.M.; Wood, G.W. 1976. Prescribed fire in scrub oak habitat 

in central Pennsylvania. Journal of Wildlife Management. 40(3): 

507-516.

Handley Jr., C.O. 1969. Fire and mammals. Tall Timbers Fire 

Ecology Conference. 9: 151-159.

Hanley, T.A.; Page, J.L. 1982. Differential effects of livestock use 

on habitat structure and rodent populations in Great Basin communities. California Fish and Game. 68: 160-174.

Hanson, E.E. 1978. The impact of a prescribed burn in a temperate subalpine forest upon the breeding bird and small mammal 

populations. M.S. thesis. Ellensburg, WA: Central Washington 

University.

Hastings, J.R.; Turner,R.M. 1965. The changing mile. Tucson: 

University of Arizona Press.

Heinselman, M.L. 1973. Fire in the virgin forests of the Boundary 

Waters Canoe Area, Minnesota. Quaternary Research. 3: 329-

382.

Howards, W.E.; Fenner, R.L.; Childs Jr., H.E. 1959. Wildlife survival 

in brush burns. Journal of Range Management. 12: 230-234.

Jain, S.; Krishna, S. 2002. Large extinctions in an evolutionary 

model: The role of innovation and keystone species. PNAS. 99(4): 

2055-2060.

Joern, A. 1986. Experimental study of avian predation on coexisting 

grasshopper populations (Orthoptera: Acrididae) in a sandhill 

grassland. Oikos. 46: 243 249.

Johnson, A.S. 1989. The thin green line: riparian corridors and 

endangered species in Arizona and New Mexico. In: G. Mackintosh, 

ed. In defense of wildlife: preserving communities and corridors. 

Washington, DC: Defenders of Wildlife: 35-46.

Kaufman, D.W.; Finck, E.J.; Kaufman, G.A. 1990. Small mammals 

and grasslands fires. In: S.L.Collins and L.L. Wallace, eds. Fire 

in North American Tallgrass Prairies, Norman and London: 

University of Oklahoma Press: 46-80.

Keay, J.A.; Peck, J.M. 1980. Relationship between fires and winter 

habitat of deer in Idaho. Journal of Wildlife Management. 44(2): 

372-380.

Keystone Species. 2004. [online] URL: http://www.bagheera.com/

inthewild/spot_spkey.htm

Keystone Species Hypothesis. 1969. [online] URL: http://www.

washington.edu/researchlpathbreakers/l969g.htm1

Komarek Sr., E.V. 1969. Fire and animal behavior. Tall Timbers 

Fire Ecology Conference. 9: 161-207.

Kotliar, N.B.; Wiens, J.A. 1990. Multiple scales of patchiness structure: a hierarchical framework for the study of heterogeneity. 

Oikos. 29: 253-260.

Kramp, B.F.; Patton, D.R.; Brady, W.W. 1983. The effects of fire 

on wildlife habitat and species. Run Wild Wildlife Habitat 

Relationships. Wildlife Unit Tech. Rep. Washington, DC: 

U.S.Department of Agriculture Forest Service. 29 p.

Lawrence, G.E. 1966. Ecology of vertebrate animals in relation 

to chaparral fire in the Sierra Nevada foothills. Ecology. 47(2): 

278-290.

Leopold, A. 1949. A Sand County almanac, and sketches from here 

and there. New York: Oxford University Press. 226 p.

Lowe, P.O.; Ffolliott, P.F; Dieterich, J.H.; Patton, D.R. 1978. 

Determining potential wildlife benefits from wildfire in Arizona 

ponderosa pine forests. Gen. Tech. Rep. RM-52. Fort Collins, CO: 

U.S. Department of Agriculture Forest Service Rocky Mountain 

Forest and Range Experiment Station. 12 p.

Lyon, L.J.; Crawford, H.S.; Czuhai, E.; Fredriksen,R.L.; Harlen, 

R.F.; Metz, L.J.; Pearson, H.A. 1978. Effects of fire on fauna: A 

state-of-knowledge review. Gen. Tech. Rep. WO-6. Washington, 

DC: U.S. Department of Agriculture Forest Service. 22 p.
12 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Lyon, L.J.; Marzluff, J.M. 1984. Fire’s effects on a small bird population. In: J.E. Lotan and J.K. Brown, comp. Fire’s effects on 

wildlife habitat-symposium proceedings. Washington, DC: U.S. 

Department of Agriculture Forest Service: 16-22.

Madden, L. 1995. Ecological effects of prescribed fire on prairie 

songbird communities. Abstract: International Conference and 

Training Workshop on Conservation and Ecology of Grassland 

Birds & 1995 Annual Meeting of the Association of Field 

Ornithologists. 26-28 October 1995. Tulsa, OK.

Marcot, B.G.; Murphy, D.D. 1996. On population viability analysis and management. In: R.C. Szaro and D.W. Johnston, eds. 

Biodiversity in managed landscapes theory and practice. New 

York and Oxford: Oxford University Press: 58-76.

Mazurek, E.J. 1981. Effect of fire on small mammals and vegetation on the Upper Sonoran Desert. M.S. thesis. Tempe: Arizona 

State University.

McCulloch, C.Y. 1969. Some effects of wildfire on deer habitat 

in pinyon-juniper woodland. Journal of Wildlife Management. 

33(4): 778-784.

McLellan, C.A.; Dobson, A.P.; Wilcove, D.S.; Lynch, J.F. 1986. Effects 

of forest fragmentation on new- and old-world bird communities: 

empirical observations and theoretical implications. In: Wildlife 

2000: Modeling habitat relationships for terrestrial vertebrates. J. 

Verner, M.L. Morrison, and C.J. Ralph, eds. Madison: University 

of Wisconsin Press: 305-314.

McNaughton, S.J. 1985. Ecology of a grazing system: the Serengeti. 

Ecological Monographs. 55: 259-294.

Miller, R.S. 1967. Patterns and process in competition. Advance 

Ecological Research. 4: 1-74.

Nelson, J.R. 1982. Relationships of elk and other large herbivores. In: 

J.W. Thomas and D.F. Toweill, eds. Elk of North America, ecology 

and management. Harrisburg, PA: Stackpole Books: 415-44.

Noss, R.F. 1992. Issues of scale in conservation biology. In: P.L. 

Fiedler and S. K. Jam, eds. Conservation Biology - The theory and 

practice of natural conservation preservation and management. 

New York: Chapman and Hall.

Noss, R.F.; Harris, L.D. 1986. Node, networks, and MUMs: preserving diversity at all scales. Environmental Management. Vol. 

10(3): 299-309.

Odum, F. P. 1971. Fundamentals of ecology. Philadelphia, PA: W.B. 

Saunders Co. 574 p.

Ohmart, R.D.; Zisner, C.D. 1993. Functions and values of riparian 

habitat wildlife in Arizona a literature review. Arizona Game and 

Fish Department. 213 p.

Paine, R.T. 1969. A note on trophic complexity and species diversity. 

American Naturalist 100: 91—93.

Parmenter, R.R.; MacMahon, J.A. 1988a. Factors influencing species 

composition and population sizes in a ground beetle community 

(Carabidae): predation by rodents. Oikos. 52: 350-356.

Parmenter, R.R.; MacMahon, J.A. 1988b. Factors limiting populations of arid-land darkling beetles (Tenebrionidae): predation by 

rodents. Environmental Entomology. 17: 280-286.

Payton, I.J.; Fenner, M.; Williams, G.L. 2002. Keystone species: the 

concept and its relevance for conservation management in New 

Zealand. Science for Conservation 203. 29 p.

Probst, J.R.; Crow, T.R. 1991. Integrating biological diversity and 

resource management. Journal of Forestry. February.

Renwald, J.D.; Wright, H.A.; Flinders, J.T. 1978. Effects of prescribed 

fire on bobwhite quail habitat in the rolling plains of Texas. 

Journal of Range Management. 31(1): 65-69.

Roppe, J.A.; Hem, D. 1978. Effects of fire on wildlife in a lodgepole 

pine forest. Southwestern Naturalist. 23(2): 279-288.

Rosenberg, K.V.; Rapheal, M.G. 1986. Effect of forest fragmentation 

on vertebrates in Douglas-fir forests. In: Wildlife 2000: Modeling 

habitat relationships for terrestrial vertebrates.

Rowe, J.S.; Scotter, G.W. 1973. Fire in the Boreal Forest. Quaternary 

Research. 3: 444-464.

Stoddart, L.C. 1987a. Relative abundance of coyotes, jackrabbits and 

rodents in Curlew Valley, Utah. Final Report, U.S. Department 

of Agriculture Project DF-931.07. Denver, CO: Denver Wildlife 

Research Center.

Stoddart, L.C. 1987b. Relative abundance of coyotes, lagomorphs, 

and rodents on the Idaho National Engineering Laboratory. Final 

Report, U.S. Department of Agriculture Project DF-931.07. Denver, 

CO: Denver Wildlife Research Center.

Su, J.C.; Debinski, D.M.; Jakubaushas, M.E.; Kindscher, K. 2004. 

Beyond species richness: community similarity as a measure of 

cross-taxon congruence for coarse-filter conversation. Conservation 

Biology. 18(1):167-173.

Thomas, D.B.; Werner, F.G. 1981. Grass feeding insects of the western 

ranges: an annotated checklist. Tech. Bull. Tempe: University of 

Arizona Agricultural Experiment Station. 243 p.

Thomas, J.W. 1979. Wildlife habitat in managed forests of the Blue 

Mountains of Oregon and Washington. Forest Service Agricultural 

Handbook 53. Portland, OR: U.S. Department of Agriculture 

Forest Service Pacific Northwest Forest and Range Experiment 

Station. 512 p.

Thomas, W.; Toweill, D.F. eds. 1982. Elk of North America, ecology 

and management. Harrisburg, PA: Stackpole Books.

Tomhack, D.F. 1986. Post-fire regeneration of krummholz whitebark 

pine: a consequence of nutcracker seed caching. Madrono. 33: 

100-110.

Urban, D.L.; O’Neill, R.O.; Shugart, H.H. 1987. Landscape ecology. 

BioScience. Vol. 37, No. 2: 119-127.

Vavra, M.; Mclnnis, M.; Sheehy, O. 1989. Implications of dietary 

overlap to management of free-ranging large herbivores. 

Proceedings Western Section American Society Animal Science. 

40: 489-495.

Williams, 0. 1955. Distribution of mice and shrews in a Colorado 

montane forest. Journal of Mammalogy. 36(2): 221-231.

Wilson, E.0. 1984. Biophilia, the human bond with other species. 

Cambridge, MA: Harvard University Press l7 p.

Winter, M. 1995. The impact of fire on Baird’s sparrow habitat. 

Abstract: International Conference and Training Workshop on 

Conservation and Ecology of Grassland Birds & 1995 Annual 

Meeting of the Association of Field Ornithologists. 26-28 October 

1995. Tulsa, OK.

Wirtz, W.O. 1977. Vertebrate post-fire succession. In: Proceedings 

of the symposium on the environmental consequences of fire and 

fuel management in Mediterranean ecosystems. Gen. Tech. Rep. 

WO-3. Washington, DC: U.S. Department of Agriculture Forest 

Service: 46-57 p.

Wisdom, M.J.; Thomas, J.W. 1996. Elk. In: Range Wildlife, 

Kausmann, P.R., ed. Denver, CO: The Society of Range 

Management. 440 p.

Wolfe, C.W. 1973. Effects of fire on a sand hills grassland environment. Tall Timbers Fire Ecology Conference. 12: 241-255.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 13

Chapter 2:

Large Native Ungulates

Bryce Rickel

Introduction_______________________

This chapter addresses the large native ungulates 

(American bison (Bos bison), elk (Cervus elaphus), 

white-tailed deer (Odocoileus virginianus), mule deer 

(Odocoileus hemionus), and pronghorn (Antilocapra 

americana) of the grasslands. The information presented includes historical background, description of 

the species’ biology and ecology, and management in 

relation to domestic animals.

American Bison____________________

Historically bison were widespread in North America 

from Alaska and Western Canada across the United 

States into Northern Mexico (fig. 2-1). They are currently found in isolated units throughout and external 

to historical range. Three separate subspecies are 

recognized: the Armican bison (Bos bison), the wood 

bison (Bos bison athabascae), and the bison (Bos bison 

bison).

The wood bison and bison are primarily distributed 

in Canada (NatureServe 2004). Population size of 

bison in North America may have been between 30 

million to 60 million about the time of Euro-American 

settlement and reduced to about 1,650 by 1903. The 

population in 1983 was estimated at 75,000 (Meagher 

1986). In Yellowstone National Park, the herd was 

estimated at 3,000 to 3,500 in 1996 (Keiter 1997); 

however, more than 1,000 were killed during the winter 

of 1996 to 1997 by agency personnel (NatureServe 

2004). Populations of American bison have fluctuated 

dramatically during the 11,000 years of the present 

interglacial (the Holocene) and probably were at their 

peak in the Great Plains when Europeans first arrived 

in the 16th century (Parmenter and Van Devender 

1995). A prolonged drought in the mid-19th century in 

concert with greater hunting pressures from Native 

American residents and American settlers and increasing competition with domestic livestock (notably 

horses and cattle) for riparian winter grazing lands, 

brought the bison to the brink of extinction. Today, 

Bison on prairie dog site. (Photo by Paulette Ford.)
14 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

there are approximately 30,000 bison in North America 

(Parmenter and Van Devender 1995). Bison currently 

play only a small role in the ecology of the grasslands 

of the Southwestern Region. They are discussed here 

because historically they had influence on some of the 

Great Plains grasslands and possibly on the desert 

grasslands (fig. 2-2) in the Region and because of their 

relationships with other wildlife species. They are also 

discussed because of their role as herbivores and the 

comparison with cattle.

Bison were common on the eastern plains of New 

Mexico in early historic times, although they did 

not inhabit the Rio Grande Valley or country to the 

west. Formerly abundant, and an important source 

of meat for Pueblo peoples, bison were essentially 

gone from New Mexico by 1860. Strips of buffalo skin 

kept among the Taos Indians were highly prized and 

were still being used primarily for decoration and 

ornamentation. Supposedly at this time, a fine buffalo 

robe would be saved by one of the Indian men for his 

burial. As with many Pueblo Tribes, the Taos Indians 

had among them a buffalo clan. The buffalo dance was 

one of the principal dances of the tribe, and it was set 

to music of drum and voice to represent the low hum 

of the grazing herd, and the thunder of a stampede 

(BISOM-M 2005). Unconfined herds of bison in the 

early 19th century moved over extensive areas of the 

Great Plains. Historical evidence shows that bison 

were present year round on short-grass, tallgrass, and 

mixed-grass regions. Herd movements were localized 

and flexible, indicating that movements were in 

response to changing forage quantity and quality, 

making regular, long-range migrations unnecessary 

(Shaw 1996).

Currently in the Southwestern Region, there are 

no large herds of bison. Where they do exist, the herds 

are comparatively small and are reintroductions or 

transplants managed for hunting or recreation or 

to maintain the species.

How Bison Modify Habitat

Bison have been identified as keystone species 

in the Great Plains (Knapp and others 1999). 

As a keystone species, bison had a principal role 

in the formation and maintenance of grassland 

ecosystems.

Grazing—Bison are primarily graminoid feeders 

(Shaw 1996, Wisdom and Thomas 1996) and avoid 

forbs and woody species. Thus, within a bison grazing area, forbs are often conspicuously left ungrazed 

and are surrounded by grazed grasses (Knapp and 

others 1999).

Bison and large herbivores influenced landscapes at 

multiple scales. Therefore, in their planning, managers 

should consider the role these large herbivores had 

historically on grasslands. At broad scales, watershed 

and landscape, the long-term consequences of bison 

activities included cover reduction, and changes in 

dominance of plant species and productivity of grasses. 

Bison grazing may result in a competitive release 

of many subdominant plant species resulting in an 

increase in the abundance of forbs, an overall increase 

in plant species richness and diversity, and increased 

spatial heterogeneity (Hartnett and others 1996).

Although alterations in plant community composition can be attributed to the direct effects of grazing 

by bison, increased plant species richness is also likely 

to be a product of increased microsite diversity generated by nongrazing activities, such as dung and urine 

deposition, trampling, and wallowing. These and other 

bison activities contribute significantly to the increase 

in the spatial heterogeneity that is characteristic of 

grazed tallgrass prairie (Knapp and others 1999).

Bison use of grasslands alters nutrient cycling 

processes and patterns of nutrient availability. Their 

effects on nitrogen cycling are critical because nitrogen 

availability often limits plant productivity in these 

grasslands and influences plant species composition 

(Knapp and others 1999). Bison and other ungulates 

in grasslands consume plant biomass that is difficult 

to digest and return labile forms of nitrogen (that is, 

urine) to soils (Knapp and others 1999). Nitrogen in 

bison urine is largely urea, which can be hydrolyzed 

to ammonium in a matter of days, facilitating nitrogen 

cycling in the grasslands (Knapp and others 1999). The 

Figure 2-1. Map of the historic range of bison (NatureServe 

2004).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 15

net effect of bison grazing on nutrient cycling appears 

to be increased rates of nitrogen cycling, coupled with 

an increase in spatial heterogeneity in nitrogen availability (Knapp and others 1999).

Are Bison and Cattle Functional Equivalents 

in Tallgrass Prairie?—Studies that have focused 

exclusively on cattle generally concur that their grazing activities can increase spatial heterogeneity and 

enhance plant species diversity, as long as stocking 

density is not too high (Hartnett and others 1996). 

Because bison grazing in tallgrass prairie has a similar 

effect, one could conclude that either herbivore can 

alter resource availability and heterogeneity and 

reduce the cover of the dominant grasses sufficiently 

to enhance the success of the subdominant species. 

Perhaps of greater importance than differences in 

foraging patterns between bison and cattle, however, 

are the numbers of nongrazing activities, such as 

wallowing and horning (rubbing on trees) that are 

associated exclusively with bison (Hartnett and others 

1996). These activities, when combined with the spatial 

redistribution of nutrients and selective consumption 

of the dominant grasses, may further increase plant 

species richness and resource heterogeneity, particularly at the landscape scale.

Nevertheless, it is likely that because bison and 

cattle are functionally similar as large grass-feeding 

herbivores, management strategies (stocking intensity and duration) will have a greater influence on 

the degree of ecological equivalency achieved than 

inherent differences in these ungulates (Hartnett 

and others 1996). The degree of overlap in diet and 

foraging patterns is greater between bison and cattle 

than between cattle and other historically important 

native herbivores (Hartnett and others 1996) such 

as antelope, deer, and elk.

The important elements of bison grazing 

activities can and should be incorporated into 

conservation and restoration strategies for remnant 

prairies (Knapp and others 1999). One approach 

to accomplish this goal is the substitution of cattle 

for bison. In Knapp and others (1999), an argument was presented that the choice of whether 

to use cattle or bison as a management tool in 

grasslands is scale- and context-dependent. Clearly, 

reintroducing bison may not be appropriate for 

small prairie remnants with public access and low 

economic resources. But cattle, managed for their 

ecological rather than their economic value, may 

be suitable in such cases.

Relationship of Bison with Other 

Herbivores

Elk have such wide feeding niches that they use 

some of the same forages as bison (Shaw 1996). 

Diets of domestic sheep overlap partially with those 

of bison (Shaw 1996). Diets of mule deer and whitetailed deer do not converge with those of bison under 

normal conditions. Deer favor browse, forbs, and 

mast, whereas bison concentrate on graminoids. Deer 

favor woodier habitats while bison more frequently 

use open meadows and grasslands. Pronghorn share 

common ranges and habitats with bison but have 

divergent diets. Pronghorns eat almost exclusively 

forbs and browse. Horses forage on a wide range of 

plant foods, and their diets can overlap with those of 

bison (Shaw 1996).

Bison in the Desert Grasslands

The role bison played in desert grasslands is not 

clear. We do know that they occurred in the Southwest, 

but to what extent is still a question. An extinct 

bison (Bison antiquus) with massive widespread 

horns was prevalent in Arizona and Sonora, Mexico, 

during the last Ice Age, the Wisconsin Glacial Period 

(Parmenter and Van Devender 1995). There have 

been several archaeological findings that indicate 

bison where in the Southwest (Parmenter and Van 

Devender 1995).

Figure 2-2. Map showing the historic range of bison overlaid 

by the States within the USDA Forest Service Southwestern 

Region, and the status of bison within each State (NatureServe 

2004) and the National Forests within the Region.
16 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

• The last confirmed hunting record of native bison 

in eastern New Mexico was in 1884 (Findley and 

others 1975).

• Bison horn cores, teeth, and bones were recovered 

in the excavation of Snaketown, a Hohokam settlement in the Gila River valley in Pinal County 

south of Phoenix that was occupied from before 

the birth of Christ until A.D. 1200 (Parmenter 

and Van Devender 1995).

• Bison bones associated with the years A.D. 1200 

and 1380 were identified from two rooms in a 1968 

excavation of the Hohokam Las Colinas site in 

Maricopa County near Phoenix (Parmenter and 

Van Devender 1995).

• Bison bones, some of them painted, found in a 

Babocomari Village excavation, were dated to A.D. 

1200 through 1450 (Parmenter and Van Devender 

1995). The Babocomari River, a tributary of the 

San Pedro River, flows through desert grassland 

on the north end of the Huachuca Mountains 

east of Elgin in Cochise County, Arizona. In the 

1970s, a bison skull was found eroding out of 

sediments in the same area (Parmenter and Van 

Devender 1995).

• At Murray Springs, in the desert grasslands in 

the San Pedro River valley near Sierra Vista on 

the eastern base of the Huachucas in Cochise 

County, bones of extinct bison were found. This 

desert grassland site was likely dominated by 

grasses prior to 1890 but is presently covered with 

velvet mesquite (Prosopis velutina). At the same 

excavation, the skeleton of a female bison with a 

near-term fetus inside was discovered (Parmenter 

and Van Devender 1995) and radiocarbon-dated 

to A.D. 1700.

Not only were bison an archeological part of the 

desert grassland fauna, they were present according to 

historical accounts at least into the 17th and even into 

the 19th century. Parmenter and Van Devender (1995) 

present a postulate that the bison expanded its range 

into Arizona from the desert grassland valleys of northwestern Chihuahua during cooler, moister climatic 

fluctuations. It is a mystery why bison disappeared 

from desert grasslands before Euro-Americans began 

recording their observations in the 1820s (Parmenter 

and Van Devender 1995).

Bison have been introduced into Arizona and New 

Mexico since the early 1900s (Chung-MacCoubrey 1996, 

Findley and others 1975). Several small bison herds 

now exist in the desert grasslands of the Southwest.

Elk________________________________

Elk are an important herbivore in North American 

rangelands. Their large body size, herding behavior, 

pioneering habits, and high mobility make them a 

conspicuous wildlife species on open grasslands. The 

potential for elk to compete with livestock makes them 

an obvious source of controversy between ranchers, 

farmers, and wildlife advocates (Wisdom and Thomas 

1996).

Elk use a broad spectrum of habitats including 

alpine, conifer, hardwood and mixed forests, grasslands, savanna, and shrubland/chaparral. They often 

frequent open areas such as alpine pastures, marshy 

meadows, river flats, and aspen parkland, as well as 

open meadows in coniferous forests, brushy clear cuts 

or forest edges, and semidesert areas. On more level 

terrain, they seek wooded hillsides in summer and 

open grasslands in winter. No special calving ground 

is typically used; calves are born in valleys or in areas 

as high as alpine tundra. Newborn initially may be 

hidden in rough terrain or dense cover.

Historical Distribution of Elk

The elk that live in North America today are direct 

descendants of red deer that migrated from Asia 

approximately 120,000 years ago. Currently, most 

biologists consider all the elk in North America to be 

the same species as the red deer in Asia and Europe 

(Rocky Mountain Elk Foundation 2004). Before 

European settlement, elk were among the most common and widely distributed of the wild ungulates in 

North America (Shaw 1996, Wisdom and Thomas 1996, 

Yoakum and others 1996). Historically, six subspecies 

of elk inhabited areas from the Atlantic to the Pacific 

Coasts. The Eastern elk (Cercus elaphus Canadensis 

extinct) once lived in Ontario, southern Quebec, and 

over much of the Eastern United States. Roosevelt 

elk (Cervus elaphus roosevelti) are found in the Pacific 

Northwest coastal forests of northern California, 

Elk photo by Lane Eskew.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 17

Oregon, Washington, and Vancouver Island. The smallest of the subspecies is the Tule or California elk (Cervus 

elaphus nanodes), which once lived in large numbers 

in California’s San Joaquin and Sacramento Valleys 

and has adapted to semidesert conditions. The Rocky 

Mountain elk (Cervus laphus nelsoni) ranges from New 

Mexico to Northcentral British Columbia and through 

the Intermountain region. It has been introduced 

into many places in North America and elsewhere 

in the world. Merriam elk (Cevus elaphus marriami

extinct) once lived in the mountains and chaparrals of 

western Texas, New Mexico and mountains of Arizona. 

Manitoban elk (Cervus elaphus manitobensis) ranges 

from the southern Prairie Provinces of Canada to North 

Dakota. (fig. 2-3) (North American Elk Subspecies 

2004, Rocky Mountain Elk Foundation 2004, Wisdom 

and Thomas 1996).

Controversy exists on whether elk once occupied 

rangelands of the Great Basin and Desert Southwest. 

This controversy centers on recent transplants of elk 

and/or recent colonization by elk into these areas 

(Yoakum and others 1996) and the potential for competition with livestock.

According to Yoakum and others (1996) and Wisdom 

and Thomas (1996), elk originally occupied most of 

the grasslands throughout Western North America, 

including major areas of the Great Basin and Desert 

Southwest. However, lack of permanent water likely 

restricted elk distribution in the driest portions 

of these regions, and these dry areas presumably 

are outside the historical range of the species (fig. 

2-3). This agrees with evidence that large, wild 

ungulates historically were absent from or sparsely 

distributed across large areas of the Intermountain 

West (Wisdom and Thomas 1996).

Population Declines and Extirpations

Shaw (1996) estimated numbers of North 

American elk at 10 million before the arrival of EuroAmerican settlers. By contrast, with the exception 

of the Rocky Mountain and Roosevelt subspecies in 

the West, elk are now distributed in small, disjunct 

populations across much of their former range 

(Wisdom and Thomas 1996) (fig. 2-3).

Elk numbers declined as Euro-American settlers 

moved west in the 17th, 18th, and 19th centuries. By 

the early 1800s, the eastern subspecies was likely 

extinct (Wisdom and Thomas 1996). Populations 

of Tule elk, once abundant in valleys and bottomlands of California, were reduced to approximately 

100 animals by 1875 (Wisdom and Thomas 1996). 

The Merriam subspecies was extirpated from the 

Southwest by 1906 (Wisdom and Thomas 1996). By 

the early 1900s, Manitoban elk of the northern prairies were close to extinction (Wisdom and Thomas 

1996). During the same period, populations of Rocky 

Mountain and Roosevelt elk were reduced to scattered, 

isolated pockets within their former range (Wisdom 

and Thomas 1996). Unregulated hunting, overgrazing 

by livestock, and conversion of habitat to agriculture 

and city land led to these broad-scale declines and 

extirpations (Parmenter and Van Devender 1995, 

Wisdom and Thomas 1996).

Current Abundance and Distribution

In 1910, an estimated 60,000 elk remained in 

North America, with fragmented populations present 

in seven Western States. By this time, most States 

had enacted legislation that prohibited the hunting 

of elk. Then, efforts to recover elk populations were 

initiated through programs of translocation. From 

1892 to 1939, more than 5,000 elk were live-trapped 

and transported from the Yellowstone area for release 

in 36 States and parts of Western Canada (Wisdom 

and Thomas 1996).

By the 1930s, elk numbers and distribution increased substantially, with the population estimated 

at 165,764 by 1937 and 236,787 by 1941. In the 1960s, 

elk populations had grown to approximately 440,000 

and increased to 500,000 by the late l970s and 600,000 

by 1987 (Wisdom and Thomas 1996).

Today an estimated 700,000 elk exist in North 

America. Populations are generally increasing and 

Figure 2-3. Historic distribution of elk in North America (Bryant 

and Maser 1982, in Wisdom and Thomas 1996) and the current 

distribution (NatureServe 2004).
18 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

translocations continue. Elk are present in at least 22 

States (fig. 2-3). Native predators, particularly wolves 

(Canis spp.), have been extirpated throughout the 

Southwestern range of elk except in limited reintroduction areas, and therefore, elk numbers are no longer 

naturally regulated. Hunting has replaced predation 

as a means for maintaining elk populations.

Elk on Public Versus Private Lands

Before Euro-American settlement, elk often made 

expansive seasonal movements in response to annual 

changes in weather and forage availability (Wisdom 

and Thomas 1996). Such movements are not possible 

today. Many seasonal habitats used historically by elk 

are now privately owned and managed exclusively for 

agriculture, livestock, and timber production (Wisdom 

and Thomas 1996). Elk use of these areas is not tolerated or tolerated only minimally. In other cases, historic 

habitats have been converted to cities, industrial developments, and recreation areas (Wisdom and Thomas 

1996). Elk now reside in “ecologically incomplete” or 

“ecologically compressed” habitats across much of their 

historic range (Wisdom and Thomas 1996).

Patterns of land ownership in the Western United 

States exemplify the problem. In the West, more than 

90 percent of elk use public lands during summer 

while they use privately owned lands for the majority of their winter range. Large numbers of elk are 

produced on publicly owned summer range, where 

they are valued for hunting and viewing. Elk then 

migrate to privately owned ranges at lower elevations during the winter. On such lands, elk often 

encounter severe shortages of habitat features and 

active programs of population control to minimize 

competition with livestock and damage to crops 

(Wisdom and Thomas 1996).

Efforts to control elk damage and consumption 

of crops and cattle forage on privately owned lands 

are intense and varied. Activities include providing 

landowners elk hunting permits to sell to hunters, government financial compensation for loss of 

revenue due to damages caused by elk, permission 

to landowners to kill “nuisance” elk, special hunts, 

fencing to exclude animals from private ranges, and 

leasing or acquisition of additional winter range to 

maintain winter carrying capacity.

The shortage of year-round, ecologically complete 

habitats for elk and other wild ungulates poses significant management problems on North American 

rangelands. Many of today’s management problems 

can be attributed to fragmented ownerships and the 

associated differences in land use goals between public 

and private lands (Wisdom and Thomas 1996).

Other Economic and Social Values

Elk hunting and viewing are highly valued experiences, both aesthetically and economically. New sources 

of revenue from consumptive and nonconsumptive 

users of elk and other wildlife, combined with growing 

political influences of such users, are causing shifts in 

economic and social values that affect management of 

public lands (Wisdom and Thomas 1996).

Elk-Livestock Interactions

Grazing by elk and livestock is perhaps the most 

significant land use issue affecting management of both 

species on public and private lands. Issues of forage 

allocation and competition are subjects of unending 

debates (Wisdom and Thomas 1996).

In relation to dietary overlap between elk and cattle, 

the following 12 generalizations can be made (Wisdom 

and Thomas 1996):

• The potential for competition between elk and 

cattle is highest on winter and spring-fall ranges 

where either forage quantity or quality is limited 

and where both species can commonly share “ecologically compressed habitats” on low-elevation 

bottomlands or foothills.

• Competition between elk and cattle is usually low 

on high-elevation summer ranges where forage 

of moderate to high quality is readily available 

during late spring and summer and where animals 

have a more expansive land base from which to 

make optimal grazing choices with carryover 

effects into winter and spring.

• The potential for competition between elk and 

cattle increases during late summer and fall on 

high elevation summer ranges following the onset 

of prolonged seasonal drought and the subsequent 

decline in forage quality.

• Elk and cattle often distribute themselves spatially in a manner that minimizes competition.

• Elk and cattle can distribute themselves temporally (seasonally) in a manner that minimizes or 

heightens competition between the two.

• On ranges where spatial or temporal distributions 

overlap, differences in the diets of elk and cattle 

can sometimes minimize competition.

• The potential for competition between elk and 

cattle is high on unproductive rangelands, 

especially in arid ecosystems. Potential for 

competition also is high on rangelands grazed 

to full or maximum use by elk or cattle, and on 

rangelands experiencing a declining trend in 

condition.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 19

• High densities of elk may induce negative effects 

on forage conditions similar to that caused by high 

stocking rates of cattle, resulting in lower animal 

performance. Intraspecific competition among 

elk, and resultant negative effects on forage, 

likely are density-dependent and manifested at 

high population densities.

• Elk show an aversion to the presence of cattle that 

may or may not restrict their grazing choices.

• On productive rangelands that were grazed historically by native herbivores, systems of cattle 

grazing can be designed to enhance forage or 

foraging conditions for elk. Likewise, grazing by 

elk can enhance conditions for cattle.

• Competition can be high at a given time and place 

during a year and low or nonexistent in the same 

place and time in subsequent years.

• Perception is rarely reality when judging competitive interactions between elk and cattle.

Elk and other Native Large Herbivores

Distributional overlap can occur with mule and whitetailed deer, bison, pronghorn, and bighorn sheep, which 

apparently partition forage and habitat resources to 

minimize competition in areas of overlapping distribution such as Yellowstone National Park (Shaw 1996). 

Often, patterns of foraging and habitat use are complementary. Grazing by elk can enhance forage conditions 

for pronghorn and mule deer (Shaw 1996).

Regardless, competition does occur in some situations 

(Wisdom and Thomas 1996). In contrast to mule deer 

and bighorn sheep, elk generally are more opportunistic 

and varied in their diet selection. Elk more easily digest 

forage of low quality, are more mobile and wide-ranging, and form larger herds, all of which may result in a 

competitive edge when resources are limited, presuming 

that elk out-compete mule deer on winter ranges that 

are limited in size and forage availability. Wisdom and 

Thomas (1996) also believed that mule deer would leave 

or avoid areas of heavy use by elk, even if forage was 

available and dietary overlap with elk is low.

Grazing Management Trade-Offs Between 

Elk and Livestock

Type of Grazing System—Little data exist about 

elk response to various systems of livestock grazing 

because there are few tests of elk response under 

experimental replication. Most research focuses on 

the effect of livestock grazing. The information that is 

available is inconclusive regarding elk preference for 

or aversion to any particular grazing system. Table 

2-1 summarizes how elk respond to various grazing 

schemes.

Effect of Stocking Rate—Stocking rate of livestock, more than any other grazing variable, influences 

the composition, quantity, and nutritive value of forage 

ungulate (Wisdom and Thomas 1996). This premise, 

combined with the potential for forage competition 

between elk and cattle, provides a framework for 

describing the effect of stocking rate on forage conditions for elk:

• Regardless of the grazing system used, the composition, quantity, and quality of forage available to 

elk is determined largely by the stocking rate of 

cattle, interacting with the density of elk, the inherent characteristics of the site, and weather.

• Light stocking rates of cattle can be neutral or 

positive in their influence on the composition and 

nutritive value of forage for elk.

• Moderate stocking rates of cattle, with allowable 

use of key species between 25 and 60 percent, can 

be positive, neutral, or negative in their effect on 

elk forage. Much depends on the timing of grazing, 

the physiological response of key forage plants 

to grazing, and the inherent productivity of the 

range and its condition.

• On relatively unproductive rangelands of the 

Great Basin and Desert Southwest, it is likely that 

Table 2-1. How elk respond to various grazing regimes (Wisdom and Thomas 1996).

Grazing system Result

Season-long, deferred-rotation Elk preferred ungrazed areas and showed no difference in their use of 

(in northeastern Oregon) season-long versus deferred-rotation pastures.

Season-long, deferred-rotation Higher elk use under season-long than deferred-rotation grazing with cattle 

grazing (in northeastern Oregon) light stocking. The opposite preference was found when stocking was high.

Rest-rotation grazing, season-long Elk preferred pastures in rest and avoided pastures actively grazed by cattle. 

grazing (in Montana) However, overall use by elk was not significantly different under rotation grazing 

 than that observed under season-long grazing.

Season-long, rest-rotation grazing No changes were found in population trends of elk following a change from 

(in east-central Idaho) season-long to rest-rotation grazing. It was concluded that elk preferred pastures 

 in rest and avoided those actively grazed by cattle.
20 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

cattle use more than 25 percent will negatively 

affect forage conditions for elk.

• High stocking rates of cattle invariably heighten 

the potential for competition with elk for limited 

forage. Similarly, high densities of elk may induce 

intraspecific competition for forage (Shaw 1996, 

Wisdom and Thomas 1996), magnifying potential 

competition with cattle.

• Regardless, cattle use of key species at levels 60 

percent or higher may significantly reduce or 

eliminate key forage plants for elk and cattle on 

nearly all range types (Wisdom and Thomas 1996). 

The remaining forage will likely be abundant 

but of low nutritive value, or scarce but of high 

nutritive value.

• On desert rangelands, both low quantity and 

quality of vegetation may result from overgrazing, negatively affecting both elk and cattle. 

These effects on elk forages may explain the 

inverse relationship between stocking rates 

of cattle and habitat use by elk (Wisdom and 

Thomas 1996). This relationship generally holds 

true for all types of grazing systems that have 

been studied.

Water Developments—Water is considered limiting to elk on many arid and semiarid rangelands. Elk 

may concentrate near water sources in extremely dry 

areas. Increasing the distribution and availability of 

water on many of the driest grasslands will likely 

enhance elk use of such areas, especially during dry 

seasons or years (Wisdom and Thomas 1996).

However, water development can be a “double-edged 

sword” if livestock have access to the water. Livestock 

use is usually highest within 1.6 km of water; this is 

the zone of most direct competition between elk and 

livestock. Improving the distribution of water—that 

is, making water more evenly and readily available 

throughout a pasture—will also result in a more 

even distribution of livestock (Wisdom and Thomas 

1996) when all things are equal. This may increase 

the potential for competition with elk, or reduce elk 

use in favor of cattle use for two reasons: First, most 

water developments have road access; elk avoid areas 

near roads. Road densities and traffic will likely to 

increase with an increase in number and distribution 

of water developments. Elk are especially vulnerable 

to human harassment during hunting seasons when 

roads facilitate hunting (Wisdom and Thomas 1996). 

The second reason is that elk avoid livestock but return 

to grazed areas after livestock have left. A more even 

distribution of livestock, resulting from a more even 

distribution of water, may hinder grazing choices by 

elk unless livestock-free areas also are available within 

a herd’s seasonal home range.

Roads—Wherever elk occur, they consistently and 

dramatically avoid areas near roads that are open to 

motorized traffic (Perry and Overly 1977, Wisdom and 

Thomas 1996). Implications of this relationship cannot 

be underestimated. All of the positive elk management that could be accomplished could be partially 

or wholly offset by the negative effects of open roads. 

This is especially true in areas of open grasslands 

and gentle terrain; here, vegetation and topography 

provide limited hiding and security from human 

activities. Aggressive programs of road management, 

obliteration and closure, is necessary to facilitate elk 

use of grasslands.

Conflict Resolution of Elk-Livestock 

Problems

Ecosystem Management of Elk—Elk are mobile, 

adaptive, and opportunistic; they use a variety of habitats that span many jurisdictions and land ownerships. 

Their annual movements can encompass hundreds of 

square kilometers, bringing them into contact with a 

multitude of landowners and land uses.

At the same time, human development continues unabated on elk ranges. Traditional users of rangelands, 

such as farmers and livestock growers, are generally 

intolerant of high elk numbers. Hunters, wildlife 

viewers, and urbanites value these same rangelands, 

desiring more elk for hunting and watching. Given 

these conditions, elkland use conflicts will continue to 

grow in frequency and intensity.

If polarized interest parties can adopt an ecosystem 

or a landscape approach to elk management, all of 

them will benefit. The geographic scale must change 

from that of an individual pasture, watershed, landowner, or allotment to that of an entire herd range, 

such as 10,000 ha or more. Within each herd range, 

partnerships such as those under way in Colorado 

(Wisdom and Thomas 1996) must be forged. Such 

partnerships are politically effective and provide an 

ecological basis upon which all interested parties can 

plan and implement the proper combination of land 

treatments and hunting regimes necessary to achieve 

the desired demography and distribution of elk across 

all land ownerships in time and space.

Deer______________________________

White-Tailed Deer

Of all the big game animals in North America, the 

white-tailed deer is the most widespread and numerous 

member of the deer family, Cervidae. It is also the most 

important big game species recreationally and economically. It occurs in all States except California, Nevada, 

and Utah (fig. 2-4) (Southeastern Cooperative Wildlife 

Disease Study 1982). There are 38 races or subspecies 

ranging from southeastern Alaska through parts of 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 21

southern Canada into practically all of the contiguous 

United States and through Central America and as far 

south as northern South America (Teer 1996).

The ecology and management of the different races 

of white-tailed deer are similar across the various 

ecosystems they inhabit. Therefore, management is 

habitat- and herd-specific and must be adapted to the 

vegetation type and other elements of their habitat.

Unlike those in forested habitats, white-tailed deer 

numbers in the grasslands in the western United 

States have increased as a result of encroachment of 

woody vegetation and agricultural fields. In the Great 

Plains, deer were associated primarily with riparian 

systems but began to extend their ranges into upland 

habitats in the 1940s. In prairie-agricultural habitat 

in east-central Montana, white-tailed deer selected 

riparian habitat and its interspersion with cropland 

and rangeland habitats (Teer 1996). White-tailed deer 

on the lower Yellowstone River are associated with 

grassland (Teer 1996) and use adjacent farmland. 

Their numbers are positively related to density of 

riparian cover.

Habitat—White-tailed deer are extremely adaptable. They thrive in close association with humans 

and their agricultural and industrial developments. 

The requirements of white-tailed deer are met in 

practically every ecological type, including grasslands, prairies and plains, mountains, hardwood, 

coniferous and tropical forests, deserts, and even 

farmland where it associates with woodlots and 

riverine habitats (Teer 1996).

Woody vegetation used by deer for cover and often 

for browse and mast is not an absolute requirement 

of deer habitats, although white-tailed deer are 

most often associated with brushlands and forested 

habitats. The species does occur in low densities 

in open pasturelands and grasslands. White-tailed 

deer in open grassland habitats occur in areas 

having screening or protective cover nearby for 

escape. Cover in such habitats may be herbaceous. 

Woody cover may be available in riparian zones 

associated with grasslands and plains.

White-tailed deer are also reported to use hardwood draws that often surge down into grasslands 

and mesic shrublands (Teer 1996). They often use 

these habitats during summer.

Management of Livestock for White-Tailed 

Deer—White-tailed deer and livestock are compatible and even synergistic, if livestock are managed 

for deer. Livestock can change the quantity and 

quality of forage by affecting erosion processes and 

nutrient losses, and consequently, even change 

the quality of the soil on which the feed is produced 

(Teer 1996). No other factors in deer ecology and 

management, outside of poaching and the plow, have 

influenced the habitat and productivity of deer on 

rangelands more than livestock (Teer 1996).

Traditionally, the economic worth of livestock made 

them the animals of choice, the priority, in grasslands. 

Livestock were often stocked in combinations on 

common-use ranges where browse, forbs, and grasses 

provide preferred foods. Until recently, white-tailed 

deer and other wildlife species were not considered in 

livestock management. Changes that have occurred in 

management of rangeland for wildlife have been the 

result of commercialization of hunting, recreation, and, 

more recently, the growing interest in wildlife viewing. The changes from traditional uses of grasslands, 

particularly of public lands, are due to environmental 

groups challenging many of the traditional uses in 

favor of management of wildlife.

Competition for Resources—Competition 

between deer and domestic livestock is important 

when populations are dense. Sheep and goat competition with deer can be especially severe because the 

seasonal and dietary overlaps of these species are 

almost 100 percent. Cattle are less competitive with 

deer because their diets are primarily grasses and 

grasslike plants, whereas, deer use grasses sparingly 

(Teer 1996).

Figure 2-4. White-tailed deer distribution in North America 

(NatureServe 2004).
22 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Good deer habitat consists of diverse vegetation. 

Availability of seasonal food sources is imperative in 

satisfying the whitetail’s nutritional requirements. 

White-tailed deer reach their greatest numbers in 

rangeland communities where seral stages are below 

climax. Thus, disturbed land is often the best deer 

range because the habitat contains an array of annual 

and perennial herbaceous plants that serve as food, 

and woody plants provide food and protection from 

enemies and weather (Teer 1996).

Behavioral Relationships with Livestock—It 

is not clear if whitetails avoid cattle in pastures that 

are grazed. Deer on the Welder Wildlife Foundation in 

southern Texas avoided pastures stocked with cattle. 

Deer did not alter their home ranges but chose pastures in unstocked portions of their range. Avoidance 

was mostly in heavily stocked, short-duration grazing 

pastures. Similar avoidance occurred with greater 

distances between deer and cattle than between deer 

and other cervids (Teer 1996).

Whitetails ignored cattle and commonly grazed 

with them. They also ignored riderless horses but 

avoided horses with riders. Deer avoided pastures 

where forage supply was decimated by heavy grazing (Teer 1996). Whitetails do not appear to avoid 

livestock when stocking is low or moderate. Deferred 

rotation pastures contained deer in higher numbers 

than pastures grazed more intensively.

In the past, criteria used by range and wildlife 

managers to judge carrying capacity or quality of the 

range for livestock had little relationship to quality 

of white-tailed deer habitat or other wildlife species’ 

habitats. Great differences occur in food and cover 

requirements of various kinds of domestic animals 

and deer, and these differences were largely ignored 

in classification systems.

Several researchers have attempted to determine 

forage relationships between livestock and whitetails. 

Animal unit equivalents proposed on the Edwards 

Plateau of Texas were five sheep, or six goats, or six 

white-tailed deer to one 489 kg cow with calf at side 

(Teer 1996). Teer (1996) reported that 13 deer were 

equivalent to one cow on the King Ranch in southern 

Texas. These ratios are viewed as rough standards 

for interpreting stocking rates and for developing diet 

relationships among various herbivores.

The impact of stocking rates of livestock on whitetailed deer is a two-pronged relationship. Overgrazing 

is obviously harmful to deer habitat, and undergrazing 

can also reduce its quality. Light to moderate grazing 

promotes plant vigor and diversity and increases 

production (Teer 1996). When grasses dry, cattle will 

switch to woody plants, which affect the quality of 

browse for white-tailed deer.

In addition to the effects grazing may have on the 

quality of white-tailed deer habitat, ranching activities also have an influence. Hood and Inglis (1974) 

found roundups disturbed white-tailed deer. Bucks 

reacted to roundups with long flights to adjacent 

pastures and, at times, extending their home range 

or left their home range for a time. Does, on the 

other hand, took a circuitous course that began and 

ended in their home range. Before the roundups the 

mean home range for does was 173 ha (427 acres), 

and after the disturbance the mean size of the home 

range had not change significantly, 190 ha (469 

acres). The mean home range for bucks enlarged 

significantly from before the roundups, 285 to 402 

ha (704 to 993 acres) (Hood and Inglis 1974). These 

before-disturbance home ranges are in close agreement 

to those determined by Rogers and others (1978) of 

a mean of 2.9 squire miles (751 acres) in semidesert 

grass-shrub community in Arizona.

Mule Deer

Population Trends and Changes with 

Habitat—Mule deer are found throughout Western 

North America (fig. 2-5). There is disagreement on how 

common they were prior to Euro-American settlement. 

Diaries and journals from the period 1820 to 1834 of 

early Euro-American explorers and mountain men 

indicate that mule deer were seen only incidentally. 

Also, researchers estimate that no more than five 

million, and possibly fewer, mule and black-tailed 

deer occupied the western United States during preColumbian times (Teer 1996).

Human settlement, unrestricted hunting, drought, 

severe blizzards, and the conversion of habitat to 

agriculture caused declines in mule deer populations. By the early 1900s, mule deer were generally 

scarce throughout much of the West. Widespread 

and intensive livestock grazing, logging, and burning 

proved beneficial to mule deer. Plants that were more 

palatable to mule deer than those that originally 

dominated native vegetation either invaded or increased in abundance (Teer 1996). This improvement 

in habitat along with strict hunting regulations, 

control of predators, and perhaps favorable weather 

conditions allowed populations to increase to all-time 

highs during the 1920s through the 1950s (Peek and 

Krausman 1996).

Since those high population levels, a variety of 

factors including overpopulation, the aging of shrub 

habitats to less productive seral stages, liberalized 

hunting regulations, and severe winters and drought 

led to declines in mule deer populations. By the mid1960s and early 1970s, mule deer populations in many 

areas of the West had declined sharply (Peek and 

Krausman 1996).

The vegetation communities that supported mule 

deer have been subjected to livestock grazing, fire 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 23

suppression, uncharacteristically hot wildfires, and 

logging. These factors combined with natural successional processes have changed the natural habitats of 

mule deer. In some areas excessive grazing reduced 

grasses and brought about a reduction of fine fuels 

necessary to carry wildfires (Peek and Krausman 1996). 

This created a favorable environment for trees and 

shrubs, allowing many browse species to become larger 

and more vigorous and establish in higher densities. 

For example, important species such as bitterbrush 

(Purshia tridentata), sagebrush, curlleaf mountainmahogany (Cercocarpus ledfolius), and rabbitbrush 

increased. Eventually, however, lack of fire or some 

other stand renewal process led to decadent/senescent 

older aged shrub stands that are less productive for 

mule deer. Dense, older aged shrub stands can restrict 

animal movements and browse availability (Peek and 

Krausman 1996).

Invasion by exotic annual grasses such as cheatgrass 

(Bromus tectorum) has caused range deterioration in 

some areas. Cheatgrass can out-compete native species and tends to burn frequently, thus eliminating 

the opportunity for woody browse species to become 

established. Ranchers and range managers encourage 

the production of grasses as forage for livestock and 

not woody browse for mule deer. Fire, mechanical 

manipulation, and herbicides are all employed to 

control and eradicate woody species. Because only a 

part of mule deer diet is composed of woody plants 

it is difficult to establish how these treatments have 

impacted mule deer numbers (Peek and Krausman 

1996).

Predation—Predators include coyotes (Canis 

latrans), bobcats (Felis rufus), golden eagles (Aquila 

chrysaetos), wolf (Canis spp.), mountain lion, grizzly 

bear (Ursus arctos), and black bear (Ursus americanus) (Peek and Krausman 1996, Shawn and Dood 

1984). The effect of predators, alone or in combination, on the dynamics of ungulate populations will 

vary with relative size of the predator and ungulate 

populations, ungulate dispersion patterns, sex and 

age, and general health of the populations. The effects of predation are difficult to isolate from all the 

factors concurrently acting on populations.

Deer population size will often influence the role 

predation plays. Predators, particularly those whose 

numbers are only slightly influenced by prey density, 

generally will have a relatively greater effect on deer 

populations whose numbers have been reduced by 

other factors (Wisdom and Thomas 1996).

Grazing and Mule Deer—Livestock grazing 

may interact with mule deer by altering plant succession to favor or reduce deer forage and cover. Changes 

in plant productivity caused by livestock grazing may 

also affect mule deer. Kie (1996) found that when cattle 

stocking rates increased during the fall and early winter 

when herbaceous forage was limited, then mule deer 

maximized energy by spending less time feeding. And 

when herbaceous plants began to grow rapidly there 

appeared to be no competition between deer and cattle, 

and increased cattle stocking rates had no effect on 

the time deer spent foraging.

Mule deer can also be directly disturbed by livestock 

and associated human activity. Mule deer responses 

may include changes in distribution patterns and 

habitat use, modification of activity, or alteration in 

population density (Peek and Krausman 1996).

Forage Choice by Mule Deer and Cattle—Peek 

and Krausman (1996) describe winter mule deer diet 

as consisting of an average of more than 74 percent 

shrubs and trees, 15 percent forbs, and 11 percent 

grasses and sedges and rushes (table 2-2). Spring diets 

include shrubs and trees (49 percent), grasses, sedges, 

and rushes (26 percent), and forbs (25 percent) (table 

2-2). Summer diets include shrubs (49 percent) and 

forbs (46 percent) and minor proportions of grasses, 

sedges, and rushes. The fall diet showed an increase 

in dietary use of shrubs and trees to an average of 60 

percent, while forbs declined to an average of 30 percent, and grasses, sedges, and rushes were 9 percent 

(table 2-2).

Figure 2-5. Mule deer distribution in North America (NatureServe 2004).
24 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

The general diet for mule deer shows high variability across the range depending on what is available. 

Habitat manipulations can stimulate changes in plant 

use. For example, burning or grazing that removes 

standing litter and stimulates early growth of bluebunch wheatgrass (Agropyron spicaturn) can increase 

wheatgrass palatability, encouraging use by mule deer 

(Peek and Krausman 1996).

Forage used by livestock changes with intensity 

of use of a range, between areas, seasons, and years, 

as with mule deer. Grasses generally predominate in 

the summer diet of cattle, but forbs may constitute 

a significant proportion of cattle diet in some areas 

(Peek and Krausman 1996). Forbs are used more in 

spring diet of cattle and when grazing is light. Highest 

forb use occurred in late spring on eastern Oregon 

grasslands (Wisdom and Thomas 1996). Foraging use 

of shrubs and trees by cattle is light during all seasons, 

but woody plants may be a significant part of the diet 

during some years and in some areas.

The contrast between cattle and mule deer forage 

preferences is best exemplified by their usages of 

grasses. Cattle concentrate on the most abundant 

grasses that are often dominants in the ecosystem 

while mule deer tend to use subdominant species 

such as the bluegrasses, fine-leafed species such as 

fescues, or tips of leaves of species such as crested 

wheatgrass.

Potential Overlaps in Forage Choice—Drought 

and intensity of grazing are significant factors that 

influence the potential for competition for forage between mule deer and livestock (Julander 1955). When 

the highly palatable forb yellow sweetclover (Melilotus 

officinalis) was reduced in abundance because of low 

precipitation, greater use of shrubs and grasses by deer 

and cattle resulted (Bowyer 1986). If grazing is heavy, 

then cattle and mule deer may be forced to compete 

for scarce forage. Diet choice may be expected to vary 

between years for both species, depending upon what is 

available and what is palatable. Managing for several 

plant species may provide the level of nutrients needed. 

In maintaining mule deer habitat, it is important to 

provide a variety of forage choices and species diversity in plant communities. Heavy grazing or drought 

that reduces plant diversity or plant productivity can 

increase the potential for forage competition among 

the different grassland wildlife species (Peek and 

Krausman 1996). Forage availability in early summer 

is much higher than later in the season when forages 

become senescent. While range managers cannot 

anticipate the weather, the implications are to keep 

grasslands productive with a diversity of forage so 

animals have opportunities to exercise free choice and 

minimize potential competition.

In winter, forage use and species abundance are 

influenced by snow depth (Peek and Krausman 1996). 

Table 2-2. Comparison of forage use by season between mule deer and cattle.

Season Mule deer Cattle

Winter

Shrubs and trees Averaged > 74% of (range = 27-100%) Shrubs and trees typically receive light use.

Forbs Averaged 15% (range = 0-43%)

Grasses and grasslikes Average 11% (range = 0-53%)

(sedges and rushes)

Spring

Shrubs and trees Average 49% (range = 6-92%) Shrubs and trees typically receive light use.

Forbs Average 25% (range = 0-43%) Forbs appear to be used more in spring and when 

 grazing is light.

Grasses and grasslikes Average 26% (range = 4-64%)

(sedges and rushes)

Summer

Shrubs Average 49% (range = 12-95%) Shrubs and trees typically receive light use.

Forbs Average 46% (range = 3-77%)

Grasses and grasslikes Very low Grasses generally predominate in the summer diet 

(sedges and rushes) of range cattle, but forbs may constitute a significant 

 proportion in some areas.

Fall

Shrubs and trees Average 60% (range = 3-97%) Shrubs and trees typically receive light use.

Forbs Average 30% (range =2-78%)

Grasses and grasslikes Average 9% (range = 0-24%)

(sedges and rushes)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 25

Although mule deer may forage on a combination of 

forbs, grasses, and browse plants during snow-free 

periods, snow conditions may limit the use of forbs 

and grasses covered with snow that might otherwise 

be selected. Releasing livestock on winter ranges 

prior to snowfall should help to minimize reductions 

in available forage for mule deer.

Plant Succession—Sagebrushes may compose a 

moderate to heavy part of the winter and spring diet 

of mule deer (Peek and Krausman 1996) when forbs 

are unavailable and grasses are cured. Mule deer 

using less nutritious woody browse forage will switch 

to green grasses and forbs when they appear in spring 

without shifting to new habitats.

Mule deer graze many exotics plants, including 

cheatgrass, tansy mustard, filaree (Erodium cicutariurn), and tumble mustard. Wyethia, dandelion, 

salsify (Tragopogon dubius), and the eriogonums 

(Eriogonurn spp.), which would be expected to increase 

with cattle grazing, are also foraged by mule deer. 

However, balsom-root (Balsamorhiza spp.), tall asters 

(Aster spp.), bluebells (Mertensia spp.), and geraniums 

(Geranium spp.), which would be expected to decrease 

with livestock grazing, may be locally important mule 

deer foods (Peek and Krausman 1996).

The semidesert grasslands of the Southwest have 

also been subject to extensive modifications due to 

grazing and reduction of fire (Peek and Krausman 

1996). There has been a general trend from grassdominated communities in semidesert grasslands to 

shrub-dominated communities, but this trend has occurred in both grazed and ungrazed areas. An increase 

in shrubs is attributed to a hotter and drier climatic 

trend (Bock and Bock 1996). These grasslands are 

extremely important mule deer habitats and probably 

have become more so as shrubs have proliferated.

Rangeland conditions have generally improved since 

the drought of the 1930s (Peek and Krausman 1996). 

However, “improvement” is often interpreted in relation 

to livestock forage and might not reflect improvement 

for mule deer. Increases in livestock forage may mean 

that ground cover has improved, even if numbers of 

palatable mule deer forage species have declined. 

Mule deer habitat may benefit in some circumstances 

where plant species diversity is enhanced (Peek and 

Krausman 1996).

Grazing Systems and Their Influence on Mule 

Deer—This section includes some recommendations 

that will benefit mule deer (Peek and Krausman 

1996):

• Construct fences that will permit passage of deer. 

These fences can be constructed to minimize 

interference with movement between pastures.

• Water developments may be used by deer, and 

guidelines exist to accommodate wildlife (Boroski 

and Mossman 1966, Peek and Krausman 1996, 

Rickel and others 1996). However, water development may allow cattle or other native ungulate 

species to graze areas previously not grazed and 

may displace deer.

• Forage use by livestock must be kept at levels 

that will ensure enough food for mule deer or 

will enhance growth of forages palatable to deer 

following grazing. This assumes that the objective 

of the grazing system is merely to integrate grazing with retention of mule deer habitat. However, 

if objectives include improvement of habitat by 

increasing shrub production, then special grazing 

management is indicated.

Peek and Krausman (1996) suggests that dual use 

of range by cattle and mule deer, when effectively 

managed, can be efficient land use even when both 

herbivores are utilizing the same forage species.

Use of Livestock to Enhance Mule Deer 

Habitat—Complete removal of livestock may not 

maintain nor increase needed shrub production on 

mule deer winter ranges, and efforts to enhance shrub 

retention and growth require active manipulation 

of plant cover (Peek and Krausman 1996). Fire and 

grazing management are logical tools for manipulating cover. They are relatively economical to apply, and 

managers have experience in their use.

Livestock grazing can be used to improve vegetation conditions for mule deer. Early observations that 

browsing stimulated production of additional twigs 

showed the potential for retention or improvement 

by grazing, recognizing that cattle might be used to 

alter shrub form and productivity to promote use by 

mule deer. The following objectives may be achieved 

by gazing livestock on mule deer habitat (Peek and 

Krausman 1996):

Mule deer photo by Lane Eskew.
26 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

• Removing old growth to stimulate new growth 

palatable to deer.

• Reducing shrub heights to levels that are available to deer.

• Removing old growth that blocks access to palatable forage for deer.

• Using forage less palatable to deer, and consequently reducing competition with forage used 

by deer.

Timing of grazing is important where mule deer and 

livestock may have similar diet preferences. Timing of 

livestock grazing can be adjusted to influence forage 

production to benefit deer. Spring grazing may be appropriate to support shrubs that compose the major 

part of mule deer’s winter diet. On ranges where the 

grass-forb components are important for deer, fall 

grazing is more appropriate. It has been found that 

sheep grazing from early summer to late summer 

improves forage quality in fall and increases forage 

quantity in spring for deer in Oregon’s Coast Range. 

Springtime sheep grazing of a Utah winter range 

increased bitterbrush current annual growth and 

reduced the standing dead grass cover, which allowed 

mule deer to select a more nutritious diet that included 

more herbaceous material (Fulgham and others 1982). 

Longhurst and others (1983) explains that grazing 

maintained oak-woodland vegetation in a productive 

seral stage for deer by promoting growth and production of important deer forages. Cattle grazing could 

be used to drive plant composition in ways similar to 

sheep management (Peek and Krausman 1996). These 

results demonstrate how livestock can manipulate 

forage composition and productivity to achieve mule 

deer habitat objectives, suggesting that complete 

exclusion of livestock may not be necessary.

Long-term changes in vegetation composition may 

occur when special grazing treatments are used, and 

these changes should be monitored to ensure that 

trends in rangeland conditions are detected and 

modifications of grazing, if necessary, are timely 

(Peek and Krausman 1996). Manipulation of timing 

and species of livestock to enhance mule deer range 

has been effectively demonstrated and should be 

more broadly applied in future habitat management 

programs for mule deer and other wildlife. While 

such range management programs may require 

more flexibility and coordination, benefits derived 

may more than offset the added efforts. We can now 

purposefully manage grasslands to benefit livestock, 

mule deer, and rangeland vegetation by establishing 

suitable objectives and developing grazing programs 

that are flexible and tailored to the specific situation 

(Peek and Krausman 1996).

Deer Management

Herd and Population Management—Maintenance, growth, and reproductive states require different 

nutritional intakes. Three seasons of the year are 

particularly critical for mule deer: fall, winter, and 

spring. Fall is important for putting on fat and for fetus 

development during the winter. A good winter diet is 

important for survival during the cold months. Then 

in spring, the plants that are early spring growers are 

critical for fetus development and lactation. Plus, aside 

from forage requirements, fawns need tall vegetation 

for hiding from predators. The importance of diets 

for fall, winter, and spring is similar for all Cervidea. 

Healthy growth of fawns after weaning is dependent 

on protein in their diets. Males have higher protein 

requirements than females. During latter stages of 

pregnancy, protein requirements are intermediate between that of growth and maintenance (Teer 1996).

Although deer can subsist for long periods on water 

obtained from vegetation, free water is an important 

component of deer habitat. Rickel and others (1996), 

in modeling mule deer habitat, concluded that the 

maximum traveling distance to water for fawns was 

1 mile, for does 2 miles, and for bucks 3 miles. Water 

is usually supplied by natural water areas or can be 

provided through stock ponds and troughs used to 

water livestock.

Starvation results from the exhaustion of food 

resources and is a consequence of poor deer herd 

management or of no management. If the habitat will 

not support the herd and if the annual crop of deer is 

not removed by other means, starvation will remove 

it. This is an axiom in deer management. Natural 

regulation of herbivore populations is difficult when 

natural systems have been disrupted by man.

Predator Control—Predators are often cited as 

the chief cause of deer declines by the general public. 

Coyotes, mountain lions, bobcats, and other carnivores 

do kill deer, and their impacts on deer numbers can be 

important. However, quality of habitat and hunting 

may be much more important than predation.

Whether predators can control numbers of deer 

has been the subject of long debate. Now, however, 

considerable evidence has accumulated to show that 

coyotes, mountain lions, and other predators can 

substantially impact, if not control, herd numbers in 

certain situations (Teer 1996).

Coyotes are omnivorous and opportunistic in relation to diet. They take vegetable materials (that is, 

fruits, mast) when available and animal prey when 

abundant and easily caught (Teer 1996). Deer fawns 

are a large proportion of coyote diets especially when 

cover is sparse and other foods are scarce.

Control of predators through some kind of bounty 

system or other subsidized control program is often 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 27

the remedy called for by misinformed citizenry. Bounty 

systems by government agencies have often been 

abused by those engaged in predator management. 

Cost and benefit evaluations of the bounty system 

indicate it is ineffective in preventing predation or 

encouraging deer numbers (Teer 1996).

Predation along with sport hunting can be used to 

control deer populations. Conversely, when deer numbers are to be increased or protected, predator control 

may be an important activity (Teer 1996). A blanket 

recommendation cannot be made about predator control; each situation must be addressed separately.

Pronghorn________________________

The pronghorn is one of the key herbivores on 

both the plains and desert grasslands (fig. 2-6). The 

pronghorn’s scientific name, Antilocapra americana, 

means “American antelope goat.” However, the deerlike pronghorn is neither an antelope nor a goat. They 

evolved in North America and are the sole surviving 

members of an ancient family dating back 20 million 

years (The Pronghorn 2004, Yoakum and others 1996). 

Pronghorn existed with bison in legendary numbers 

when Lewis and Clark made their historic journey 

across the continent (fig. 2-6).

Today, an estimated 98 percent of pronghorn share 

their habitat with domestic livestock (Yoakum and 

others 1996). An estimated 60 percent of all pronghorn 

live on private lands and the remainder on Federal or 

State government-administrated lands.

Pronghorn Distribution

When Euro-Americans began to explore North 

America, they found pronghorn from the plains 

of south-central Canada (Alta., Saskatchewan, 

Manititoba), south through most of the Western 

United States and into Mexico (fig. 2-6). Herds 

ranged from the Mississippi River to the Pacific 

Ocean in central California. Herds extended from 

the Gulf of Mexico in Texas to the Pacific Ocean in 

lower California.

Pronghorn populations reached greatest densities 

with bison on grasslands of the Great Plains. Smaller 

populations occupied Intermountain and desert 

regions. They did not occur east of the Mississippi 

River (Yoakum and others 1996). Pronghorn now 

occupy many of their historic grasslands but in 

greatly reduced numbers. Contemporary herds 

are confined to smaller isolated habitats (fig. 2-6), 

unable to make historical seasonal movements 

because of freeways, railroads, fencing, and other 

anthropogenic constraints. The pronghorn is the 

most representative big game species dependent 

upon Western grasslands (Yoakum and others 

1996).

Pronghorn Abundance

Yoakum and others (1996) estimated there were 30 

to 60 million pronghorn at the beginning of the 19th

century. Some reports indicate they were as numerous 

or possibly more abundant than bison (Shaw 1996, 

Yoakum and others 1996). During the late 1800s, sport 

and commercial hunters hunted pronghorn herds and 

killed animals regardless of sex or age. Much of the 

best habitat was lost to the plow, and pronghorn movements became increasingly restricted by fences and 

other human-made impediments. Pronghorn also were 

subjected to livestock diseases and parasites to which 

they had little resistance. Numbers dropped from an 

estimated 35 million in 1800, to perhaps 13,000 in 1910 

(Yoakum and others 1996). Public concern arose, protective laws were enacted, and supporting conservation 

and management practices were implemented. Within 

a decade, populations more than doubled and have 

continued to recover. Currently, pronghorn antelope 

number approximately 1 million (Pronghorn 2004). 

They are now second only to deer in large herbivore 

abundance and harvest in the United States.

Two subspecies—Sonoran pronghorn (A. a. 

sonoriensis) and Chihuahuan (Mexican) pronghorn 

(A. a. mexicana)—are found in the southern part of 

Arizona and New Mexico. The Sonoran pronghorn is 

Federally endangered. Of the total pronghorn population in North America, less than 1 percent is classified as 

Figure 2-6. Pronghorn antelope distribution in North America 

(NatureServe 2004).
28 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

endangered (Yoakum and others 1996). The endangered 

populations are on the grasslands along the MexicoUnited States border (BISON-M 2005, Ockenfels and 

others 1996, Yoakum and others 1996).

Chihuahuan (Mexican) Pronghorn—Historically 

this subspecies ranged throughout southeastern and 

south-central Arizona and inhabited grass-shrub valleys and grasslands. In Arizona by 1900, Chihuahuan 

pronghorn populations had been reduced, and they 

were not reported in Arizona after the 1920s (BISOM-M 

2005). Chihuahuan pronghorn populations in Arizona 

were probably extirpated by uncontrolled subsistence 

hunting and changing land-use patterns (BISON-M 

2005). Populations were reintroduced from Texas stock 

in several areas within historic range from 1981 through 

1985 and 1987, including Empire Ranch near Sonoyta, 

San Bernardino Valley, Fort Huachuca, and Buenos 

Aires National Wildlife Refuge (most western extent 

of its range) (Arizona’s Natural Heritage Program:

Heritage Data Management System 2005).

The Chihuahuan pronghorn in Arizona are classified as big game, permitting it to be hunted. Only 

bucks are harvested annually (Arizona’s Natural 

Heritage Program:Heritage Data Management System 

2005).

Sonoran Pronghorn—Sonoran pronghorn has 

been described as the smallest of the five subspecies. 

Some females lack horns (Arizona’s Natural Heritage 

Program: Heritage Data Management System 

2005).

The historic range is difficult to determine because 

the subspecies was not described until 1945, many 

years after the population had declined and marginal 

populations were extirpated (BISON-M 2005). Herds 

that were observed along the lower Gila River, Arizona, 

by early travelers are presumed to have been Sonoran 

pronghorn. They are thought to have ranged from 

Hermosillo to Kino Bay, Mexico, to the south; Highway 

15, Mexico to the east; Altar Valley and the Tohono 

O’odham Indian Reservation to the North; and Imperial 

Valley, California to the west (BISON-M 2005).

Presently, in Arizona, they are found on the Cabeza 

Prieta National Wildlife Refuge, the Organ Pipe Cactus 

National Monument, the Luke Air Force Barry M. 

Goldwater Gunnery Range, and possibly the Tohono 

O’odham Indian Reservation. In Mexico, they are 

believed to be confined to the northwest part of Sonora 

(Arizona’s Natural Heritage Program: Heritage Data 

Management System 2005).

Based on a study conducted in 1984 using collared 

Sonoran pronghorn, four males had home ranges ranging from 64.5 km2 to 1,213.6 km2 (24.9 miles2

 to 468.6 

miles2 ) an average of 799.7 km2 (179.8 miles2), while 

six females had home ranges ranging from 40.7 km2 

to 1,143.7 km2 (150.7 miles2 to 441.6 miles2

) and an 

average of 465.7 km2 (176.3 miles2

). The large variation 

in home range size for this study appears to be tied 

to forage and possibly water availability (BISON-M 

2005). These animals must use large tracts of land to 

obtain adequate forage.

Sonoran pronghorn exhibit the same social doe/fawn, 

territorial, and flight behaviors as noted for other pronghorn. A heightened response to human traffic has been 

noted. Once aware of an observer, Sonoran pronghorn 

are quick to leave the area (BISON-M 2005).

Fawning for Sonoran pronghorn takes place from 

February to May, and as early as January for populations in Mexico. Although the stress of summer rutting 

on pronghorn is higher, spring drop is desirable to 

coincide with temperate weather and spring forage. 

Sonoran pronghorn fawns are nursed for 60 days, unlike northern populations, which nurse up to 90 days 

(Arizona’s Natural Heritage Program: Heritage Data 

Management System 2005).

Sonoran pronghorn were observed browsing on forbs, 

shrubs, and cacti. Forbs and cholla (Cactaceae ssp) 

were browsed on during the summer and fall seasons, 

while shrubs, cholla, and ocotillo were browsed on the 

remainder of the year (Arizona’s Natural Heritage 

Program: Heritage Data Management System 2005). 

Cholla appears to be a key succulent forage item in 

their diet during the summer, constituting nearly 50 

percent of their diet, apparently to meet their water 

requirements (Arizona’s Natural Heritage Program:

Heritage Data Management System 2005).

Free-standing water is limited within range of 

the Sonoran pronghorn in Arizona. In 1984, collared 

Sonoran pronghorn were observed at water troughs in 

November, January, and August. Tracks were observed 

leading up to, then away from, seasonal potholes during 

the monsoon season. The collared pronghorn exhibit 

movements apparently tied to water, as well as forage, 

availability. During a study conducted in 1995 using 

collared animals, Sonoran pronghorn were observed 

using an ephemeral supply of water on a daily basis 

in a crater on the Barry M. Goldwater Gunnery Range 

A doe and buck pronghorn. (Courtesy of NatureServe© Larry 

Master)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 29

(Arizona’s Natural Heritage Program: Heritage Data 

Management System 2005).

In Arizona, the reason for population decline is 

attributed mainly to loss of habitat and drought. The 

drying of major rivers and overgrazing significantly 

altered Sonoran pronghorn habitat in southwestern 

Arizona by the 1930s. The population has not recovered 

since the establishment of three large public land 

withdrawals and the removal of cattle from these 

areas in the early 1980s (Arizona’s Natural Heritage

Program: Heritage Data Management System 2005). 

The only significant loss of habitat in recent years 

in Arizona occurred on the Tohono O’odham Indian 

Reservation where severe overgrazing by cattle, 

coupled with drought, resulted in the loss of large 

areas of pronghorn habitat (Arizona’s Natural Heritage

Program: Heritage Data Management System 2005). 

In Mexico, it is believed that economic exploitation 

of habitat (grazing and agriculture) and poaching are still causing population and habitat losses 

(Arizona’s Natural Heritage Program: Heritage Data

Management System 2005).

Measures have been taken to protect and enhance 

this endangered pronghorn. The establishment of 

three large public land withdrawals in Arizona, Cabeza 

Prieta National Wildlife Refuge, Organ Pipe Cactus 

National Monument, and Luke Air Force Gunnery 

Range (Barry M. Goldwater Gunnery Range). The 

removal of hunting from these sites, and the restriction of vehicle traffic further protects the Sonoran 

pronghorn. The removal of fencing between the Cabeza 

Prieta NWR and Organ Pipe Cactus NM, and within 

the Cabeza Prieta NWR in the 1990s has allowed 

for easier natural movement of Sonoran pronghorn 

(Arizona’s Natural Heritage Program: Heritage Data 

Management System 2005).

Predators

Predators of all pronghorn include coyotes, bobcats, 

mountain lions, golden eagles, and wild dogs. Coyotes 

are the primary predator of fawns in Arizona and New 

Mexico. Losses of pronghorn due to predation vary with 

pronghorn and predator numbers, habitat type, and 

availability of alternative food sources for predators. 

A pronghorn’s speed and exceptional eyesight are its 

main defense from predators; pronghorn can detect 

movement up to 4 miles away. A fawn’s best defense 

from predators is to choose good cover to lay down and 

to lie motionless (Arizona’s Natural Heritage Program:

Heritage Data Management System 2005, BISON-M, 

Yoakum and others 1996).

Pronghorn Habitat Requirements

Currently, 68 percent of the herds inhabit grasslands, 31 percent shrub-steppes, and 1 percent deserts 

(Arizona’s Natural Heritage Program: Heritage Data 

Management System 2005, BISON-M, Yoakum and 

others 1996)). Most populations historically occurred 

in grasslands and shrub-steppes with large expanses 

of flat or low rolling terrain without major physical 

barriers to seasonal movements.

Pronghorn forage studies show that pronghorn 

prefer forbs, then shrubs, and that grasses were the 

least preferred forage. Thus, pronghorn management 

should promote an abundance of preferred forb species. 

However, an abundance of less preferred forbs and 

reduced grass cover resulting from excessive livestock 

grazing is detrimental to pronghorn.

Pronghorn select palatable, nutritious, succulent 

forage, and being selective, take small bites of preferred 

leaves, flowers, and terminal parts. Rarely do they 

feed extensively in one place but move frequently as 

they forage. Pronghorn consume less then 1 percent of 

forage produced on Western grasslands in the United 

States (Yoakum and others 1996).

The quality and quantity of vegetation appear to 

be major factors affecting pronghorn densities and 

production. Habitat quality is directly related to proper 

percentages, quantities, and distribution of physiographic and vegetative characteristics; too little or too 

much of any component may limit pronghorn production and survival. Pronghorn population dynamics 

between the shrub-steppes of the Great Basin and the 

grasslands of the Great Plains were compared, showing 

that fecundity was 190 fawns per 100 producing does for 

both ecosystems (Yoakum and others 1996). But fawn 

survival was twice as high, and grass and forb production 

was higher on the Great Plains than in the Great Basin. 

Nutritive values (particularly protein) of grasses and 

forbs were greater than shrubs during late spring and 

early summer when fawning occurred. Fawn survival 

was higher on the Great Plains because of abundant, 

nutritious grasses and forbs during late gestation and 

early lactation (Yoakum and others 1996). The lesser 

amount of grasses and forbs in the Great Basin, partly 

because of livestock grazing, apparently resulted in 

grasslands of lower carrying capacity for pronghorn.

Pronghorn Relationships with Other 

Grassland Wildlife

Prairie Dogs—Because forb production is high in 

the centers of prairie dog colonies, it is speculated that 

prairie dogs (Cynomys spp.) may enhance grasslands 

for pronghorn by consuming grasses and disturbing 

soils, thereby increasing the abundance and variety 

of forbs (Yoakum and others 1996).

Predators—Predatory birds and mammals kill 

some pronghorn, but predation usually is significant 

only on marginal grasslands or sites where predator 

numbers are high in relation to pronghorn numbers. 

Most predator losses occur among fawns 1 to 3 weeks 
30 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

of age, while separated from their dams (Yoakum and 

others 1996).

Bison—Bison and pronghorn shared the same 

range, foraging and watering on the vast grassland 

prairies (Yoakum and others 1996). Apparently bison 

did not coexist with pronghorn historically in the Great 

Basin or on deserts. In comparing diets of bison and 

pronghorn in Colorado, diet differences were attributed 

to the species rather than to the grasslands (Yoakum 

and others 1996). Diets of bison and pronghorn studied 

in Montana overlapped minimally: bison consumed 

mostly grasses while pronghorn ate forbs and shrubs 

(table 2-3) (Shaw 1996). Bison disturbed dominant 

grass communities by trampling, wallowing, and 

grazing that resulted in greater production of forbs 

and shrubs favored by pronghorn.

Elk—There were larger range and habitat overlaps 

between elk and pronghorn historically than now. Elk 

and pronghorn have been observed foraging together 

with no acts of aggression and little dietary overlap 

(table 2-3) (Yoakum and others 1996). There is little 

overlap in habitat use between elk and pronghorn in 

Montana and little competition for space, water, or 

forage (Yoakum and others 1996). Elk forage primarily 

on grasses, whereas pronghorn prefer forbs and shrubs; 

consequently, dietary overlap was low.

Deer—Pronghorn occur on grasslands with whitetail 

and mule deer; however, their distributions usually do 

not overlap (Yoakum and others 1996.). All have similar 

diets (primarily forbs and shrubs) but overlaps are 

minimal (table 2-3) because of differences in habitat 

occupancy patterns. Behavioral interactions are also 

minimal because pronghorn are generally diurnal 

while deer are crepuscular and nocturnal.

Pronghorn Relationships with Livestock

Pronghorn and Livestock Forage Competition—An estimated 98 percent of pronghorn populations 

share grasslands with domestic or feral livestock during 

some of the year (Yoakum and others 1996). Livestock 

on Western grasslands are either domestic animals such 

as cattle, sheep, and horses, or feral stock, primarily 

horses. Table 2-4 is a summary of the dietary overlaps 

between pronghorn and cattle, horses, and domestic 

sheep on grasslands and shrub-steppes.

Livestock grazing may alter pronghorn habitats 

more than any other human-controlled activity by 

changing vegetation structure and composition (Shaw 

1996). When pronghorn and livestock grazed grasslands in southeastern New Mexico, they used many 

of the same forage species. Pronghorn did not switch 

to less preferred forage classes and were adversely 

affected when forbs and shrubs were depleted through 

drought or heavy livestock grazing (Shaw 1996).

Pronghorn use of plants noxious to livestock has 

been well documented (Shaw 1996), and for this reason 

pronghorn may be considered beneficial to livestock 

on grasslands. Spines of bull thistle (Cirsium spp.), 

Russian thistle (Salsola spp.), and cacti are less 

palatable to livestock, but pronghorn frequently eat 

them. Other noxious plants that livestock do not eat 

but are consumed by pronghorn include: in Oregon—

woolly locoweed (Astragalus mollissimus), larkspur 

(Dephinium spp.), lupine (Lupinus spp.), death camas 

(Zigadimus spp.) (Shaw 1996); in Texas—locoweed 

(Astragalus spp.), woody senecio (Senecio spp.), Riddell 

groundsel (Senecio spp.), (Yoakum and others 1996); 

in Colorado—cocklebur (Xanthium spp.), snakeweed 

(Gutierrezia spp.), bull thistle, Russian thistle, cacti, 

Table 2-3. Dietary overlap for forage classes between pronghorn and bison, white-tailed deer, mule deer, and elk (Yoakum and 

others 1996).

% dietary overlap

Species Location Biome Grasses Forbs Shrubs Annual

Bison Wichita Mountain National Wildlife Refuge, OK Grassland 1 1 2

Pawnee Grasslands, CO Grassland 41 3 0 44

National Bison Range, MT Grassland 2.8 1.4 0.1 4.3

Elk Wichita Mountain National Wildlife Refuge ,OK Grassland 0.1 24 0 24.1

National Bison Range, MT Grassland 2.9 4.6 3.8 11.3

Trickel Mountain, CO Shrub-steppe 3.5 11.5 45.5 60.5

Whitetailed deer Wichita Mountains National Wildlife Refuge, OK Grassland 0 99 0 99

National Bison Range, MT Grassland 2.9 24.8 18.5 46.2

Mule deer Yellowstone National Park, WY Shrub-steppe 16.6 9.2 52 77.8

National Bison Range, MT Grassland 2.3 27.1 33.3 62.7

Northeast California, northwest Nevada Shrub-steppe 2.3 4.2 82.3 88.8

Trickle Mountain, CO Shrub-steppe 3.5 10.2 79.5 93.2

Sheldon National Wildlife Refuge, NV Shrub-steppe 5 31 39 75
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 31

chokecherry (Prunus spp.) (Hoover and others 1959); in 

Texas—tarbush (Flourensia cernua) (Hailey 1979); and 

in Utah—halogeton (Halogeton spp.) (Shaw 1996).

A potential exists for forage competition between 

pronghorn and domestic sheep. Yoakum and others 

(1996) reported intense competition for preferred 

forbs. A study of pronghorn and domestic sheep forage competition conducted during the early 1980s in 

New Mexico determined that livestock used as much 

as 40 percent of available forbs (Shaw 1996). As forbs 

declined in number or decreased in moisture content, 

livestock increased their use of grasses, increasing 

the potential competition for forbs. Overlap is highest 

when forbs are most available and lowest when they 

are least available. Yoakum and others (1996) reported 

that moderate use of shrubs on the cold desert of 

southwestern Utah by domestic sheep during winter 

created unfavorable conditions for pronghorn until the 

spring regrowth occurred.

Horses predominantly feed on grasses while pronghorn prefer forbs and shrubs. Five studies listed dietary 

overlap as low to moderate (11 to 36 percent) (table 2-4), 

suggesting that competition for food was limited.

Pronghorn and Livestock Behavioral 

Relationships—The behavioral relationship between 

pronghorn and livestock is nonaggressive while both 

species are feeding, drinking, or resting. Pronghorn does 

avoided cattle during fawning, resulting in selection of 

less favorable fawn-production sites (Shaw 1996).

Domestic and feral horses occupy grasslands with 

pronghorn (Yoakum and others 1996). Both animals 

have been observed to water together, with pronghorn 

giving ground only when directly approached by horses 

(Shaw 1996). Pronghorn and feral horse relationships 

in the Great Basin have shown various instances where 

pronghorns were displaced by horses.

Pronghorn and Livestock Diseases—Exposure 

to diseases is a concern when pronghorn and cattle 

share grasslands because their spatial distributions 

usually overlap. There is concern that either species may 

serve as a reservoir for diseases that affect the health 

of the other (Yoakum and others 1996). Blue tongue is 

probably the most serious disease of pronghorn, and 

cattle are a primary agent for spreading this disease 

to pronghorn. Blue tongue is fatal to domestic sheep. 

Neither pronghorn nor sheep are important reservoirs 

Table 2-4. Forage class dietary overlap among pronghorn and cattle, horses, and domestic sheep (Yoakum and others 2006).

Percent dietary overlap

Class of livestock Location Biome Grasses Forbs Shrubs Annual

Cattle Trans-Pecos, TX Grassland 4 7 9 20

Southwest, MT Grassland 3 20.3 1.6 24.9

Winnett, MT Shrub-steppe 3 13 0 16

Rawlins, WY Shrub-steppe 7.1 0.2 39.4 46.7

Pawnee, CO Grassland 46.3 2 6 54.3

Northeast California, northwest Nevada Shrub-steppe 4.1 3.4 3.7 11.2

Southwest, UT Shrub-steppe 0 0 27 27

Roswell, NM Grassland 4 15.2 1.3 20.5

Panhandle, TX Grassland 2.5 19 8.5 30

Trickle Mountain, CO Shrub-steppe 3.5 5 14 22.5

Little Lost-Birch Creek, ID Shrub-steppe 3 2 11 16

Burns Junction, OR Shrub-steppe 9.4 4.6 0.2 14.2

Sheldon National Wildlife Refuge, NV Shrub-steppe 5 9 1 15

Horse Sheldon National Wildlife Refuge, NV Shrub-steppe 3 23 2 28

Northeast California, northwest Nevada Shrub-steppe 2.5 5.2 5 12.7

Trickle Mountain, CO Shrub-steppe 3.5 2 31 36.5

Burns Junction, OR Shrub-steppe 13.5 2.1 0.3 15.9

Sheldon National Wildlife Refuge, NV Shrub-steppe 5 6 0 11

Sheep Trans-Pecos, TX Grassland 4 19 10 33

Red Desert, WY Shrub-steppe 3.2 2.6 28.4 34.2

Southeast, MT Grassland 3 27.3 25.3 55.6

Rawlins, WY Shrub-steppe 7.1 1.2 39.9 48.2

Pawnee, CO Grassland 46.3 18 3.3 67.6

Southwest, UT Shrub-steppe 0 0 46 46

Little Lost-Birch Creek, ID Shrub-steppe 3 14 39 56

Roswell, NM Grassland 4 50.2 6 60.2
32 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

for this disease because few animals survive to become 

carriers (Yoakum and others 1996). Leptospirosis causes 

some mortality in pronghorn, but evidence is insufficient 

to implicate either pronghorn or cattle as a primary 

reservoir of infection (Yoakum and others 1996).

Pronghorn had higher rates of parasitism on grasslands grazed by domestic sheep than grasslands grazed 

by cattle. In Wyoming, illness and deaths of pronghorn 

fawns have been attributed to parasitic infections that 

were prevalent on grasslands grazed heavily by sheep 

(Yoakum and others 1996).

Pronghorn Management

Grassland management practices and improvements can be beneficial or detrimental to pronghorn 

populations, depending upon how they are planned 

and implemented. If the biological requirements of 

pronghorn are not met, then any management action 

could be detrimental.

Effective management systems should consider 

the control of livestock and range improvements for 

livestock, determine how many livestock will use 

certain grasslands, and take into account seasons of 

livestock use, results of monitoring studies, physiology 

of plants, and the effects and needs of wildlife on the 

grasslands.

For grasslands used by pronghorn and livestock, 

the following guidelines are recommended for designing livestock grazing systems (Yoakum and others 

1996):

1. When allotting forage, the habitat requirements of 

pronghorn should be considered.

• Adequate amounts of preferred plant species 

should be reserved as forage for pronghorn. These 

include grasses, forbs, and shrubs determined 

from diet studies in the same or similar ecosystems. Consideration should be given to proper 

use of key forbs and shrubs. Make sure that 

these are not grazed beyond their physiological 

tolerance levels.

• Natural vegetation should be managed to provide 

an abundance and variety of forage classes. Most 

livestock graze grasses; pronghorns primarily 

consume forbs and shrubs. Grasslands producing 

mixtures of grasses, forbs, and shrubs will best 

serve livestock and wildlife. The challenge is to 

maintain existing grasslands in good ecological 

condition comprised of native vegetation. When 

deteriorated sites require rehabilitation, practices 

that restore vegetation to natural diversity are 

more desirable than practices that bring about 

monocultures and other unnatural conditions.

• Practices that increase availability of long-term 

drinking water sites are highly beneficial to 

both animal groups. Water should be available 

every 1.5 to 6.5 km. Water improvements can be 

designed in a number of ways; however, those 

that simulate natural waters are favored and 

cause fewer problems.

• Both pronghorn and livestock experience problems 

with predation, diseases, and parasites. Managers 

need to recognize how these factors affect the 

animals and coordinate control techniques beneficial to both groups.

2. Drinking water should be available during all seasons 

that pronghorn are in the area.

3. Fence construction should meet specifications to 

allow movements year-round for all pronghorn age 

classes.

4. When livestock grazing systems are designed using 

the “key plant species” concept, forbs and/or shrubs 

preferred by pronghorn should be included as key 

species.

5. Livestock use should be limited on pronghorn natal 

areas during the fawning season.

6. Livestock grazing systems that restrict, alter, limit, 

or deleteriously affect the habitat requirements 

of pronghorn should include mitigating measures 

and alternate procedures for enhancing pronghorn 

habitat.

Acknowledgments_________________

I thank Mark A. Rumble, Fred B. Samson, and 

John Sidle for reviewing this chapter and providing 

insightful comments. I express my appreciation to 

Deborah M. Finch for editing the chapter and to Nora 

C. Altamirano for helping prepare this chapter for 

publication. A thank you also goes to Kerry Mich for 

help on the range maps.

References________________________

Arizona’s Natural Heritage Program: Heritage Data Management 

System (HDMS). 2005. [Online] http://www.azgfd.gov/w_c/edits/

species_concern.shtml.

BISOM-M, New Mexico Game and Fish. Biota Information System 

of New Mexico. 2005.

 [Online] http://nmnhp.unm.edu/bisonm/bisonquery.php

Bock, C.E.; Bock, J.H. 1996. Factors controlling the structure and 

function of desert grasslands: a case study from southeastern 

Arizona. In: The future of arid grasslands: identifying issues, 

seeking solutions. B. Tellman, D.M. Finch, C. Edminster, and R. 

Hamre, eds. Proc. RMRS-P-3. Fort Coillins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Research Station.

Boroski, B.B.; Mossman, A.S. 1966. Distribution of mule deer in 

relation to water sources in northern California. Journal of 

Wildlife Management. 60(4)770-776.

Bowyer, R.T. 1986. Habitat selection by southern mule deer. 

California Fish and Game Bulletin: 72(3): 153-169.

Chung-MacCoubrey, A.L. 1996. Grassland bats and management in 

the Southwest. In: D.M. Finch, ed. 1996 Ecosystem disturbance 

and wildlife conservation in western grasslands – A symposium 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 33

proceedings, September 22-26, 1994; Albuquerque, NM. Gen. 

Tech. Rep. RM-GTR-285. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Forest and Range 

Experiment Station. 82 p.

Findley, J.S.; Harris, A.H.; Wilson, D.E.; Jones, C. 1975. Mammals 

of New Mexico. A.buquerque: University of New Mexico Press.

Fulgham, K.0.; Smith, M.A.; Malachek, J.C. 1982. A compatible 

grazing relationship can exist between domestic sheep and mule 

deer. In: I.M. Peek and P.D. Dalke, eds. Wildlife-livestock relationships symposium: proceedings 10. Moscow: University of Idaho, 

Forestry, Wildlife and Range Experiment Station: 458-478.

Hailey, T.L. 1979. A handbook for pronghorn antelope management 

in Texas. Houston: Texas Parks and Wildlife Department. 59 p.

Hood, R.E.; Inglis, J.M. 1974. Behavior responses of white-tailed deer 

to intensive ranching operations. Journal of Wildlife Management. 

38(3): 488-478.

Julander, 0. 1955. Deer and cattle range relations in Utah. Forest 

Science. 1: 130-139.

Keiter, R.B. 1997. Greater Yellowstone’s Bison: unraveling of an 

early American Wildlife Conservation achievement. Journal of 

Wildlife Management. 61: 1-11.

Kie, J.G. 1996. The effects of cattle grazing on optimal foraging in mule 

deer (Odocoileus hemionus). Forest Ecology and Management. 

88(1996): 131-138.

Knapp, A.K.; Blair, J.M.; Briggs, J.M.; Collins, S.L.; Hartnett, D.C.; 

Johnson, L.C.; Towne, E.G. 1999. The keystone roll of bison in North 

American tallgrass prairie. BioScience vol. 49 no 1: 39-50.

Meagher, M. 1986. Bison. Mammalian Species. 266: 1-8.

NatureServe. 2004. [online] http://www.natureserve.org/explorer

North American Elk Subspecies. 2004. Pennsylvania Elk Herd 

[online] http://www.pennsylvaniaelkherd.com/elk00076.htm

Ockenfels, R.A.; Ticer, C.L.; Alexander, A.; Wennerlund, J.A.; Hurley, 

P.A.; Bright, J.L. 1996. Statewide evaluation of pronghorn habitat 

in Arizona. W-78-R Final Rep. Phoenix: Arizona Game and Fish 

Department Fed Aid Wildife. Rastoration Project. 296 p.

Parmenter, R.R.; Van Devender, T.R. 1995. Diversity, spatial variability, and functional roles of vertebrates in desert grasslands. 

In: The Desert Grasslands. M.P. McClaran and T.R. Van Devender, 

eds. Tucson, AZ: The University of Arizona Press: 196-230.

Peek, M; Krausman, J.M. 1996. Grazing a mule deer. In P.R. 

Krausman, ed. Rangeland Wildlife. Denver, CO: The Society of 

Range Management: 183-192.

Perry, C.; Overly, R. 1977. Impact of roads on big game distributions 

in portions of the Blue Mountains in Washington. 1972-1973. Appl. 

Res. Bull. Olympia: Washington Department of Game. 39 p.

Pronghorn. 2004. Great Plains Nature Center. [online] http://www.

gpnc.org/pronghor.htm.

Rickel, B.W.; Anderson, B.; Pope, R. 1996. Using Fuzzy systems 

object-oriented programming, and GIS to evaluate wildlife habitat. 

AI Applications. 12(1-3):31-40.

Rocky Mountain Elk Foundation. 2004. [online] http://www.rmef.

org

Shaw, J.H. 1996. Bison. In: P.R. Krausman, ed. 1996. Rangeland 

Wildlife. Denver, CO: The Society of Range Management: 

227-235.

Shawn, J.R.; Dood, A.R. 1984. Summer movements, home range, 

habitat use, and behavior of mule deer fawns. Journal of Wildlife 

Management. 48(4): 1302-1310.

Southeastern Cooperative Wildlife Disease Study. 1982. White-tailed 

deer populations 1980. A map prepared in cooperation with the 

Emergency Programs, Veterinary Services, Animal and Plant 

Health Inspection Service, U.S. Department of Agriculture, 

through cooperative agreement 12-16-5-2230. Athens: University 

of Georgia. 1 p.

Teer, J.G. 1996. The white-tailed deer: natural history and management. In: P.R. Krausmann, ed. 1996. Rangeland Wildlife. Denver, 

CO: The Society of Range Management: 193-209.

The Pronghorn. 2004. Desert USA. [online] http://www.desertusa.

com/mag99/may/papr/pronghorn.html

Wisdom, M.J.; Thomas, J.W. 1996. Elk. In: P.R. Krausmann, ed. 

1996. Rangeland Wildlife. Denver, CO: The Society of Range 

Management: 157-181.

Yoakum, J.D.; O’Gara, W.; Howard, V.W. 1996. Pronghorn on the 

western rangelands. In: P.R. Krausmann, ed. 1996. Rangeland 

Wildlife. Denver, CO: The Society of Range Management: 

211-226.
34 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 35

Introduction_______________________

This chapter focuses on small mammals, reptiles, 

and amphibians that inhabit the grasslands within 

the Southwestern Region of the USDA Forest Service. 

The chapter is not intended to be an all inclusive list 

of species, but rather to address the species that play 

important roles in grassland ecosystems and that often 

are associated with the management of grasslands. 

Among the larger rodents discussed here are prairie 

dogs and pocket gophers. The small rodents include 

deer mice, voles, kangaroo rats, and pocket mice. 

Nonrodent species described in this chapter include 

the endangered black-footed ferret, as well as cottontail rabbits, jackrabbits, and bats. The herpetofauna 

include turtles and tortoises, lizards, snakes, frogs, 

toads, and salamanders.

The species discussed in this chapter serve important 

ecological roles and are considered important to the 

health and function of grassland ecosystems. Some species—for example, prairie dogs and kangaroo rats—are 

frequently identified as keystone species by scientists 

and ecologists because they influence ecosystems 

processes and populations of other species.

Distribution maps included in this chapter provide 

assistance to managers as to what species may be 

of concern when managing grasslands. All of the 

species distribution maps can be downloaded from 

a NatureServe’s Web site: http://www.natureserve.

org/getData/animalData.jsp.

Rodents are the largest and most diverse component 

of the mammalian faunas on grasslands. They range 

from the small harvest mouse (Reithrodontomys spp.) 

and pocket mouse (Perognathus spp.) to the large 

porcupine (Erethizon dorsatum) and beaver (Castor 

canadensis). The diets of small mammals reflect a 

diverse selection of food types that vary by mammal 

species, behavior, activity schedule, habitat, and location. Small mammal species range from being strictly 

herbivorous to omnivorous to mostly carnivorous. 

Rodents are nocturnal, diurnal, and crepuscular, depending on species. Most grassland rodents are strictly 

terrestrial and mostly fossorial (burrowing), while 

others are semiaquatic. Jones and Manning (1996) 

demonstrated that general habitat type (for example, 

riparian, tallgrass, shortgrass) influenced species 

distribution of rodents more than either the presence 

or absence of particular species of plants. Many of 

the heteromyid rodents (pocket mice and kangaroo 

rats) inhabit overgrazed areas and sparsely vegetated 

areas on sandy soils. The pocket mouse (Chaetodipus 

hispidus) occurs frequently in areas of early seral 

stage. In the arid Southwest, the species composition 

of rodent communities can be habitat-specific (Findley 

1989, Parmenter and Van Devender 1995).

Rodents have important roles in influencing habitat 

structure and composition of grasslands. These roles 

include dispersal of seeds, consumption and shredding of vegetation contributing to the deposition of 

humus, and mixing and aeration of soils by burrowing 

Bryce Rickel

Chapter 3:

Small Mammals, Reptiles, and 

Amphibians
36 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

activities. In addition, numerous rodents are major 

sources of food for predators.

General Effects of Grazing on 

Rodents and Other Small Mammals__

Effects of livestock grazing on small mammals can 

be variable, depending on the level of grazing, the 

type of grassland, and the particular small mammal 

species involved. Moderate grazing may have little or 

even a positive effect on many species, but overgrazing depresses populations of most small mammals. 

For example, heavy grazing and repeated fires in 

sagebrush range caused the establishment of nearly 

pure stands of annual grasses (Fagerstone and Ramey 

1996, Jones and Manning 1996) that support only a 

few species of deer mice (Peromyscus maniculatus) 

and Great Basin pocket mice (Perognathus parvus). 

In southern Idaho, rodent burrow numbers were significantly higher on ungrazed than on heavily grazed 

pastures. In a seldom-grazed pasture in Arizona, the 

total rodent population was roughly twice as high as 

on a heavily grazed pasture (Fagerstone and Ramey 

1996). On grasslands that are grazed heavily and 

are used continuously for decades, as they have been 

in areas of the Southwest, the resulting soil erosion 

reduces the quality of habitat for even grazing-tolerant 

species such as kangaroo rats (Dipodomys spp.) and 

prairie dogs (Cynomys spp.).

Results from studies suggest that the general composition of small mammal communities is determined 

primarily by structural attributes of the habitat. 

Livestock grazing affects many aspects of grassland 

ecosystems, including plant cover or biomass, plant 

species composition and diversity, primary productivity, 

soil compaction, and soil moisture. Plant cover probably 

has the most influence on small mammal populations 

because it provides food, nests, and protection from 

predators. Plant cover also influences behavioral interactions such as fighting and dispersal, and moderates 

ground level humidity, temperature, and soil moisture. 

There has been reported a significant positive relationship between small mammal abundance and canopy 

cover in sagebrush-grass grassland in Montana, where 

all areas were managed on a rest-rotation grazing 

system. Research has also shown that the percentage 

of forb cover was most consistently correlated with 

small-mammal species abundances, grass coverage 

was of lesser importance, and tree cover was not 

related to species abundance. These findings were 

consistent with that of other researchers, who reported 

that rodent abundance and diversity increased with 

vegetation cover and density and that overgrazing by 

cattle decreased vegetation complexity (Fagerstone 

and Ramey 1996).

Small mammal population responses to grazing 

depend on site characteristics and original composition 

of mammal species, and therefore, responses differ 

greatly among grassland types. Where there is sufficient vegetation in ungrazed grasslands to support 

herbivorous, litter-dwelling species, for example voles 

(Microtus spp. and Clethrionomys spp.), the small 

mammal communities are changed significantly by a 

reduction in cover caused by grazing. These changes in 

rodent communities are true for tallgrass and montane 

grasslands, which have significantly greater standing 

vegetation, greater annual net primary production, 

and greater abundance of mammals than shortgrass 

and bunchgrass grasslands (Fagerstone and Ramey 

1996).

Often when there are habitat modifications, small 

mammal communities shift in species composition and 

abundance. Decreases in vegetation cover in tallgrass 

and montane grasslands result in a decrease in total 

number of small mammals, an increase in small mammal species diversity, and a shift from litter-dwelling 

species with relatively high reproductive rates to surface-dwelling species with relatively low reproductive 

rates (Fagerstone and Ramey 1996). Microtines—voles 

and lemmings—dominated ungrazed tallgrass habitats 

(Payne and Caire 1999), with cricetines—harvest mice, 

deer mice, grasshopper mice, and woodrats—roughly 

half as abundant as microtines. In contrast, grazed 

tallgrass habitats were dominated by sciurids (chipmunks, marmots and squirrels) and heteromyids 

(pocket mice and kangaroo rats). In montane grasslands, grazing-induced reduction in cover resulted in 

similar decreases in total small mammal biomass and 

changes in species composition from litter-dwelling 

species to surface dwelling species, but the reduction 

in cover also resulted in a decrease rather than an 

increase in mammal species diversity. At montane 

sites, microtines dominated the ungrazed area, but 

cricetines dominated the grazed area. In shortgrass 

and bunchgrass grasslands, numbers of small mammal 

species and abundance were not changed drastically by 

reduction in vegetation cover by grazing (Fagerstone 

and Ramey 1996).

Small mammal communities of shortgrass and 

bunchgrass often are composed primarily of surfacedwelling, granivorous, and omnivorous species adapted 

to open habitats. On bunchgrass sites, sciurids and 

heteromyids were dominant, and on shortgrass sites, 

biomass was greatest for cricetines and sciurids, followed by heteromyids. A reduction in cover resulting 

from grazing may improve conditions for granivorous 

mammal species by promoting the abundance and seed 

production of annual grasses and forbs rather than 

perennial grasses (Fagerstone and Ramey 1996).

By affecting plant species diversity and vegetation 

structure, livestock grazing can influence rodent 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 37

species diversity. The effects of grazing on rodents can 

vary by habitat. Despite long-term protection from 

grazing, vegetation patterns on desert grasslands 

do not return to their original grass cover. This is 

because recovery takes a long time in desert (xeric) 

environments and because, once established, woody 

plants may competitively restrict the reestablishment of herbaceous cover and perennial grasses. For 

example, granivorous foragers such as the least chipmunk (Tamias minimus), Great Basin pocket mouse 

(Perognathus parvus), and deer mouse (Peromyscus 

maniculatus) have increased most in mesic habitats. 

In contrast, reduction of herbaceous vegetation by 

livestock grazing has resulted in a reduction in plant 

diversity and rodent diversity in xeric communities 

(Fagerstone and Ramey 1996).

Research has demonstrated that small mammals 

that used riparian areas may be significantly impacted 

by grazing. In a comparison of small mammal populations before and after late season (late August to 

mid-September) grazing, population estimates were 

lower in all grazed riparian habitats than in ungrazed 

habitats (table 3-1). The significant difference between 

small mammal populations in grazed versus ungrazed 

riparian areas was apparently related to loss of cover 

due to forage removal. Reduced cover resulted in 

increased predation on small mammals and their 

emigration from grazed habitats into neighboring 

ungrazed habitats (Fagerstone and Ramey 1996).

Sylvatic Plague_____________________

Sylvatic (bubonic) plague (Yersinia pestis) has been 

in the United States for approximately 100 years and in 

black-tailed prairie dog populations for approximately 

50 years. This exotic disease was first observed in wild 

rodents in North America near San Francisco, CA, in 

1908. The first reported incidence of plague in blacktailed prairie dogs (Cynomys ludovicianus) occurred 

in Texas in 1946 (Gober 2002). Some rodents may act 

as hosts or carriers of the disease and show little or 

no symptoms, but prairie dogs do not develop effective 

antibodies or immunity to the disease (Gober 2002). 

The plague is transmitted by fleas and decimates 

prairie dogs (Knowles 2002). It has spread through 

the West and Southwest where a variety of rodent 

species serve as reservoirs for passing the disease 

to humans and wildlife (Brand 2002). Of the three 

major factors (habitat loss, poisoning, and disease) 

that currently limit the abundance of black-tailed, 

white-tailed (C. leucurus), Gunnison’s (C. gunnisoni), 

and Utah (C. parvidens) prairie dogs, sylvatic plague is 

the one that is currently beyond human control (Cully 

and Williams 2002). When colonies are infected, the 

mortality of prairie dogs is often as high as 90 to 100 

percent (Brand 2002).

The plague has the potential to reduce prairie dogs 

to levels lower than encountered during organized 

poisoning campaigns. And directed prairie dog poisoning, in concert with the plague, has the potential of 

extirpating prairie dogs from large areas, resulting 

in fragmented and isolated prairie dog populations 

persisting over the long term. Repeated catastrophic 

events (plague epizootics and poisoning) will progressively drive prairie dog populations toward extinction 

(Knowles 2002).

The plague is the major reason for the declines 

in Gunnison’s prairie dog populations today. The 

only area where plague appears not to have had an 

impact is possible Aubrey Valley, Arizona, which has 

no documented plague outbreaks and retains large 

prairie dog colonies. Plague entered the range of the 

Gunnison’s prairie dog during the late 1930s to the late 

1940s. Published accounts for Gunnison’s prairie dogs 

show that mortality from plague frequently exceeds 

99 percent. Bureau of Land Management biologists 

who have Gunnison’s prairie dogs within their area of 

jurisdiction have reported that, due to plague, there 

are no large colonies, 200 acres (81 ha) being the upper 

size limit of a colony (Knowles 2002).

Some populations have had no significant recovery, 

such as Gunnison’s prairie dogs in South Park, CO. In 

northern New Mexico, Gunnison’s prairie dogs partially 

recovered following an initial plague epizootic but failed 

to recover following a second epizootic. Other reports 

suggest a sequence where colonies are regularly lost 

due to plague, then new colonies develop and grow in 

other areas; this pattern may yield populations that are 

stable over a larger geographic area. Similar reports 

have come from northern Arizona, where there have 

been substantial declines due to plague. However, at 

the same time, Arizona’s largest complex has been 

increasing 8 percent annually since 1992. There 

are concerns that plague cycles result in successive 

population peaks that are progressively lower than 

the previous peak. There are also concerns that with 

each new epizootic, the loss of colonies from plague 

will exceed the rate of establishment of new colonies 

(Knowles 2002).

Observations of these patterns to date are largely 

anecdotal and not based on careful mapping. However, 

Table 3-1. Rodent populations in three grazed and ungrazed 

riparian habitats (Fagerstone and Ramey 1996).

Community Mammals/ha

 type Grazed Ungrazed

Hawthorne (Crataegus spp.) 800 to 83 690 to 136

Meadow 450 to 60 235 to 463

Cottonwood-mixed conifer 129 to 42 118 to 254
38 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

in New Mexico and Colorado, plague impacts for the 

Gunnison’s prairie dog are well documented. South 

Park, CO, was described as containing 913,000 acres 

(369,480 ha) of Gunnison’s prairie dog colonies in 

1941 prior to the advent of plague. Plague entered 

this area in 1947, and by only 2 years later plague 

had reduced the prairie dog acreage by more than 95 

percent. Epizootics of plague continued in this area 

through the 1950s and 1960s and prairie dogs were 

nearly eliminated from South Park. Currently this 

area contains only a few hundred acres of prairie dog 

colonies. Former colonies are now occupied by Wyoming 

ground squirrels (Spermophilis elegans) and thirteenlined ground squirrels (Spermophilis trzdecemlineatus)

(Knowles 2002).

The black-tailed prairie dog is less susceptible to 

plague than the Gunnison’s species (Knowles 2002). 

Plague affects black-tailed prairie dog populations by 

reducing colony size, increasing population variance 

within colonies, and increasing intercolony distances 

within colony complexes. In the presence of plague, 

black-tailed prairie dogs will probably survive in 

complexes of small colonies greater than 3 km (1.9 

mi) from their nearest neighbor colonies or colonies 

that undergo severe population fluctuations (Cully 

and Williams 2002).

Currently plague is widespread throughout 66 

percent of the historic range of the black-tailed prairie dog including all of Arizona, Colorado, Montana, 

New Mexico, Texas, and Wyoming, and portions of 

Kansas, Nebraska, North Dakota, and Oklahoma. 

South Dakota is the only State within the range of 

the species where plague in black-tailed prairie dogs 

has not been documented, although plague antibody 

titers (Knowles 2002) have been detected in badger 

(Taxidea taxus), coyote (Canis latrans), and red fox 

(Vulpesfulva) collected in the southwestern portion of 

the State (Gober 2002).

The endangered black-footed ferrets (Mustela 

nigripes) are extremely susceptible to sylvatic plague 

(CBSG 2004, Hatfield-Etchberger and others 2002). The 

threat to the black-footed ferret comes from both direct 

mortality from plague and indirectly from decimation 

of prairie dogs, their sole food source (Brand 2002).

Plague can infect humans also. Some 10 to 15 cases 

of plague in humans have been reported each year in 

the United States since 1975. Wild rodents, particularly rock squirrels (Spermophilus variegatus), are 

frequently shown or implicated to be the reservoir for 

infecting fleas that then transmit the plague bacterium 

to humans via domestic cats. Increased risk for plague 

in humans is associated with expansion of residential 

areas into areas populated by rodent reservoirs, and 

with pastoral human life styles, particularly among 

Native Americans on reservations in the southwestern 

United States.

Plague control in wildlife in the United States 

has been attempted on numerous occasions in direct 

response to human cases of plague or proactively to 

reduce risk of transmission to humans within a relatively localized geographic area. These programs often 

involve rodent and vector population suppression in 

addition to public education and medical surveillance. 

While these efforts to reduce the incidence of plague 

have met with varying degrees of success, in general 

they do not contain the disease for long periods or over 

broad geographical areas.

Large Rodents_____________________

Prairie Dogs

Prairie dogs (Cynomys) are unique to North America. 

Five species within the genus inhabit grasslands of 

central North America from southern Canada to northeastern Mexico (Hof and others 2002, Wagner and 

Drickamer 2004). The Mexican prairie dog is the only 

one that does not occur in the United States. The four 

species that do reside in the United States are the blacktailed prairie dogs and three species of white-tailed 

prairie dogs—white-tailed prairie dog, Gunnison’s 

prairie dog, and Utah prairie dog. Black-tailed and 

Gunnison’s prairie dogs occur in the Southwestern 

Region of the Forest Service. Populations of the blacktailed prairie dog are distributed in New Mexico, 

Texas, Oklahoma, Kansas, and Colorado grasslands; 

Gunnison’s prairie dogs are found in Arizona and New 

Mexico (fig. 3-1) (Fagerstone and Ramey 1996). In New 

Mexico, black-tailed prairie dogs occurred historically 

in the southwestern, southeastern, and northeastern 

parts of the State, while Gunnison’s and Utah prairie 

dogs occurred in the Great Basin.

Prairie dogs typically live in towns of 1,000 acres 

(400 ha) or larger. The rodent occupied up to 700 million acres of western grasslands in the early 1900s. 

The largest prairie dog colony on record, in Texas, 

measured nearly 25,000 square miles (65,000 km2) 

and contained an estimated 400 million prairie dogs 

(Knowles 2002). Larger towns are divided into wards 

by barriers such as ridges, lines of trees, and roads. In 

a ward, each family or “coterie” of prairie dogs occupies 

a territory of about 1 acre (0.4 ha). A coterie usually 

consists of an adult male, one to four adult females, 

and any of their offspring that are less than 2 years 

old. Members of a coterie maintain unity through a 

variety of social activities, for example calls, postures, 

displays, grooming, and other forms of physical contact 

(Hygnstrom and Virchow 1994).

Prairie dogs are largely herbivorous, feeding

preferably on grasses, 62 percent to 95 percent of 

their diet, and on forbs when they are the dominant 

vegetation. During certain periods of the year, they 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 39

may feed heavily on seeds (Hygnstrom and Virchow 

1994, Knowles 2002). All prairie dogs are capable of 

living without free water, obtaining their water from 

what they eat. Gunnison’s prairie dogs are hibernators 

and may even estivate during late summer (Knowles 

2002).

Prairie dogs are most active during the day. In the 

summer during the hottest part of the day, they go 

below ground where it is cooler. Black-tailed prairie 

dogs are active all year, but may stay under ground 

for several days during severe winter weather. The 

Gunnison’s prairie dogs hibernate from October 

through February (Hygnstrom and Virchow 1994, 

Sevilleta LTER: Data 1998c).

Since 1900, prairie dog populations have been 

reduced by as much as 98 percent in some areas and 

totally eliminated in others. This demise is largely the 

result of cultivation of prairie grasslands and control 

programs implemented in the early and mid-1900s (Hof 

and others 2002, Hygnstrom and Virchow 1994) and the 

plague (Brand 2002). Because prairie dogs feed upon 

gasses and upon a variety of annuals, they compete 

with domestic livestock for food. As a result, humans 

have made great efforts to eliminating the “pest” from 

rangeland, chiefly by poisoning. For example, in 1908, 

Vernon Bailey, working for the U.S. Biological 

Survey, traveled from Deming to Hachita and 

through the Animas and Playas Valleys in what 

is now southern Hidalgo County. He reported that 

the area was one continuous prairie dog town, and 

estimated that the county contained 6.4 million 

animals. In numerous trips through exactly the 

same region from 1955 through 1972, workers from 

the Museum of Southwestern Biology never saw a 

single prairie dog. Similar devastation has occurred 

in many parts of the mammal’s former range. In 

this respect, black-tail prairie dogs have suffered 

more than Gunnison’s. The latter species may be 

seen more or less regularly in various parts of 

northwestern New Mexico. Both species are subject 

to the plague and are periodically decimated by the 

disease (Sevilleta LTER: Data 1998c). Population 

increases have been observed in the 1970s and 

1980s, possibly due to the increased restrictions 

on and reduced use of toxicants (Hygnstrom and 

Virchow 1994).

Early accounts of the black-tailed prairie dog 

suggest that this was an abundant species on the 

Great Plains. Although we lack similar accounts 

of the white-tailed and Gunnison’s prairie dogs, it 

is assumed that these were also highly successful 

within their distributional range (Knowles 2002). The 

1900s saw drastic declines for all prairie dog species. 

Although the prairie dog distributional range has 

not contracted greatly, it is estimated that overall 

black-tailed prairie dog populations have declined by 

98 to 99 percent (Hof and others 2002, Knowles 2002, 

Wagner and Drickamer 2004).

In addition to the plague and poisoning, recreational 

shooting has affected prairie dog densities and populations (Knowles 2002). Pauli (2005), in his study on the 

effects of recreational shooting on black-tailed prairie 

dogs, found shooting caused a reduction in a colony of 

30 percent in 1 year. He also found:

• Survivors exhibited an eight-fold increase in 

alert behavior.

• Aboveground activity was reduced by 66 percent, 

which reduced the time spent foraging.

• These behavioral changes resulted in 35 percent 

decrease in the body condition of the survivor.

• Flea load increased 30 percent.

• Fecal corticosterone—a steroid hormone produced 

in the adrenal glands that functions in the metabolism of carbohydrates and proteins—increased 

80 percent.

• After shooting, the pregancy rates declined 50 

percent, and the reproductive output decreased 

by 76 percent.

The results from Pauli’s study indicate that the stress 

caused by the shooting is long lasting and affects the 

colony and not just individuals.

Figure 3-1. Distribution of black-tailed (Cynomy ludovicianus) 

and Gunnison's (C. gunnisoni) prairie dogs in North America 

(adapted from Fagerstone and Ramey 1996).
40 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Importance of Prairie Dogs—Knowles (2002) 

appropriately states that the importance of prairie 

dogs to the grassland ecosystems of North America is 

matched only by the degree to which that importance 

is misunderstood, misrepresented, and minimized. 

They probably had a more profound influence on the 

physiognomy and composition of native grassland communities than most other mammal species (Fagerstone 

and Ramey 1996, Hof and others 2002). Probably the 

only other species that played such a significant role 

in grassland structure on the Great Plains was the 

bison. Prairie dog colonies often encompassed huge 

grassland expanses and their maintenance of these 

areas for colonial use influenced both abiotic and biotic 

conditions (Fagerstone and Ramey 1996).

Prairie dogs are identified as keystone species in the 

ecosystems they inhabit (Cook and others 2003, Kotliar 

2000). A keystone species is one whose impact on its 

community or ecosystem is large, and disproportionately large relative to its abundance (Payton and others 

2002) (see discussion in chapter 1). Prairie dogs play a 

keystone role in maintaining grassland ecosystems. For 

example, dozens of species of mammals, birds, reptiles, 

and amphibians are dependent to one degree or another 

on prairie dogs for food, shelter, or both. Without the 

prairie dog, the vast American grassland ecosystems 

cannot survive (Knowles 2002). Much of the research 

on these associated species has been conducted within 

the range of the black-tailed prairie dogs. Reports of 

up to 117 wildlife species associated with prairie dogs 

may overestimate the total number, but many species 

are benefited by prairie dogs. These close associates 

appear to use white-tailed and Gunnison’s prairie dog 

colonies, as well as black-tailed prairie dog colonies 

(Knowles 2002).

Effects of Prairie Dogs on Grassland—Prairie 

dogs colonize sites where the vegetation is low due 

to heavy grazing or to other disturbance that reduce 

vegetation height and density, thus allowing a good 

view of predators. In well-established prairie dog 

colonies, large areas of bare soil are common. Where 

there is low vegetation, they often clip shrubs and 

other tall vegetation to maintain a condition where 

plant species composition, biomass, and productivity 

of vegetation differ from uncolonized areas (Fagerstone 

and Ramey 1996).

High densities of prairie dogs may negatively influence native perennial grasses by causing shifts in 

plant species composition toward shorter grasses and, 

ultimately, toward annual and short-lived perennial 

forb species (Fagerstone and Ramey 1996, Severe 

1977). Often buffalograss (Buchloe dactyloides) is 

the dominant plant on prairie dog colonies, and the 

taller western wheatgrass (Agropyron smithii) and 

blue grama (Bouteloua gracilis) are most common on 

uncolonized mixed-grass prairie sites. In areas with 

the greatest prairie dog activity, annual forbs, shrubs, 

and cacti often replace most of the original grass 

cover. The formation of forb-dominated communities 

in prairie dog colonies is related to the length of time 

since colonization and the level of prairie dog activity; 

forb domination is usually greatest in the center of 

the colony (Fagerstone and Ramey 1996). They will 

forage on the following forbs: scarlet globe- mallow 

(Sphaeralcea coccinea), prickly pear (Opuntia spp.), 

kochia (Kochia scoparia), peppergrass (Lepidium spp.), 

and wooly plantain (Plantagopatagonica) (Hygnstrom 

and Virchow 1994).

Positive effects of prairie dogs on grassland productivity include greater soil aeration, changes in 

community structure, increased plant species diversity, 

and greater forb production. It has been postulated 

that burrowing decreases soil compaction, increases 

water absorption, aerates soil, and promotes soil 

formation. Soils in prairie dog colonies are richer in 

nitrogen, phosphorus, and organic matter than soils 

in adjacent grasslands (Fagerstone and Ramey 1996, 

Severe 1977).

Prairie dog foraging removes aging leaves and may 

stimulate growth of new plant tissue, which usually 

has a higher nutritional value (increased nitrogen 

concentration) than older tissue. Prairie dog colonies, 

therefore, have been found to contain better quality 

forage and growing conditions than uncolonized areas 

(Fagerstone and Ramey 1996).

On February 4, 2000, the USDI Fish and Wildlife 

Service announced its 12-month finding for a petition to list the black-tailed prairie dog as Threatened 

throughout its range under the Endangered Species Act 

Black-tailed prairie dog. (Photo by Jeff Venuga)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 41

of 1973. They determined that listing was warranted 

but precluded by other higher priority actions. The 

black-tailed prairie dog was added to the candidate 

species list.

Competition Between Prairie Dogs and 

Cattle—The degree of dietary competition and overlap 

between prairie dogs and cattle can be high. Both eat 

mainly grasses, followed by forbs and shrubs. However, 

eliminating prairie dogs has had little effect on increasing the amount of food available for cattle (Fagerstone 

and Ramey 1996). At a prairie dog repopulation rate 

of 30 percent, controls have not been economically 

feasible, and annual maintenance costs are greater 

than the amount of forage gained. Controlling blacktailed prairie dogs on depleted grasslands in western 

South Dakota did not increase the amount of forage 

produced after 4 years, whether or not cattle were 

allowed to graze (Fagerstone and Ramey 1996). To 

improve range conditions, long periods, up to 10 years, 

of total exclusion from prairie dogs and livestock may 

be required when the range is in a low condition class. 

In a study discussed by Fagerstone and Ramey (1996), 

prairie dog-cattle competition was found to have no 

differences in forb production on steers-only pastures 

compared to pastures with steers and prairie dogs. They 

did find significant reductions in availability of blue 

grama, sand dropseed (Sporobolus cryptandrus), and 

other grasses on pastures with prairie dogs. Uresk and 

Paulson (1988) estimated the carrying capacity and 

forage utilization for cattle in western South Dakota 

when prairie dogs were present, but pastures were 

maintained in good condition at a near climax stage 

of mixed perennial cool-season grasses. They found 

that carrying capacity for cows and for cow-calf units 

decreased as the number of hectares occupied by prairie 

dog colonies increased; the decrease was approximately 

three AUMs or two cow-calf units for every additional 

20 ha of prairie dogs. The researchers showed that on 

such sites, needle leaf sedge (Carex eleocharis) and 

needlegrasses (Stipa spp.) could become major limiting 

factors in determining cow carrying capacity.

Interactions Between Prairie Dogs and Other 

Wildlife Species—As a keystone species, prairie dogs 

have great influence on other wildlife species. Through 

modifications of aboveground vegetation, prairie 

dogs influence the densities, foraging patterns, and 

nutritional dynamics of other animals. Because these 

habitat modifications can be extensive, researchers 

refer to “prairie dog ecosystems”—that is, they are 

systems comprised of prairie dogs and other associated 

plants and animals. Studies have reported 64 to 163 

vertebrate species associated with prairie dog colonies 

(Fagerstone and Ramey 1996, Hygnstrom and Virchow 

1994). Five classes of invertebrates were identified 

on prairie dog colonies in South Dakota, which may 

explain why more insectivorous rodent species are 

found on prairie dog colonies than on surrounding 

grassland (Fagerstone and Ramey 1996, Hygnstrom 

and Virchow 1994).

One of the most important features of a prairie dog 

colony is the burrow system. Prairie dog burrows serve 

as homes for various small mammals, reptiles, birds, 

amphibians, and invertebrates whose numbers are 

usually higher on prairie dog colonies (Fagerstone and 

Ramey 1996, Hygnstrom and Virchow 1994, Knowles 

2002). On mixed-grass sites in South Dakota and 

Oklahoma, small rodent abundance was found to be 

greater on than off colonies, but small rodent species 

richness was significantly lower. Lower species richness 

can probably be attributed to changes in vegetation 

structure and composition in colonies (Fagerstone and 

Ramey 1996).

Prairie dogs are also important prey for some mammalian predators. The black-footed ferret has been 

historically an important predator of prairie dogs. The 

population decline of black-footed ferrets to the point 

where they are now listed as Federally Endangered 

is related to the decrease in prairie dog populations. 

In the absence of ferrets, the badger (Taxidea taxus) 

is the main prairie dog predator (Fagerstone and 

Ramey 1996). Others mammalian predators include 

coyotes (Canis latrans), bobcats (Lynx rufus), foxes 

(Vulpes spp.), occasionally mink (Mustela vison), and 

long-tailed weasels (Mustela frenata) (Fagerstone and 

Ramey 1996). Many avian predators feed on prairie 

dogs, including golden eagles (Aquila chrysaetos), bald 

eagles (Haliaeetus leucocephalus), Ferruginous hawks 

(Buteo regalis), red-tailed hawks (B. jamaicensis), 

rough-legged hawks (B. lagopus), marsh hawks (Circus 

cyaneus), and other species (Fagerstone and Ramey 

1996). Young prairie dogs may be taken by prairie rattlesnakes (Crotalus viridis) and bullsnakes (Pituophis 

catenifer), but rarely are adult prairie dogs prayed 

upon by snakes (Hygnstrom and Virchow 1994).

Bird species diversity and abundance are significantly higher on prairie dog colonies than on mixed-grass 

sites (Fagerstone and Ramey 1996). They attributed 

the higher numbers to “patchiness” or structural diversity on prairie dog colonies, to increased forb seed 

production, and to lower amounts of mulch and lower 

vegetation height, which may result in greater visibility 

of macroarthropods and seeds. Bird species that are 

significantly more abundant on prairie dog colonies 

include horned larks (Eremophila alpestris), mourning 

doves (Zenaidura macroura), killdeer (Charadrius 

vociferus), barn swallows (Hirundo rustica), and burrowing owls (Athene cunicularia) (Fagerstone and 

Ramey 1996).

Prairie Dog Management Programs—Land 

managers are learning to account for the positive and 

negative effects of prairie dogs on grasslands and on 

other wildlife species in land management planning. 
42 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Recognizing the important ecological roles 

prairie dogs play within grasslands is critical 

for the overall management of these ecosystems. Management of prairie dog colonies and 

grasslands should take into consideration a 

number of factors including range conditions 

and trends, season of livestock use, prairie 

dog and livestock densities, how much area is 

available for colony expansion, maintaining 

habitats with a large component of appropriate vegetation, and potential interactions with 

other species.

Prairie dogs most frequently colonize sites 

that have been overgrazed or otherwise disturbed. Intense livestock grazing promotes 

high prairie dog densities, and colonies tend 

to expand under heavy grazing. Depending on 

climatic factors, prairie dog colony expansion 

rates can be decreased by increasing vegetation 

cover around colonies through reducing grazing 

and human disturbance.

Pocket Gophers

Pocket gophers are fossorial (burrowing) 

rodents that have gnawing teeth for chewing 

(Pocket Gopher 2005c, Wiscomb and Messmer 1988). 

The name pocket gopher comes from the pouches on 

their cheeks. They belong to the family Geomyidae 

(Geomyidae 2005, Pocket Gophers 2005a). There 

are 33 species of pocket gophers represented by five 

genera in the western hemisphere. The two primary 

genera of pocket gophers discussed here are Geomys

spp. and Thomomys spp. Geomys are present from 

the Rocky Mountains east to the Mississippi River, 

and from southern Canada to southern Texas. The 

three main Geomys species in North America are the 

plains pocket gopher (Geomys bursarius), the desert 

pocket gopher (G. arenarius), and the Texas pocket 

gopher (G. personatus). The plains pocket gopher is 

the most widespread (fig. 3-2). Thomomys species occur 

generally in the Western States. The northern pocket 

gopher (T. talpoides) (fig. 3-2) is widely distributed 

from Canada south to northern California and New 

Mexico, and from the West Coast east to the Dakotas 

(fig. 3-2) (Fagerstone and Ramey 1996).

Pocket gophers occur on pastures, grasslands, 

prairies, roadsides, and railroad rights-of-way (Pocket 

Gopher 2005c) or any disturbed land. They live in a 

broad range of habitats from deserts to mountain 

meadows, in soils ranging from sand to clay, with loam 

preferred. In valleys and mountain meadows, they 

prefer loamy soil, but some occur in sandy or rocky 

situations. The soil in which a gopher will dig its burrow seems to be dependent on the size of the animal 

and related to depth and friability of the soils. Larger 

gophers lived in deep, soft soils and small animals in 

shallow, rocky ones (Sevilleta LTER: Data 1998b).

In Arizona and New Mexico, botta’s pocket gophers 

live in nearly every habitat within the States so long 

as sufficient tuberous roots and plant material are 

available and soil is suitable for digging tunnels. They 

are found near sea level up to 11,000 feet. They live in 

extremely xeric deserts through all of the vegetative 

types to near timberline in the mountains (Sevilleta 

LTER: Data 1998b, Sullivan 2005).

These rodents, weighing less than a pound, are living mining machines. Where the digging is easy, they 

are able to tunnel as much as 200 to 300 feet (61 to 91 

m) in a single night. A burrow may be occupied by the 

same animal for several years, and burrows may occur 

in densities of up to16 to 20 per acre (6.4 to 8 per ac). 

Burrow systems consist of a main tunnel from 4 to 18 

inches (10 to 46 cm) below the surface with a number 

of lateral tunnels branching off from the main tunnel 

(Forest Preserve District of Cook County 1973, Pocket 

Gophers 2005c, Sullivan 2005). Lateral tunnels end at 

the surface where the soil mound is created.

Pocket gophers usually construct one nest that 

contains a number of toilets and a number of food 

cache chambers in deeper tunnels that branch off from 

Figure 3-2. Distribution of the northern (Thomomys talpoides), 

plains (Geomys burarius), and desert (G arenarius) pocket 

gophers (NatureServe 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 43

the main tunnel. A nest chamber also is lined with 

vegetation. Nest chambers and food caches have been 

found as deep as 5 to 6 feet (1.5 to 1.8 m) below the 

surface. During the breeding season a male’s burrow 

may be more linear because its sole purpose is to intercept a female’s burrow. A single pocket gopher may 

construct as many as 300 soil mounds in a year while 

moving more than 4 tons (3629 kg) of soil. Burrows are 

continually changing, with old tunnels being sealed off 

and new ones excavated. A single tunnel system may 

consist of as much as 200 yards (183 m) of tunnels. In 

habitat with poorer vegetation, longer tunnels must be 

excavated to meet food needs (Forest Preserve District 

of Cook County 1973). Gophers seal the openings to 

the burrow system with earthen plugs (Pocket Gophers 

2005b,c, Sullivan 2005).

Pocket gophers feed on a wide variety of herbaceous 

material (Pocket Gopher 2005c, Sevilleta LTER: Data 

1998b, Wiscomb and Messmer 1988). Above ground, 

from the vicinity of burrow openings, they take leafy 

vegetation, generally preferring herbaceous plants, 

shrubs, and trees to grass; most commonly they feed 

on roots and fleshy portions of plants while digging 

underground devouring succulent roots and tubers. 

They often prefer forbs and grasses, but diet shifts 

seasonally according to the availability and needs for 

nutrition and water. For example, water-laden cactus 

plants may become a major dietary component during the hot and dry summer months in arid habitats. 

Gophers will pull entire plants into their burrow from 

below. In snow-covered areas they may feed on bark 

several feet up a tree. Pocket Gophers are active all 

year, day and night, and guard their burrows and 

territories fiercely (Pocket Gopher 2005a,b,c, Forest 

Preserve District of Cook County 1973, Wiscomb and 

Messmer 1988). Gophers do not hibernate (Pocket 

Gopher 2005b, c). Although pocket gophers are usually 

solitary (Pocket Gopher 2005c, Sevilleta LTER: Data 

1998b), occasionally a male and female will be found in 

the same burrow on the same day. This probably occurs 

most frequently during their breeding season. A male 

probably mates with several females, especially those 

with burrow systems adjacent to his. This polygamous 

behavior results in a large number of females in the 

population. Some males practice serial monogamy; 

researchers found four cases where a male and female 

were sharing a nest (Sevilleta LTER: Data 1998b).

Gophers are prayed upon by hawks, owls, snakes, 

badgers, foxes, and coyotes (Sullivan 2005). Badgers 

and coyotes hunt pocket gophers by digging out their 

burrows, while weasels and snakes may pursue them 

underground. Other predators include skunks, owls, 

bobcats, and hawks (Desert USA 2005).

Pocket gophers are found throughout most of the 

grasslands in the United States (Fagerstone and 

Ramey 1996, Pocket Gopher 2005c). That pocket 

gophers play a vital role in the functioning of grassland ecosystems becomes evident as we consider their 

roles as an ecosystem engineer and prey species, their 

influence in loosening, stirring, and enriching the soil 

(Forest Preserve District of Cook County 1973, Pocket 

Gophers 2005a, Reichman 2004), their effects on microtopography and in creating habitat heterogeneity, 

their effects on plant species diversity and primary 

production, and their role as providers of habitat for 

other wildlife species, such as, rabbits, ground squirrels, mice, skunks, snakes, lizards, and toads (Pocket 

Gophers 2005a, Sullivan 2005).

Gopher’s Effects on Grasslands—Pocket gophers 

are an important element controlling ecosystem 

structure and development. It has been argued that 

Geomyidae is a dynamic force to direct the biogeochemical attributes of the North American grasslands. The 

activities of gophers may provide an explanation for 

the genesis of North American Prairie soils. Native 

plant life on hill and mountainside in canyon and 

mountain meadow would soon begin to depreciate if 

gopher populations were to be completely destroyed 

(Huntly and Inouye 1988).

Gopher effects on the productivity, heterogeneity, 

and trophic structure of ecosystems, of various temporal and spatial scales have been described. Gophers 

influence the physical environment, altering patterns 

and rates of soil development and nutrient availability, 

microtopography, and the consequent abiotic environment. They affect the demography and abundance 

of plant species, changing vegetational patterns and 

diversity. They affect the behavior and abundance of 

other herbivores, from grasshoppers and ground squirrels to large grazers (Huntly and Inouye 1988).

Pocket gophers may be a keystone species in grasslands. The plains pocket gopher (Geomys bursarius) 

turns over as much as 5 percent of the tallgrass 

prairies per year. This disturbance creates openings 

in the grassland canopy that may allow seedlings to 

establish. It has been demonstrated that gopher mound 

building negatively affects the activity of meadow 

voles (Microtus pennsylvanicus), a major aboveground 

herbivore. This, in turn, allows a greater proportion of 

seedlings to escape predation by voles. Thus, through 

these direct and indirect effects, gophers may be instrumental in structuring the prairie plant and animal 

communities as well as maintaining prairie diversity 

(Geomyidae 2005).

Pocket gophers affect grassland in three important 

ways: (1) by burying plants; (2) by transporting nutrients to the soil surface during burrowing and mound 

formation activities; and (3) by feeding, which decreases 

biomass of forage plants and alters plant species composition. Pocket gophers compete directly with livestock 

by consuming range plants, above and below ground. 

Consumption of forage by gophers is much higher than 
44 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

for other small mammals (Fagerstone and Ramey 1996, 

Forest Preserve District of Cook County 1973).

Pocket gophers may be the primary non-ungulate 

consumer of forage in grasslands, frequently harvesting 

and storing more vegetation than they actually eat. 

What they do not eat, they store in underground food 

caches (Forest Preserve District of Cook County 1973, 

Pocket Gophers 2005a,b). The plains pocket gopher 

differs from other pocket gophers in that forbs comprise 

a smaller portion of their diet (Fagerstone and Ramey 

1996). In contrast to Thomomys species, Geomys species 

frequently thrive in grassland areas with few forbs. 

Various studies have shown that grasses were either 

the majority or near majority portion of Geomys diet 

(Fagerstone and Ramey 1996).

Several studies demonstrate that pocket gophers can 

decrease grassland forage production by consumption, 

clipping, burying litter and vegetation, and reducing 

plant vigor. The decrease in production varies between 

shortgrass and tallgrass and between range condition classes within sites. Plains pocket gophers have 

been shown to significantly impact forage production 

on western Nebraska grasslands, decreasing overall 

production between 18 percent and 49 percent. In 

Texas, biomass increased 22 percent when plains pocket 

gophers were excluded from grasslands (Fagerstone 

and Ramey 1996).

Besides changing forage availability, pocket gophers can alter the vegetation species composition 

of grasslands by feeding, burying herbage, and by 

altering the microenvironment. Plant species favored 

by gophers tend to decrease on grasslands while 

unpalatable species increase. Pocket gopher feeding 

and burrowing activity promotes the presence of 

annual grasses, annual forbs, and perennial forbs, 

while decreasing the frequency of perennial grasses. 

These changes in plant composition are related to 

precipitation. The greatest changes occur in areas with 

low precipitation. Pocket gophers have been known 

to cause major changes in vegetation composition on 

high mountain grasslands, suppressing productivity 

of some livestock forage species such as common 

dandelion (Taraxacum officinale), lupine (Lupinus

spp.), agoseris (Agoseris glauca), and aspen peavine 

(Lathyrus leucanthus), and increasing production 

of orange sneezeweed (Helenium hoopesi), which is 

poisonous to sheep and unpalatable to cattle. Two 

grass species palatable to livestock, slender wheatgrass (Agropyron pauctflorum) and mountain brome 

(Bromus carinatus), benefited from gopher activity. 

Range condition may decline following pocket gopher 

occupation as desirable perennial grasses decline, 

accompanied by an increase in annual grasses and 

forbs. This decline may cause gophers to move into 

previously unoccupied areas and abandon the weedy 

areas (Fagerstone and Ramey 1996).

As pocket gophers dig, they deposit soil that may 

bury vegetation and prevent growth of the underlying 

vegetation. Subsequent plant succession on denuded 

areas may be slow and may continually provide colonization sites for early successional species. Vegetation 

density on pocket gopher mounds increases rapidly 

over time as perennial species replace less desirable 

annuals and forbs. The first plants to appear are usually annuals, followed by perennial dicots. It has been 

observed that herbaceous perennial dicots benefit from 

pocket gopher disturbance by germinating and surviving in greater numbers on mounds than off mounds 

(Fagerstone and Ramey 1996).

Digging by gophers may cause higher erosion rates 

than those attributed to other processes in an area. 

Most digging occurs in late summer and fall when young 

gophers establish their own burrow systems and when 

adults extend their burrows in search of underground 

food. Estimates on the amount of soil brought to the 

surface by pocket gophers range from 4,483 kg to 85,200 

kg (4.9 to 94 tons). This huge variance in displacement 

of soil by gophers may result in formation of mima 

mounds (Fagerstone and Ramey 1996, Geiger 2002). 

Such mounds are usually 0.3 to 0.9 m (1 to 3 ft) high 

and 4.5 to 30 m (14.7 to 98.4 ft) in diameter and are 

formed over decades of gopher burrowing activities 

that tend to move soil toward the nest. The soil on 

mima mounds may differ considerably from adjacent 

soils, having a lower bulk density, higher water permeability, higher organic matter content, and a lack 

of definite structure in the topsoil. Stones of the sizes 

pocket gophers can move are concentrated in mounds, 

and vegetation on mounds is usually denser and more 

effective in retarding soil erosion than that off mounds. 

In a Colorado range seeding project, grasses produced 

two to five times more herbage on mounds than between 

mounds (Fagerstone and Ramey 1996).

Results from different studies conflict as to how 

nutrient cycling may be affected by deposition of soil by 

pocket gophers. Increases in organic matter, nitrogen, 

and phosphorus in areas occupied by pocket gophers 

have been reported. In contrast, other researchers 

reported reduced nutrients in occupied areas. Still 

other studies looked at the nutrient content of soil 

samples from old mounds, new mounds, and away from 

mounds (controls), and found that old mounds were 

often significantly lower in nutrient concentrations 

than new mounds, which were lower than control sites. 

Gophers reduced the average nitrogen concentration 

near the soil surface and increased the variability in 

soil nitrogen. Soil deposited by pocket gophers was 

lower in nutrients (that is, phosphorus, nitrate, and 

potassium) than randomly collected samples, possibly 

because nutrients were leached out or drawn from soil 

by plant roots. Deficiency of nutrients in mounds may 

also occur because mounds lack the litter layer that 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 45

is important in holding moisture and nutrients near 

the surface (Fagerstone and Ramey 1996).

Pocket gophers may benefit grasslands by loosening 

compacted soil, allowing better aeration, improving 

water infiltration, and increasing soil fertility by adding 

excrement and burying vegetation. The decrease in 

plant biomass caused by mound-building was partly 

compensated for by increased production in areas 

immediately adjacent to mounds, where production 

was higher than vegetative production near the edge 

of mounds. Researchers hypothesized that increased 

density of vegetation near mounds was a response to 

increased nutrient availability caused by leaching of 

nutrients from mounds into surface soil (Fagerstone 

and Ramey 1996).

Grazing Effects on Pocket Gophers—Pocket 

gophers are attracted to grasslands in good to excellent range condition where they use vigorous plants 

with large root systems. Their densities appear to be 

dependent on plant biomass. Lower plant biomass 

may require gophers to burrow more extensively to 

locate food. The impacts of grazing on pocket gophers 

are variable. In two studies of northern pocket gopher 

populations, no significant differences were found 

between grazed or ungrazed areas on mountain 

rangeland. Other studies have shown higher pocket 

gopher numbers on ungrazed areas. However, other 

comparisons involving grazing intensity have shown 

heavily grazed range to have higher gopher densities 

than lightly grazed range.

Although these results seem contradictory they may 

have a biological basis. During the summer grazing 

season, both sheep and cattle consume large quantities 

of forbs, which are also the preferred summer foods 

of northern pocket gophers. Forb availability may be 

highest for gophers on ungrazed range versus lightly 

or moderately grazed range and may allow for higher 

pocket gopher densities on the ungrazed areas. On the 

other hand, higher pocket gopher densities on heavily grazed range may be an effect of long-term heavy 

grazing, which can promote greater abundance of forb 

species than moderate or light grazing (Fagerstone 

and Ramey 1996).

Pocket Gopher Management—Range management can favor plains pocket gophers, which are 

attracted to areas of improving and good range condition, where gophers use vigorous plants. Once present, 

pocket gophers interact with grasslands and livestock 

in ways that can decrease grassland productivity by 

25 to 50 percent. Managers should be aware that the 

presence of pocket gophers may require reduction in 

levels of livestock grazing to maintain good range 

condition (Fagerstone and Ramey 1996).

Even though grassland production is lowered by 

gophers, many researchers believe that gophers are 

not a significant problem on well-managed grasslands. 

In some areas, such as high mountain grassland, the 

total ecological effects of pocket gopher populations may 

be beneficial rather than detrimental. In areas where 

livestock were excluded, grass biomass increased most 

at sites having pocket gophers; so it is possible that 

pocket gophers may actually improve depleted range 

(Fagerstone and Ramey 1996).

Pocket gopher control is rarely practiced on Western 

grasslands. Gopher control is more frequently recommended for improving deteriorated grasslands than 

for maintaining grasslands that are well managed and 

productive. Where range conditions are poor, it may 

be advantageous to reduce pocket gopher populations. 

The most widely used approach to control pocket gophers is poisoning. Control of forbs, which frequently 

have large underground storage structures, can be an 

effective method for minimizing damage to grassland 

by northern pocket gophers. Application of herbicides 

can indirectly reduce pocket gopher populations by 

80 to 90 percent. Herbicide success is attributed to 

decreased forb production and resulting starvation of 

pocket gophers. Where vegetative composition after 

herbicide treatment remained relatively stable, with 

a grass dominance for 5 years, repopulation of pocket 

gophers in treated areas was slow (Fagerstone and 

Ramey 1996).

Small Rodents_____________________

Deer Mice

The deer mouse (Peromyscus maniculatus) belongs 

to the family Muridae and subfamily Sigmodontinae. 

The deer mouse is found almost everywhere in North 

America (fig. 3-3). Because it occurs over a large 

geographic area and range of habitats and is highly 

variable in appearance, more than 100 subspecies 

have been described (Mammals 2005). The deer mouse 

inhabits woodlands but it also turns up in desert areas 

(CDC 2004).

Like other small rodents, deer mice are heavily 

preyed upon and are quite secretive (Deer Mouse 

2005). Because of their abundance, deer mice are a 

major food source for almost every bird and mammal 

predator. When predators are reduced or absent, the 

mice can become pests (Cato 2005).

They are primarily nocturnal emerging from their 

nest to feed (Deer Mouse 2005). They are energy efficient, reducing their body temperature when in their 

burrows. Lowering their metabolism means they need 

less food. Deer mice do not hibernate during the winter 

(Cato 2005). Deer mice usually make their nest in a 

cavity found inside a tree, stump, under logs, and sometimes even in abandoned squirrel nests. Their nests are 

often found under rocks, boards, and haystacks. The 

nest, about the size of two cupped hands, is made of 
46 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

coarse outer materials with a soft inner lining of plant 

fibers, fur, feathers, moss, shredded tree bark, leaves, 

and other material (Deer Mouse 2005, Forest Preserve 

District of Cook County 1974).

Deer mice are widespread and adaptable, with 

broad diets (Fagerstone and Ramey 1996). The food 

of the deer mouse is mostly seeds of grasses, weeds, 

and berries. Their diet may also include buds, insects, 

spiders, centipedes, land snails, and many other 

foods (Forest Preserve District of Cook County 1974, 

Mammals 2005). They will carry food in their cheek 

pouches and often store by the quart (0.9 l). These small 

rodents need more food in proportion to their weight 

than do larger warm-blooded animals. A 1 oz (28 grams) 

mouse will eat 0.5 oz (14 grams) of food per day (Forest 

Preserve District of Cook County 1974). Food selection 

is dependent on both habitat and season. Deer mice 

feed heavily on larvae from lepidopterans (moths and 

butterflies) and other insects in the spring. They can 

eat large volumes and are capable of ridding an area of 

many insects that may be detrimental to trees. In the 

fall, seeds become a major food source and are stored 

in caches for use during the winter (Cato 2005).

Their habitat selection ranges from native prairie to 

farm fields (Fagerstone and Ramey 1996). Throughout 

their range, they occur in nearly all ecological communities and life zones from the desert floor to the high 

mountains (Cato 2005). These small rodents are pioneer 

species that occur in most vegetation types during most 

stages of plant succession, but usually not in large 

numbers (Fagerstone and Ramey 1996). At times, they 

can be highly abundant, numbering as many 

as 10 per acre (4 per ha) (Cato 2005). They 

are sometimes referred to as a “weed” species 

because disturbances that result in early seral 

stages favor population increases. Deer mice 

usually are the most abundant small mammals 

in severely disturbed areas. Therefore, grazing is generally beneficial to deer mice, which 

select areas with low cover and are common 

in habitats with bare soil surface and open 

vegetation such as grazed prairie. Researchers 

have found that heavy grazing in big sagebrush 

habitat promotes an increase in deer mice 

numbers (Fagerstone and Ramey 1996). The 

total small mammal population declined in 

grazed communities, but the density estimates 

of deer mice increased; they were dominant 

after the grazing season whereas they were 

found in only minor proportions before. Deer 

mice have used microhabitats with high shrub 

density, which is sometimes a consequence of 

grazing (Fagerstone and Ramey 1996).

Range depletion does not always favor an increase in 

deer mouse populations in all habitats. Fagerstone and 

Ramey (1996) found increased deer mice populations 

with increasing forb cover. In some studies, more deer 

mice were found in ungrazed than in grazed riparian 

habitat, but fewer deer mice were found in ungrazed 

than in grazed short-grass prairie uplands (Fagerstone 

and Ramey 1996). From the different studies conducted 

on deer mice and grazing, it can be concluded that deer 

mice have differential responses to grazing, decreasing in the most xeric habitats and increasing in mesic 

habitats (Fagerstone and Ramey 1996).

Deer mice do not normally have noticable effects 

on grassland vegetation. However, seed predation 

by mice may be an important factor in grasslands. 

During reseeding efforts, deer mice may consume or 

cache considerable quantities of seeds, resulting in 

poor plant establishment. In contrast, Fagerstone and 

Ramey (1996) concluded that seed caches may result in 

clumps of seedlings, and that 50 percent of bitterbrush 

resulted from rodent seed caches.

Deer mice may be carriers of Hantavirus. When 

present, this virus is spread through the rodent’s urine 

and feces. Although the mice do not become ill from 

the virus, humans can become infected when they 

are exposed to contaminated dust from the nests or 

droppings. Humans are advised to not camp nor sleep 

where mouse droppings are abundant and to clean 

indoor areas where mice live (Cato 2005).

People should eliminate or minimize contact 

with rodents in homes, workplaces, or campsites. 

If structures used by humans are inhospitable for 

mice, then humans will have less contact with mice. 

Recommendations include (1) Seal up holes and gaps 

in homes or garages. (2) Place traps in and around 

Figure 3-3. Distribution of the deer mouse (Peromyscus maniculatus) (NatureServe 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 47

homes to decrease rodent infestation. (3) Clean up 

any easy-to-get food (CDC 2004).

Recent research results show that many people 

who became ill with HPS developed the disease after 

having been in frequent contact with rodents and/or 

their droppings around a home or a workplace. On 

the other hand, many people who became ill reported 

that they had not seen rodents or rodent droppings at 

all. Therefore, if you live in an area where the carrier 

rodents are known to live, try to keep your home, vacation place, workplace, or campsite clean (CDC 2004). 

Never vacuum or sweep mouse droppings; thoroughly 

wet the area with a disinfectant, then carefully wipe 

up the droppings with a wet cloth (Mammals 2005).

Voles

Voles (Microtus spp.), which are commonly known 

as meadow or field mice, belong to the rodent family 

Muridae. The range for each species is limited by specific habitat conditions. Four species of voles inhabit 

grasslands in Arizona and New Mexico: long-tailed 

vole (Microtus longicaudus), montane vole (Microtus 

montanus), prairie vole (Microtus ochrogaster), and 

meadow vole (Microtus pennsylvanicus) (fig. 3-4). The 

most widely distributed species is the meadow vole. 

Voles normally occupy areas with dense ground cover, 

Figure 3-4. Distribution of long-tailed (Microtus longicaudus), 

montane (M. montanus), prairie (M. ochrogaster), and meadow 

(M. pennsylvanicus) voles (NatureServe 2005).

grasses, grasslike plants, or litter, and they 

are active day and night, year-round (Andelt 

and Ahmed 2004, Saimon and Gorenzel 2002). 

The meadow vole will inhabit stream and lake 

shores and is a good swimmer (Meadow Vole 

2005).

Vole numbers fluctuate from year to year. 

Populations are influenced by dispersal, food 

quality, climate, predation, physiological stress, 

and genetics. Under favorable conditions their 

populations can increase rapidly (Andelt and 

Ahmed 2004, Saimon and Gorenzel 2002). In 

some areas their numbers are cyclical, reaching 

peak numbers every 3 to 6 years before dropping 

back to low levels. When populations go through 

a high, numbers can soar to several thousand 

per acre (Saimon and Gorenzel 2002). During 

cyclic population peaks, voles (M. longicaudus, 

M. montanus, and M. pennsylvanicus) can reach 

densities as high as 7,400 voles per ha (2960 

per ac) (Fagerstone and Ramey 1996).

Voles make nests in clumps of grass, using 

materials such as dry grass, sedges, and weeds. 

From their nests, they build “runways”, akin to 

tunnels beneath the grass and plants (Meadow 

Vole 2005). They construct many surface 

runways and underground tunnels with numerous 

burrow entrances. The surface or subsurface burrows 

and tunnels are 1 to 2 inches (2.5 to 5 cm) wide, in a 

relatively small area, and contain numerous adults and 

young. Home-range size is usually less than 0.25 acre 

(0.1 ha) and varies with habitat quality, food supply, 

population levels, and season (Andelt and Ahmed 2004, 

Bryan 2005, Saimon and Gorenzel 2002).

Voles are primarily herbivores and forage on 

grasses, flowers, vegetables, fruits, bulbs, and roots; 

on occasion they will eat insects and snails (Saimon 

and Gorenzel 2002). They eat virtually constantly, 

concentrating on green vegetation during the summer 

and switching to mostly grains and seeds in the fall 

(Meadow Vole 2005). During winter months, voles do 

not hibernate, but instead make tunnels beneath the 

snow and gnaw on shrubs and tree bark for nutrition 

(Andelt and Ahmed 2004, Meadow Vole 2005, Saimon 

and Gorenzel 2002).

Voles are an important part of the ecosystem. They 

are preyed upon by hawks, owls, foxes, cats, snakes, 

crows, herons, shrews, skunks, bullfrogs, snapping 

turtles, largemouth bass, and raccoons (Meadow Vole 

2005).

Most vole species select sites with relatively high 

cover, so increased canopy cover is likely to increase 

Microtus populations. They do well in ungrazed or 

only lightly grazed grasslands but disappear from 

areas with moderate to heavy grazing (Fagerstone and 

Ramey 1996). In a study reported by Fagerstone and 

Ramey 1996, prairie voles were captured only on mixed 
48 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

grass prairie sites and did not occur on shortgrass sites 

such as prairie dog colonies. Mountain vole (Microtus 

montanus) was drastically reduced in numbers or 

disappeared from grazed habitats (Fagerstone and 

Ramey 1996). One hypothesis is that there may be a 

cover threshold required by voles before populations 

show significant fluctuations, and that a lower cover 

threshold may be needed before voles can establish 

resident breeding populations (Fagerstone and Ramey 

1996). Researchers have postulated that cover provides 

favorable conditions for population buildups by providing food, reducing antagonistic contacts between 

voles, and moderating microhabitat humidity and 

temperature (Fagerstone and Ramey 1996).

During normal years, voles have little influence on 

grasslands, although they may have a direct impact 

on soil. Voles at a density of 200 to 400 per ha (600 

to 900 per ac) probably dislodge 1,000 m2 of earth per 

hectare (1196 yard2

 ac) per year. Because this activity 

is restricted to the top 40 cm (15 in) of soil, such activities have minimal influence on microtopography and 

surface water runoff (Fagerstone and Ramey 1996). 

They normally have little effect on vegetation cover 

because the amount of standing crop vegetation they 

remove is usually quite small (Fagerstone and Ramey 

1996). However, after a literature review and after 

studying the California vole (Microtus californicus) 

in a field, Fagerstone and Ramey (1996) concluded 

that during the late increase and peak phases of a 

population cycle, grazing by voles can have a marked 

effect on vegetation cover. Grazing by microtine rodents 

removed current-season stem primordia of perennial 

grasses. In a series of exclosure experiments, grazing 

by voles kept the habitat open and increased plant 

species diversity; when voles were excluded, grasses, 

their preferred food, increased and became dominant. 

In some instances, voles can have severe effects on 

vegetation. Studies have found that a population of 

California voles that exceeded 1,500 voles per ha (3700 

per ac) removed 85 percent of the volume of vegetation 

for wild oats (Avenafatua), Italian ryegrass (Lalium 

mulhflorum), and ripgut grass (Bromus rigidus). Heavy 

cropping by small rodents of plants during reproduction 

suppressed flowering and caused a 70 percent seed 

loss (Fagerstone and Ramey 1996). Seed predation by 

mice, including voles, may be an important regulating 

factor for some plant species (Janzen 1971).

During high population levels, voles can kill and 

damage sagebrush and other shrub species (Andelt 

and Ahmed 2004, Bryan 2005, Fagerstone and 

Ramey 1996, Mueggler 1967, Saimon and Gorenzel 

2002) by stripping bark from plants and girdling 

stems and branches. Damage is greatest when a 

dense, ungrazed herbaceous understory exists that 

favors increases in vole populations, and when the 

snowpack persists throughout the winter (Fagerstone 

and Ramey 1996, Mueggler 1967). Usually voles kill 

only portions of the crowns of individual plants, but 

occasionally they kill entire plants. During 1962 to 

1964, Mueggler (1967) observed an irruption of voles 

on southwestern Montana grassland that caused damage to a number of shrubs, including big and silver 

sagebrush (Artemisia spp.), sumac (Rhus trilobata), 

bitterbrush (Purshia tridentata), mountain mahogany, 

and serviceberry (Amelanchier spp.). Mueggler (1967) 

recorded crown kills of 35 to 97 percent of sagebrush 

on extensive areas. A similar population explosion of 

long-tailed voles (M. longicaudus) in Utah in 1969 

killed 59 percent of sagebrush plants and damaged 

another 28 percent (Fagerstone and Ramey 1996). 

But as natives of the sagebrush-grass ecosystem, 

vole populations at normal levels have little impact 

on grassland function.

Kangaroo Rats and Pocket Mice

The Heteromyidae is a family of rodents consisting 

of kangaroo rats and pocket mice. Despite their names, 

they are neither rats nor mice; and in spite of their 

mouselike appearance, they are not closely related to 

any other species of North American rodent (Pocket 

Mice 2005). Most Heteromyidae live in complex burrows 

within the deserts and grasslands of western North 

America, though species within the Heteromys and 

Liomys genera are also found in forests and extend 

south as far as northern South America. They feed 

mostly on seeds and other plant parts, which they 

carry in their cheek pouches to their burrows. Although 

they are different in physical appearance, the closest 

relatives of the Heteromyidae are pocket gophers in 

the family Geomyidae (Heteromyidae 2005).

Kangaroo rats and pocket mice are all nocturnal, 

burrowing animals with external fur-lined cheek 

pouches for storing and transporting the seeds that 

are their primary food. They are all well adapted to 

living in arid environments and most of them never 

need to drink water. They also have efficient kidneys 

that can conserve fluids by concentrating the urine 

(Heteromyidae 2000).

Because there are many of these little rodents and 

they are closely related to each other, each species 

has evolved to have different foraging times and 

places, which minimizes competition. Bailey’s pocket 

mouse (Chaetodipus baileyi), for example, climbs up 

into desert wash vegetation to find seeds and berries 

still on the plants, while the desert pocket mouse 

(Chaetodipus penicillatus) hunts along the ground in 

washes and open areas for seeds. Merriam’s kangaroo 

rat (Dipodomys merriami), a creature of open, creosote 

flats, tends to dash from one clump of bushes to the 

next, overlooking seeds out in the open spaces, leaving 

those for other mice to find. In this way many species of 

heteromyid mice and rats can share the same environment (Heteromyidae 2000, Burgess 1996).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 49

Pocket Mice—Pocket mice are well adapted 

to arid desert life (fig. 3-5). They seldom drink, 

and can conserve water in a number of ways. 

They spend the days underground in the burrow where in summer the humidity is higher 

and the temperature lower than aboveground. 

The entrance hole is usually plugged to keep 

the moisture from escaping to the dry air 

above. Their kidneys concentrate the urine 

to a viscous consistency, reducing water loss. 

When temperatures become extreme, some 

pocket mice go into a torpor state. They appear 

to be active through most of the year in the 

southern part of their range, but they probably “hibernate,” or are at least holed up, in 

winter in northern Texas (Pocket Mice 2005). 

These animals are solitary and defend small 

territories, often fighting when they encounter 

each other (Heteromyidae, 2000).

They burrow in friable soil. Their holes have 

been described as resembling auger holes bored 

straight into the ground. Usually all the dirt 

excavated from the burrow system is piled near 

one opening. A burrow excavated in Brazos 

County, Texas, had two openings, neither of 

which was plugged, connected by a single tunnel that descended to a depth of about 40 cm 

(16 in). A side branch contained food and nest 

chambers. Another burrow was found opening under 

a log that served as a roof for the nest chamber. These 

mice have been known to inhabit deserted burrows, 

and in Texas they were using burrows of Mexican 

ground squirrels (Spermophilus mexicanus) (Pocket 

Mice 2005).

The nest of the Hispid pocket mouse (Chaetodipus 

hispidus) is composed of shredded dry grasses and 

weeds. In captivity, the mice pile the nesting material 

into a loose heap and then mat it down by sleeping on 

top of the structure. They seem to behave likewise in 

the wild (Pocket Mice 2005).

Although they feed almost entirely on vegetation, 

and principally seeds, gaillardia, cactus, evening 

primrose, and winecup are found in their caches, an 

additional 23 other species of plants have been utilized 

(Pocket Mice 2005). Animal matter makes up only 

a small part of their diet, including grasshoppers, 

caterpillars, and beetles (Pocket Mice 2005).

In farming areas, pocket mice can do considerable 

damage by digging up and carrying away planted 

seeds. In range and pasture lands they perform a 

service by eating seeds of weeds (Pocket Mice 2005). 

Some pocket mice species (Chaetodipus spp. and 

Perognathus spp.) prefer a heavy protective cover of 

grass and some shrubs (table 3-2) (Fagerstone and 

Ramey 1996). In desert grasslands, some favor cover 

forage under and around large shrubs and clumped 

vegetation (Fagerstone and Ramey 1996). Within these 

habitats, pocket mice densities are highest when soils 

are sandy, which allows for easier digging (Pocket Mice 

2005, Fagerstone and Ramey 1996). Pocket mice are 

most common in nongrazed dense grass communities or 

areas that are lightly grazed with heavy cover and high 

seed production. Pocket mice populations are reduced 

in habitats with sustained heavy grazing (Fagerstone 

and Ramey 1996). Grazing itself, by reducing the height 

of bunchgrasses, produces a less favorable habitat, and 

lowers pocket mouse numbers.

Kangaroo Rats—There are 23 species of kangaroo 

rats (genus Dipodomys) in North America (fig. 3-6); 14 

occur in the lower 48 States (Howard 1994). Kangaroo 

rats inhabit semiarid and arid regions throughout 

most of the Western and Plains States. The Ord’s 

kangaroo rat is the most common and widespread of 

the kangaroo rats (Howard 1994). Throughout the 

Southwest, kangaroo rats occur in great numbers in 

desert shrub-grasslands. Several species of kangaroo 

rat inhabit Arizona and New Mexico, including the 

banner-tailed kangaroo rat (Dipodomys spectabilis), 

Merriam kangaroo rat, and the Ord’s kangaroo rat (D. 

ordii), which has the widest range (Kangaroo Rat 2005a, 

Figure 3-5. Distribution of Bailey's (Chaetodipus baileyi), 

desert (C. penicillatus), and Hispid (C. hispidus) pocket mice 

(NatureServe 2005).
50 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Sevilleta LTER: Data 1998d). Merriam’s kangaroo rat 

inhabits warm deserts and grasslands in southern New 

Mexico where it is associated with mesquite or other 

leguminous shrubs (Sevilleta LTER: Data 1998d). In 

those areas of southern New Mexico where creosote 

bush dominates large areas of degraded grassland, 

Merriam’s rat is one of the few mammals that occur 

with any regularity. The Ord’s kangaroo rat has a 

wide distribution that includes revegetated habitats 

with low production and is most abundant where 

disturbed earth provides easy tunneling (Fagerstone 

and Ramey 1996).

Most kangaroo rats prefer areas with less dense 

herbaceous vegetation and soils that are sandy or 

sandy loam, which are easy for burrowing. Where Ord’s 

kangaroo and Merriam’s rats inhabit the same area, 

the latter are usually excluded from the more friable 

soils and are most common on desert pavements or 

other hardened, stonier soils. A common relationship 

is for Merriam’s to occupy the gravelly bajadas and 

for Ord’s to be found in the loose soils along arroyos 

or around wind-ablated playas (Sevilleta LTER: Data 

1998d). Merriam’s kangaroo rat will burrow on sandy 

soils, clays, gravels, and even among rocks, mostly in 

low deserts with scattered vegetation (Fagerstone and 

Ramey 1996, Kangaroo Rat 2005a,c, Sevilleta LTER: 

Data 1998d).

Banner-tailed kangaroo rat (Dipodomys spectabilis) 

inhabits well-developed grasslands, preferring heavier 

soils, and avoids basins where basal cover of grass is 

low. Light soils may be unable to support the fairly 

complex and deep burrow systems usually constructed 

by these rats (Sevilleta LTER: Data 1998d).

Kangaroo rat burrows can be simple in design, 

shallow, and with openings near the base of shrubs, 

or quite extensive with separate living, nesting, and 

food storage areas. Nests are constructed of plant 

fibers. Rats live in these during the daytime and rear 

their families. Usually, only one adult occupies each 

burrow system. If burrows are occupied, entrance 

holes are plugged (Kangaroo Rat 2005c, Howard 1994). 

Tracy and Walsberg (2002) determined: (1) burrows 

are much hotter during the summer than previously 

thought, 30 to 35 o

C (86 to 95 oF), (2) kangaroo rats 

remain in shallow burrows that are less than a meter 

(3 feet) below the surface at relatively high ambient 

temperatures (above 35 °C, 95 o

F) throughout the 

daytime in summer instead of residing deep within 

the soil as once assumed, (3) they do not restrict their 

activity to the coolest periods of the night but are active 

immediately following sundown, during the hottest 

time of the night, (4) burrows are not persistently 

humid but can be quite dry, and (5) insects and succulent vegetation constitute a significant portion of a 

kangaroo rat’s diet and may be key to their survival 

in the hot desert environment.

These nocturnal rodents are solitary rather than 

communal, with a home range of less than 0.5 acre (0.2 

ha); the female’s home territory is usually smaller than 

the male’s (Kangaroo Rat 2005a,b). Kangaroo rats do 

not have large home ranges, 0.04 to 0.07 acres (0.02 

to 0.03 ha) and rarely exceeding 0.14 acres (0.06 ha). 

Table 3-2. Selected pocket mice species, their response to grazing, and brief description of their habitats (Fagerstone and Ramey 

1996).

Species Response to grazing Habitat description

Great Basin pocket mice Positive Reside only in relatively dense cover in sagebrush or 

Perognathus parvus greasewood communities.

Price pocket mice Negative Are most abundant in dense stands of perennial grasses.

Perognathus penicillatus Their numbers are reduced on grazed ranges.

and 

Bailey pocket mice 

P. baileyi

Arizona pocket mouse Positive Associated with open habitat and with increased grazing.

P. amplus

Silky pocket mice Both negative and positive They were found in greatest densities where there was the

P. flavus sparsest grass cover. Adversely affected by protection

of the playa grassland from grazing.

Hispid pocket mice Positive Are commonly found in areas with open vegetation or

Chaetodipus hispidus hispidus where the prairies are intensely grazed and erosion has 

removed much of the topsoil. Hispid pocket mice inhabit a 

wide variety of habitats, from native prairie to cropfields, 

and would be expected to be affected by disturbance. 

Severe disturbance limits their populations except 

in heavily grazed prairie dog colonies.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 51

found in their cheek pouches, as well as green 

vegetation, succulents, and insects (Howard 

1994, Kangaroo Rat 2005a,b,c, Sevilleta 

LTER: Data 1998d, Tracy and Walsberg 

2002). A study of Merriam’s kangaroo rats in 

the Guadalupe Mountains showed that seeds 

made up 64 percent of the diet, with seeds of 

shrubs constituting 23 percent, those of forbs 

24 percent, those of grasses 4.5 percent, and 

those of succulent plants 12 percent.

Kangaroo rats are opportunistic feeders. 

Green vegetation is most important in midsummer, while insects are eaten in greatest 

abundance in winter (Kangaroo Rat 2005c). In 

the spring when annuals are producing seeds 

their pouches are filled with these seeds. In 

late summer, seeds of the perennial grasses 

are more abundant in their pouches. At some 

places and at some times, insects may make up 

as much as 15 to 20 percent of the diet. Green 

vegetation is sometimes consumed, especially 

during the breeding season (Sevilleta LTER: 

Data 1998d).

During winter, Merriam’s kangaroo rats 

open surface caches where seeds are stored. 

Seeds may be cached as far as 200 feet from 

where they were found. Apparently Merriam’s 

kangaroo rats do not normally store food within a burrow system but rather in separate caches on the surface 

(Sevilleta LTER: Data 1998d, Preston and Jacobs 2001). 

Kangaroo rats may harvest more than 75 percent of an 

entire seed crop (Parmenter and Van Devender 1995). 

In some years, Merriam kangaroo rats are sufficiently 

abundant to eat nearly all large perennial grass seed 

produced (Fagerstone and Ramey 1996).

Because grass seeds contain little water, adaptations 

have developed to conserve what little water rats take 

in. Their skin has no sweat glands and their urine is 

about three times more concentrated than humans 

(Kangaroo Rat 2005c). Even with their ability to conserve water, they still need to obtain water by eating 

insects and green plants (Tracy and Walsberg 2002).

Kangaroo rats are prey for owls, coyotes, foxes, 

badgers and snakes. Badgers and coyotes dig them out 

of the ground, and snakes may get some by entering 

their burrows (Kangaroo Rat 2005b,c).

 Kangaroo rats are a keystone guild: through seed 

predation and soil disturbance they have major effects 

on biological diversity and biogeochemical processes, 

facilitating the establishment of annuals and shrubs 

by selectively foraging on large seeds, and by seed 

caching and burrowing activities(Sevilleta LTER: 

Research 2001). Fields and others (1999) showed that 

banner-tailed kangaroo rats have important effects on 

both species dominance and composition of different 

vegetation patch types and may provide a mechanism 

for small-scale dominance patterns at an ecotone. This 

Figure 3-6. Distribution of banner-tailed (Dipodomys spectabilis), Merriam's (D. merriami), and Ord's (D. ordii) kangaroo 

rats (NatureServe 2005).

They may move nearly 1 mile (1.6 km) to establish a 

new home range (Howard 1994).

Kangaroo rats are generally solitary animals, although they often occur in aggregations that appear 

to have little if any social organization among them. 

Burrows are spaced to allow for adequate food sources 

within normal travel distances. Spacing of mounds will 

vary according to abundance of food, but well-defined 

travel lanes have been observed between neighboring 

mounds. Both the number of burrows and individuals 

per acre can vary greatly depending on locality and 

time of year. There are usually many more burrow 

openings than there are rats. Each active burrow 

system, however, will contain at least one adult rat. 

There could be as many as 35 rats per acre (14 per 

ha) in farmlands. In rangelands, 10 to 12 rats per 

acre (4 to 5 per ha) is more common (Howard 1994). 

These territorial rats will engage in fierce battles if a 

prowler is caught trying to pilfer from his neighbor’s 

stores (Kangaroo Rat 2005a).

Their food, which is held within two cheek pouches, 

is almost entirely seeds. Seeds of mesquite, fescue 

grass, shadscale (Atriplex confertifolia), creosote bush 

(Larrea tridentate, Cryptantha angustifolia), purslane, 

ocotillo (Fouquieria splendens), desert scrubs, Russian 

thistle (Salsola spp.), and grama grass have been 
52 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

provids further support for their role as keystone species in desert grasslands.

Other Small Mammals______________

Black-Footed Ferret

The black-footed ferret (Mustela nigripes) is the 

most endangered mammal in North America and one of 

only three species of ferrets in the world (Black-Footed 

Ferret 1995). Even before their numbers declined, 

black-footed ferrets were rarely seen: they were not 

officially recognized as a species by scientists until 

1851, following publication of a book by naturalist 

John James Audubon and Rev. John Baclunan. Even 

then, their existence was questioned because no other 

black-footed ferrets were reported for more than 20 

years (Black-Footed Ferret 2005c, CBSG 2004).

Black-footed ferret is a member of the mustelid 

(musk-producing animals) family that includes mink, 

skunks, badgers, martens, fishers, stoats, polecats, 

and wolverines (Ferret Facts 2005). They are loners, 

except during breeding season, and are nocturnal 

predators, living in or near prairie dog colonies. They 

use prairie dog burrows for shelter and travel (BlackFooted Ferret 1997, 2005c,b, Fagerstone and Ramey 

1996) and will move into vacant burrows or prey on 

the current resident and then move in (Black-Footed 

Ferret 1997). Prairie dogs make up the main staple 

of the ferret’s diet although they occasionally eat mice 

and other small animals (Black-Footed Ferret 2005a,b, 

Fagerstone and Ramey 1996). They have also been seen 

chasing birds and catching moths (Black-Footed Ferret 

2005a). A single family of four black-footed ferrets eats 

about 700 prairie dogs each year and cannot survive 

without access to large colonies of them (Black-Footed 

Ferret 2005c).

In the wild, black-footed ferrets spend 99 percent of 

their time underground (Black-Footed Ferret 2005b). 

During the night they hunt for prairie dogs that are 

sleeping in their burrows. Sometimes the ferret is the 

casualty when a group of prairie dogs attack and drag 

a ferret underground (Black-Footed Ferret 2005b). 

They will travel, from burrow to burrow, hunting for 

prairie dogs, sometimes traveling more than a mile 

in a night. They leave scent to mark their territory, 

which averages 150 acres (61 ha) for a female with a 

litter (Black-Footed Ferret 1997). The ferret is well 

adapted to slither around its prairie environment. Its 

color and markings blend so well in grassland soils and 

plants that it is hard to detect until it moves (Ferret 

Facts 2005).

Black-footed ferrets can be detected by looking for 

snow tracks or trenching. Because of their short legs, 

as ferrets dig, they cannot throw dirt between their legs 

like dogs do when they dig. They hold the dirt against 

their chests when they are digging, then back out of 

the hole, leaving a furrow of dirt. These trenches are 

usually made during winter, perhaps when ferrets dig 

after hibernating (Black-Footed Ferret 1997).

Ferrets are born in May or June, usually in a litter 

of three or four kits. Kits look like mice when they 

are born and their eyes are shut. When they are 6 to 

8 weeks old, the mother starts taking the kits out of 

their burrow. Before winter, the kits are on their own, 

and they leave their home territory and their mother 

(Black-Footed Ferret 1997, 2004).

The original distributional range of the black-footed 

ferret corresponded closely to that of prairie dogs, and 

historically they were found throughout the Eastern 

and Southern Rockies and the Great Plains, from 

Southern Saskatchewan to Texas (Fagerstone and 

Ramey 1996, Black-Footed Ferret 1994, 2004, 2005b, 

Naylor 1994). This range included portions of 12 States 

(Naylor 1994).

The black-footed ferret is an important member of 

grassland ecosystems. As a predator, they kept prairie 

dog populations in check. As with all native species, 

it evolved having a unique niche within grassland 

ecosystems. The demise of the ferret and other prairie 

species is a reminder that the grassland ecosystem 

itself may be threatened (Naylor 1994).

The decline in black-footed ferret numbers is linked 

primarily to (1) reductions in prairie dog populations 

and to (2) secondary poisoning by eating poisoned 

animals—both programs to eradicate prairie dogs 

(Black-Footed Ferret 2004, CBSG 2004, Fagerstone 

and Ramey 1996). They are susceptible to the Sylvatic 

plague and canine distemper (Black-Footed Ferret 

1995, 2004, CBSG 2004). They also are a casualty 

to plowing and fragmenting of the grasslands for 

agriculture (Black-Footed Ferret 2004, CBSG 2004). 

They are reported to be prey for owls and coyotes, and 

as with other wildlife, they become victims to vehicles 

(Black-Footed Ferret 2004).

Captive Breeding and Recovery Program—The 

black-footed ferret became extinct in the wild in 

Canada in 1937 (Black-Footed Ferret 2005c). In the 

United States, it was listed in 1967 as an endangered 

species. By the 1970s, only a few ferrets were known to 

exist, and by 1980 the species was feared to be extinct 

(Naylor 1994).

Then in 1981, a small population was discovered in 

Meeteetse, WY (Black-Footed Ferret 2005b,c, CBSG 

2004, Naylor 1994). A black-footed ferret captive 

breeding program was initiated in October 1985 by 

the Wyoming Fish Department, in cooperation with 

the USDI Fish and Wildlife Service, and a year later a 

recovery plan for captive breeding and reintroduction of 

black-footed ferrets was formed. Six black-footed ferrets 

were captured near Meeteetse to start the program. 

The ferrets were taken to the Department’s Sybille 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 53

Wildlife Research Center near Wheatland, WY, and 

eventually all died of canine distemper (Black-Footed 

Ferret 1995, 1997, CBSG 2004). The disease was then 

confirmed among Meeteetse’s wild ferrets (CBSG 2004, 

Saving a Species 2005).

Biologists launched an emergency effort to capture 

all remaining animals. Five were captured in late 1985, 

12 more in 1986, and by February 1987, the last known 

wild black-footed ferret was captured. All the animals 

were vaccinated and quarantined, and all 18 survived 

(CBSG 2004, Saving a Species 2005).

No kits were born during the ferret breeding season 

in captivity in 1986. However, eight were born in 1987 

at Sybille. Seven survived and were followed by 34 

surviving kits in 1989 and 66 in 1990 (CBSG 2004, 

Saving a Species 2005).

In 1988, the USDI Fish and Wildlife Service developed the “Black-Footed Ferret Recovery Team” that 

emphasized species preservation through natural 

breeding, development reproductive technology, and 

establishment of multiple reintroduction sites. The 

objective of the captive breeding program was to maintain 240 ferrets (90 males, 150 females) of 1 through 

3 years old and subdivide the captive population into 

different groups to avoid catastrophic loss at a single 

facility. The strategy for the reproductive program 

was to support captive breeding efforts by developing 

artificial insemination. One high priority for protecting 

genetic diversity was to have a frozen repository of 

sperm from genetically valuable males (CBSG 2004, 

Saving a Species 2005).

After evaluating eight Gunnison’s prairie dog complexes across northern Arizona, the Aubrey Valley was 

selected as the best site for black-footed ferret reintroduction. In 1997, the prairie dog acreage estimate was 

29,653 acres. With the release of 35 ferrets (nine kits, 

26 adults) in 1996, Aubrey Valley in Arizona became 

the fourth reintroduction site and the first to develop 

and evaluate onsite acclimation pens to precondition 

release candidates. No ferrets were released in 1997, 

26 in 1998, 52 in 1999, 19 in 2000, 12 in 2001, and six 

in 2002. Survivorship has been generally low. In 2001, 

the first wild-born black-footed ferret kits were found 

in Arizona following a spring release of animals bred 

prior to release (CBSG 2004).

Current Status—Captive-bred ferrets have been 

reintroduced to the Shirley Basin in Wyoming, UL 

Bend National Wildlife Refuge and the Fort Belknap 

Reservation in Montana, the Badlands National Park 

and Buffalo Gap National Grasslands in South Dakota, 

and Aubrey Valley in Arizona (Black-Footed Ferret 

2005b,c, CBSG 2004). Approximately 1,000 blackfooted ferrets live in captivity at breeding facilities, 

while another 80 exist in the wild following release by 

the Federal government (Black-Footed Ferret 2005b, 

Naylor 1994).

Despite these population gains in some areas of the 

country, both habitat loss and the continued decline of 

their prey base, the prairie dog, continue to threaten 

the black-footed ferret (Black-Footed Ferret 2005b). 

Conversion of grasslands to agricultural uses, widespread prairie dog eradication programs, and plague 

(Black-Footed Ferret 1995) have reduced ferret habitat 

to less than 2 percent of what once existed. Remaining 

habitat is now fragmented, with prairie dog towns 

separated by great expanses of cropland and human 

development (Ferret Facts 2005). Preservation of large 

prairie dog colonies will be essential for recovery of 

the black-footed ferret.

Rabbits and Hares

Approximately 50 species of rabbits and hares

form the family Leporidae of the order Lagomorpha 

(Lagomorphs 2005b). This group of mammals is largely 

diurnal or crepuscular (Order Lagomorpha 2005). The 

order Lagomorpha, with a fossil history dating back 

to the Oligocene (33.8 to 23.7 million years ago), comprises two modern families: Ochotonidae (pikas) and 

Leporidae (hares and rabbits). Members of all genera 

except Lepus are usually referred to as rabbits, while 

members of Lepus are usually called hares (Fagerstone 

and Ramey 1996). Rabbits—as distinguished from the 

related hares—are altricial, having young that are 

born blind and hairless; many also live underground 

in burrows (Lagomorphs 2005b). No lagomorphs 

hibernate (Order Lagomorpha 2005). The distinction 

between these two common names does not correspond 

completely with current taxonomy, because the jackrabbits are members of Lepus, whereas members of 

the genera Pronolagus and Caprolagus are sometimes 

called hares (Lagomorphs 2005b).

Rabbits and hares feed almost entirely on vegetable 

matter—grasses, forbs, bark of trees and shrubs, and 

so forth. Because of their usually large size and food 

preference, lagomorphs frequently come into conflict 

with grazing, agriculture, and forestry interests (Order 

Lagomorpha 2005). Leporids are conspicuous mammals 

in various habitats on grasslands. These herbivores 

generally occur in areas where short grasses and herbs 

are abundant and clumps of tall grasses or brush are 

available for cover. Several taxa of lagomorphs are 

major components of the wildlife of grasslands. The 

mountain cottontail (Sylvilagus nuttallii) occurs on 

montane “islands”; the desert cottontail (S. audubonii) 

is an inhabitant of upland, grazed areas; the eastern 

cottontail (S. floridanus) seems restricted mostly to 

ungrazed, riparian-edge habitats. The white-sided 

jackrabbit (Lepus callotis) occurs in Southwestern 

desert grasslands; however, its current range in the 

United States apparently is restricted to about 120 

km2 (46 ac) in southwestern New Mexico (Jones and 
54 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Manning 1996). The grassland habitats for some of 

these species have expanded as a result of alterations 

of the environment by humans.

Rabbits are prolific breeders and some, especially 

black-tailed jackrabbit (L. californicus), may become 

abundant enough that at times they contribute to 

overgrazing of grasslands and even become a nuisance 

to agricultural crops (Jones and Manning 1996). 

Rabbits are major dispersers of seeds of some important 

plants, especially dropseed (Sporobolus spp.), and they 

are important components of the food chain in that 

they serve as major prey species for some carnivores. 

Lagomorphs are even of some value to humans as food 

(Fagerstone and Ramey 1996).

Rabbits—Seven genera in the family are classified 

as rabbits, including the European rabbit (Oryctolagus 

cuniculus) and cottontail rabbits (Sylvilagus spp.) 

(Lagomorphs 2005b). They are well-known for digging 

networks of burrows, called warrens, where they spend 

most of their time when not feeding. Rabbits can be 

gregarious, while hares are often solitary. In areas 

with high densities of rabbits, females form dominance 

hierarchies, with the dominant females suppressing the 

reproduction of subdominants by denying them access 

to nest sites and by physical intimidation (Lagomorphs 

2005a). Cottontail females (doe) and young share territory only until the young are independent (Eastern 

Cottontail 2005).

Rabbits are also well-known for their breeding rate, 

another factor that differentiates them from hares; in 

theory, a doe can produce from three to seven live young 

per month, during the first half of the year, although a 

more common rate is half that (Lagomorphs 2005b). In 

the warmer parts of their range, eastern cottontails can 

breed from the months of February through September, 

with three or four litters per year. The female rabbit 

is entirely responsible for her young. The doe makes a 

shallow nest that she lines with grasses, twigs, and fur 

she pulls from her own coat. She visits her nest only 

at dusk and dawn (Desert Cottontail 2005, Eastern 

Cottontail 2005).

Cottontails—Two primary species of cottontail 

rabbits are of interest to Southwestern grasslands: 

the eastern cottontail (S. floridanus) and the desert 

cottontail (S. auduboni). Cottontails are among the 

most widely distributed of North American mammals, 

with eastern cottontail the most widely distributed 

of the cottontail rabbits. Eastern cottontails are the 

most common rabbits in North America. The eastern 

cottontail is found from the Eastern Seaboard west 

to the Rocky Mountains and from southern Canada 

south to Costa Rica (fig. 3-7) (Eastern Cottontail 2005, 

Fagerstone and Ramey 1996). It occurs throughout 

the Plains region primarily in riparian ecosystems 

(Fagerstone and Ramey 1996) and ubiquitously in the 

Eastern deciduous forests.

Desert cottontails are distributed widely throughout the arid areas of Western North America, from 

Montana south to central Mexico, and from the High 

Plains of Oklahoma, Kansas, and Nebraska to the 

Pacific Coast (fig. 3-7) (Fagerstone and Ramey 1996). 

They occur in a wide variety of habitats, including dry 

desert grasslands and shrublands, riparian areas, and 

pinyon-juniper forests. They may occur in the same 

areas as black-tailed jackrabbits (Lepus californicus) 

(Desert Cottontail 2005).

Cottontails are preyed upon by a number of predators, including golden and bald eagles, great horned 

owls, ferruginous hawks, badgers, coyotes, foxes, bobcats, and humans. Badgers, weasels, and rattlesnakes 

may prey on the young (Desert Cottontail 2005).

Cottontails can be found almost anywhere, fields, 

woods, and farmlands but they especially inhabit areas 

where there are thickets and brush for shelter and for 

hiding (Eastern Cottontail 2005). Typical habitat in 

the Great Plains includes weedy margins of fields and 

pastures, brushy areas, and dry ravines.

Cottontail densities are positively correlated with 

increased biomass of herbaceous vegetation and 

with areas ungrazed by livestock (Fagerstone and 

Ramey 1996). Periodic cycles may have occurred in 

historic populations. For example, eastern cottontail 

population densities have been reported to be as high 

as 17 to 25 per hectare (42 to 62 per ac) with 8 to10 

years in cyclic tendencies (Fagerstone and Ramey 

1996, Lagomorphs 2005a). However, recent land use 

changes have more profound impacts on population 

densities than any natural processes. For one thing, 

nesting density is habitat specific. A Pennsylvania 

study illustrated how human changes in the landscape 

can alter population densities: nests occurred every 

1.5 acres (0.6 ha) in unkempt orchards, were 7 acres 

(2.8 ha) apart in hayfields, 13.5 acres (5.5 ha) apart 

in woodlands, and 14 acres (5.7 ha) in pasture lands 

(Lagomorphs 2005a).

Home range for eastern cottontails of 2.5 to 5 acres 

(1 to 2 ha) is normal, with a range from 1 to10 acres 

(0.4 to 4 ha). The home range will depend on populations, habitat quality, season, and sex, with the male 

having a slightly larger range. Ranges for males are 

the largest in the main breeding period of late spring to 

early summer, while females have the smallest range 

at this same time. Dominant males have the largest 

ranges. Ranges are generally smaller in spring (before 

breeding) and in winter, reflecting lush vegetation 

and severe limiting weather. These mammals are not 

territorial, and their ranges often overlap (up to 50 

percent for males and 25 percent for females in spring), 

especially in winter, when they tend to concentrate in 

areas offering the best combination of food and cover. 

Females have little or no overlap of home ranges during 

breeding season (Lagomorphs 2005a).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 55

Eastern cottontails, like most other rabbits, are 

solitary, generally not territorial, except for females 

in the immediate vicinity of a nest. Males have a 

dominance hierarchy in which the most dominant have 

more aggressive encounters with other males and do 

most of the mating. Dominance hierarchy of males 

allows the strongest males to fertilize more females 

than subordinates and also minimizes fighting. Most 

aggressive behavior is exhibited between the dominant 

male and the individual immediately below it in social 

status. Females have a less rigid hierarchy. Females 

exhibit dominance over males except during estrus 

(Lagomorphs 2005a).

Rabbits are herbivores, eating a wide range of vegetation. The diet of the eastern cottontail varies with 

habitat and seasonal availability of forage. Eastern cottontails prefer herbaceous plants when available during 

the growing season, including bluegrass, orchard grass 

(Dactylis glomerata), timothy (Phleum pratense), clover 

(Trifolium spp.), and alfalfa (Eastern Cottontail 2005, 

Fagerstone and Ramey 1996). Succulent new growth 

supplies much of the cottontail’s water requirements. 

Woody species are preferentially eaten during the 

dormant season (Fagerstone and Ramey 1996).

The desert cottontail is a crepuscular to nocturnal 

forager. Desert cottontails can survive droughts by 

obtaining water from cacti and forbs (Desert 

Cottontail 2005, Fagerstone and Ramey 1996). 

They also can survive drought conditions 

because they are coprophagic, meaning they 

eat their own feces. Because grass is difficult 

to digest, the rabbits eat the first-formed set 

of pellets after a meal. These pellets, rich 

in protein and B vitamins, are reingested 

directly from the anus. Additional nutrition is 

extracted during the second digestive process. 

Pellets from the second set are hard, fibrous, 

and lack nutritive value (Desert Cottontail 

2005, Lagomorphs 2005a). This practice allows 

the animals to spend relatively little time 

exposed to predators because while they are in 

the field feeding, they consume green vegetation rapidly and then make optimum use of 

it in the safety of their cover. This process is 

also called “pseudo-rumination” because it is 

functionally the same as cows chewing their 

cud. Coprophagy is also practiced by beaver 

and voles and, apparently, by some shrews.

Desert cottontail rely seasonally on grasses, 

sedges, rushes, shrubs (for example, blackberry [Rubis allegheniensis), and trees (for 

example, willow [Salix spp.] and oak [Quercus spp.]). 

Their annual diet is similar to that of prairie dogs 

and cattle in the kinds of plants eaten, but differs in 

the relative proportions preferred (Fagerstone and 

Ramey 1996).

Cottontail Economic Status—Eastern cottontails 

are the most widely hunted game mammals in the 

United States (Jones and others 1985), and their high 

reproductive rates allow them to withstand high hunting pressure. Eastern cottontails are responsible for 55 

percent of tularemia cases reported in Americans, due 

to direct contact while skinning and dressing animals. 

A few cases of plague (Pasteurella pestis) have also 

been reported in cottontail rabbits. Cottontails are 

not an important contributor to grassland overgrazing. Overall, the ecological, economic and recreational 

benefits from hunting by humans outweigh the minor 

damage cottontails do to crops, nurseries, and orchards 

(Fagerstone and Ramey 1996).

As with other rabbits, cottontails disperse seeds 

widely owing to their high abundance, intensive use of 

small annual and perennial herbs, and production of 

fecal pellets. Seed dispersal by rabbits may influence 

the distribution and long-term dynamics of some plant 

species (Fagerstone and Ramey 1996).

Cottontail Association with Livestock—Cottontails 

were reported to be significantly greater on the ungrazed bottomlands paralleling the South Platte River 

in eastern Colorado and were almost nonexistent on 

grazed areas. The desert cottontail is negatively impacted when pinyon-juniper habitat is cleared during 

Figure 3-7. Distribution of the eastern cottontail (Sylvilagus 

floridanus) and the desert cottontail (S. auduboni) (NatureServe 2005).
56 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

operations for increased livestock production 

(Fagerstone and Ramey 1996).

Cottontail Management—Habitat management to increase populations should emphasize 

moderate grazing, clumps of shrubs and 

small trees, and possibly rock and brush 

piles (Fagerstone and Ramey 1996). Because 

cottontails utilize successional vegetation 

primarily, habitat management techniques 

such as prescribed burning and sharecropping 

may be useful for controlling succession and 

increasing cottontail numbers.

When cottontails are found in high concentrations, they may damage crops, nurseries, 

and orchards. Controls include hunting and exclusionary methods such as tree trunk guards, 

fencing, repellents, habitat modifications, and 

trapping (Fagerstone and Ramey 1996).

Jackrabbits—Jackrabbits are a prominent 

grassland herbivore throughout the West. 

Two principal species occur on grassland: the 

black-tailed jackrabbit and the white-tailed 

jackrabbit (L. townsoni) (fig. 3-8). Jackrabbit 

densities are dependent on vegetation, climate, 

season, and other factors (Fagerstone and 

Ramey 1996). Reported densities have ranged from 

0.01 per ha (0.025 per ac) in southeastern Colorado to 

35 per ha (88 per ac) in agricultural areas in Kansas. 

Jackrabbit densities are significantly higher near 

cultivated crops than on isolated grassland (Fagerstone 

and Ramey 1996).

Fluctuations in jackrabbit density have been reported in the literature with cycles of approximately 

5 to 10 years. Populations in local areas can become 

extremely large during population irruptions. Some 

researchers believe that the populations are not actually cyclic, but that drought and food availability or 

drought and overgrazing concentrate the jackrabbits. 

Evidence now suggests that the key parameters associated with population fluctuations are much more 

complex than previously thought. There appear to be 

geographic trends in jackrabbit frequency of fluctuations, and these include the interactions between many 

features of the jackrabbit habitat, for example, food 

availability, and natural phenomena such as weather 

(Fagerstone and Ramey 1996).

The black-tailed jackrabbit is the most common 

jackrabbit in the Western and Central United States 

(fig. 3-8), ranging from the Pacific Coast to western 

Missouri and Arkansas, and from the prairie and 

grassland regions of Idaho to South Dakota to the 

Mexican border (Dunn and others 1982, Fagerstone 

and Ramey 1996). Black-tailed jackrabbits occur in 

many diverse habitats but are primarily associated 

with shortgrass prairie and open country. Black-tailed 

jackrabbits avoid areas of heavy brush or woods, 

which limit their principal means of defense— keen 

eyesight and escape speed (Fagerstone and Ramey 

1996, Mueggler 1967).

The white-tailed jackrabbit occurs in close association with the flora of the northern Great Plains 

and open areas of the Great Basin (Dunn and others 

1982, Fagerstone and Ramey 1996). They range from 

Southern Canada to Colorado and from Michigan to 

the high mountain slopes of the Rockies, Cascades, 

and Sierras (fig. 3-8). White-tailed jackrabbits once 

ranged south across the Plains States to southern 

Kansas, but they now occur generally north of the 

Platte River (Fagerstone and Ramey 1996) and are 

considered imperiled in New Mexico.

White-tailed jackrabbits prefer large expanses of 

croplands and pastures interspersed with brush and 

heavy vegetation in open flats. In recent years, the 

white-tailed jackrabbit’s range has declined, and it is 

now extirpated from Kansas and is rare in Missouri. 

These jackrabbits are not adapting to the general 

climatic warming of the Great Plains and are less 

able to use cultivated areas than are the black-tailed 

jackrabbits. Some researchers have theorized that 

the expansion of black-tailed jackrabbits into areas 

formerly occupied by white-tailed jackrabbits was due 

Figure 3-8. Distribution of the black-tailed jackrabbit (Lepus 

californicus), white-tailed jackrabbit (L. townsoni), and whitesided jackrabbit (L. callotis) (NatureServe 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 57

to the changing habitat associated with prairie cultivation by humans. When both species came into contact on 

the shortgrass prairie, white-tailed jackrabbits selected 

more sparsely vegetated upland habitats and occupied 

higher elevations than black-tailed jackrabbits. It has 

been suggested that the black-tailed jackrabbit is more 

efficient than the white-tailed jackrabbit in foraging. 

Thus, the black-tailed jackrabbit, with its greater 

adaptability and feeding efficiency, may be able to 

displace the white-tailed jackrabbit (Fagerstone and 

Ramey 1996).

Jackrabbits feed in open areas that have a supply 

of succulent young plants, interspersed with patches 

of dense weeds; the open areas allow them to detect 

danger, and the dense weeds serve as cover. Foraging 

begins during twilight, increases during early night, 

and begins to decrease as dawn approaches. This general pattern of foraging is influenced by season, ambient 

temperature, and phase of the moon (Fagerstone and 

Ramey 1996).

Although the plant species eaten vary throughout 

the jackrabbit’s range, their diets have some seasonal 

features in common: they show greater use of shrubs 

in winter than in other seasons and greater use of 

grasses and forbs in the spring and summer. Various 

cereal crops and other cultivated crops (such as winter 

wheat [Triticum aestivum], alfalfa, crested wheatgrass 

[Agropyron cristatum]) are used when available. Plant 

phenology was a major factor in determining food 

preferences of grassland jackrabbits; 85 percent of their 

diet was composed of grasses in the spring, grasses and 

forbs were nearly equal in diets in early summer, and 

forbs and shrubs increased to 71 percent of the diets in 

late summer. Grasses and sedges composed 49 percent 

of their overall diet (Fagerstone and Ramey 1996).

In contrast, in arid areas of the Southwest, mesquite 

made up 54 percent of the annual diet. Needleand-thread grass (Stipa comata) was preferred in 

sagebrush habitat, while yarrow (Achillea lanulosa) 

was preferred in bitterbrush habitat. Grasses are the 

largest component of jackrabbit diets in semidesert 

grassland, particularly during the summer growing 

season. Herbaceous weeds were not preferred dietary 

items. Most investigators agree that jackrabbits select 

plants for their succulence, particularly during times 

of water stress (Fagerstone and Ramey 1996).

Positive and Negative Effects of Jackrabbits—Blacktailed jackrabbits have both positive and negative 

relationships with humans. The following are the 

positive effects (Fagerstone and Ramey 1996):

• They are used for sport hunting, food, and fur.

• They are also an important part of the prairie 

ecosystem and a major food for mammalian 

predators, particularly coyotes.

• They are important to secondary succession on 

old fields and denuded ranges by dispersing seeds 

in fecal pellets.

• Jackrabbits are also important in increasing the 

viability of some seeds in their pellets by their 

digestive processes.

The following are the negative effects (Fagerstone 

and Ramey 1996):

• In some areas they cause extensive damage to 

agriculture. Cultivated crops adjacent to grassland are particularly vulnerable to damage by 

jackrabbits given the grassland provides daytime 

resting areas for them.

• Jackrabbits have been reported to cause damage to seedling trees, grains, and cotton, range 

rehabilitation efforts, and vegetables.

White-tailed jackrabbits are generally viewed as 

having positive effects. Some of the positive benefits 

of these jackrabbits have been associated with sport 

hunting, food, and fur. Their fur has had some commercial value. They are also important for maintenance 

of ecological balance, biodiversity, and aesthetics of 

grassland ecosystems.

Jackrabbit Association with Livestock—The relationship between livestock grazing and jackrabbit 

population numbers is not well understood, and there 

are mixed opinions as to the costs and benefits of cohabitation of jackrabbits and livestock. The varied 

conclusions noted in table 3-3 may be the result of 

differences in study areas, grazing systems, or other 

factors. Studies have shown the interactive relationship 

of herbivores and their pastures, and that herbivores 

affect the biomass, growth, and species composition 

of the pasture (Fagerstone and Ramey 1996). As the 

density of herbivores increases, a point is reached 

when plant species of low quality and palatability 

also increase and make the pasture less suitable for 

herbivores, including jackrabbits. The influences that 

range conditions have on the diets of jackrabbits and 

cattle (Lagomorphs 2005b) are: jackrabbit densities 

were highest on ranges that were in good condition 

(Eastern Cottontail 2005); range condition influenced 

the amount of forbs and shrubs in a jackrabbit’s diet, 

but not the grass component; and jackrabbits consumed 

less grasses and more shrubs than cattle, producing a 

moderate dietary overlap (Order Lagomorpha 2005). 

In early studies of competition, scientists estimated 

the foraging capacity of black-tailed jackrabbits to be 

one cow equivalent to 148 black-tailed jackrabbits or 

62 white-tailed jackrabbits, and one sheep equivalent 

to 6 black-tailed or 15 white-tailed jackrabbits. The 

greatest direct competition for forage between cattle 

and black-tailed jackrabbits was in early spring when 

both species preferred green forage such as western 

wheatgrass, needle-and-thread, and sunsedge (Carex 
58 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

heliophila), and least in late fall and winter (Fagerstone 

and Ramey 1996).

The vegetation on various grasslands has changed 

during the past century, partly because of poor grazing 

practices (Fagerstone and Ramey 1996). Reports have 

documented shifts from grassland to shrub-dominated 

vegetation in New Mexico. In Great Basin shrub communities protected for 15 years from domestic sheep, 

with or without protection from jackrabbits, plant 

community changes were slow to nonexistent, and 

protection from jackrabbits had no effect (Burgess 

1996, Fagerstone and Ramey 1996).

Bats

The more than 900 species of bats worldwide belong to the Order Chiroptera. Bats are found almost 

everywhere on Earth, except in extremely hot desert 

environments and the cold Polar Regions. The United 

States is known to have 15 genera, totaling 44 species 

of bats. Those that inhabit the Southwestern deserts 

comprise 11 genera and more than 18 species. The 

diverse habitats of New Mexico and Arizona support 

up to 28 species of bats (Parmenter and Van Devender 

1995, Chung-McCoubrey 1996), many of which are 

found regularly or occasionally in grasslands.

Bats are unique in the animal kingdom because they 

are the only mammals to have evolved true flight. Bats 

are often thought of as flying mice, but they are more 

closely related to primates than to mice. As with most 

other mammals, their bodies are covered by hair, with 

the exception of their wings. Although bats have the 

same basic arm and hand bones found in humans and 

most other mammals, their hand and finger bones are 

long and slender, and there are only four digits. The 

skin between the arms, fingers, body, legs, and feet 

looks delicate, but it is extremely resistant to tearing 

by sharp objects.

The bats in temperate North America are nocturnal and are mostly insectivorous, but a few are 

nectarivorous and play an important role in the pollination of certain plants (Chung-MacCoubrey 1996). 

As insectivores, bats are important because they feed 

on nocturnal flying and terrestrial insects, and they 

likely play a role in regulating insect populations 

(Chung-McCoubrey 1996) and insect-related ecological 

processes. Bats help maintain balances of relationships within the insect community, and between 

insects and plants, animals, and other entities. Bat 

droppings (guano) support entire ecosystems of unique 

organisms, including bacteria useful in detoxifying 

wastes, improving detergents and producing gasohol 

and antibiotics. Bats are integral to the function and 

integrity of many ecosystems (Chung-McCoubrey 

1996).

Mating may occur two or even three times a year, 

in late fall, just before hibernation, in midwinter if 

the roost is warm enough, and again in spring. Birth 

takes place in the spring or summer after a gestation period of 50 to 60 days in May, June, and July. 

Within a week after its birth, the female will carry 

her pup on nightly hunts; the pup grasps her fur and 

feeds at one of her two nipples. The young bats are 

weaned in 2 to 3 weeks, and then they may be fed 

on regurgitated food brought home by the mother. 

In 3 to 4 weeks, the young bat is hunting on its own 

and is only 5 weeks away from full growth. Females 

typically mate at the end of their first summer, males 

at the end of their second. Multiple births up to four 

occur in some species.

Many people see bats as a threat because they fear 

bats carry rabies. Like any other animal, a bat that 

contracts rabies will die. What is unique about bats 

with rabies is that they rarely become aggressive. 

Dogs and cats pose a far more dangerous threat as 

transmitters of rabies to humans. Humans are rarely 

endangered by bats except in cases where sick bats 

are handled.

Bat Habitat—Grassland habitats for bats are 

limited by the availability of roosting sites, such as, 

crevices, caves, trees, buildings, mines, bridges, other 

artificial structures, and by availability of water. The 

type of roosts selected varies by bat species. Their 

mobility allows them to select habitats at a large landscape level and to utilize habitats that are separated 

by significant distances (Chung-McCoubrey 1996). 

Their small size allows them to exploit a large variety 

of sheltered sites.

Roosts: The annual energy budget of bats makes 

it difficult for them to balance body size, energetic 

demands of flight, fat storage given limited ability, and 

response to seasonal abundance of prey. One of the keys 

to managing their energy expenditures is selection 

Table 3-3. Some results of studies on the relationship between 

jackrabbits and livestock (Fagerstone and Ramey 1996).

Relationship Explanation

Negative The combination of 

livestock and jackrabbits 

can cause severe destruction to grassland.

Positive with heavy grazing Some researchers believe 

heavy grazing of grassland 

increases suitable habitat 

for jackrabbits and encourages their presence.

Positive with moderate grazing It has been reported that 

jackrabbits prefer moderately grazed pastures.

Positive with no grazing It has been stated that 

jackrabbits prefer nongrazed grassland.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 59

of roost sites (Chung-McCoubrey 1996). Roost sites 

typically have the following characteristics:

• Suitable microclimates that facilitate gestation 

and rapid growth of the young (Chung-McCoubrey 

1996).

• Exposed to minimal disturbance.

• Relatively close to food and water.

Reproductive success and over-winter survival of 

individuals and populations may largely depend on 

suitable roosts (Chung-McCoubrey 1996). Winter hibernation sites minimize the potential for disturbance 

and arousal and maximize use of energy reserves. 

Other factors, such as threat by predators, distance to 

or availability of local hibernacula, and sensitivity to 

human disturbance may influence local distributions 

and sites

Bat roosts within grasslands include crevices in and 

under stones and rocks, excavated or natural holes 

in the ground, and the foliage of scattered shrubs 

and trees. Such habitats are interspersed within 

grasslands, patches of other habitat types such as 

rock escarpments, talus slopes, cliff faces, lava flows 

and tubes, caves, open mines, barns, and bridges, 

providing a host of different roost environments for 

grassland bats. In addition, bats may roost within 

the foliage, bark, and cavities of riparian vegetation along arroyos, tributaries, and rivers that pass 

through grasslands. Table 3-4 provides a summary of 

grassland bat species (including scientific names of 

bats) by Federal status and types of summer roosts 

(Chung-MacCoubrey 1996).

Water: Surface water for drinking is critical for bats. 

Due to their high protein diet, insectivorous bats require 

water to excrete toxic nitrogenous waste products. 

Bats residing in dry environments have high rates of 

evaporative water loss through wing membranes and 

respiratory exchange. Xeric species such as California 

myotis (Myotis californicus), western pipistrelle 

(Pipistrellus hesperus), pallid bat (Antrozous pallidus), 

and Mexican free-tailed bat (Tadarida brasiliensis) 

have high urine-concentrating abilities and are efficient 

at conserving water. Species that live in more mesic 

grasslands and deserts (for example, long-legged myotis 

(Myotis volans), fringed myotis (Myotis thysanodes), 

little brown myotis (Myotis lucifugus), Yuma myotis 

(Myotis yumanensis), and Townsend’s big-eared bat 

(Corynorhinus townsendii)) have low urine-concentrating abilities and probably select habitats where water 

is readily available (Chung-McCoubrey 1996).

Historically, bats relied on naturally occurring 

water sources, sparsely distributed springs, seeps, 

and permanent water sources. Water availability may 

have limited the geographic distribution of species not 

adapted for water conservation or for long-distance 

flight. Construction of waterholes and stock tanks in 

Southwestern grasslands has increased the quantity 

and distribution of available water. Areas originally 

devoid of water may have become viable roosting and 

foraging habitat for bat species that were historically 

absent (Chung-McCoubrey 1996). Because few records 

document bat distributions prior to Euro-American 

development of the Southwest, it is impossible to 

confirm whether geographic ranges of such species—for example, fringed myotis and long-legged 

myotis—expanded into grasslands and deserts due to 

the increased number of water holes and stock tanks, 

or whether they were always present.

Food: Food along with water availability determine 

bat species distribution and habitat use. Insects may 

generally be so abundant as to preclude competition 

among bats (Fagerstone and Ramey 1996). Dietary partitioning may occur among insectivorous bats as evident 

from their wide range of sizes, flight styles, echolocating 

abilities, and the vertical and horizontal partitioning 

of space during foraging (Chung-McCoubrey 1996). 

Although bats forage on diverse insects, some select 

particular orders of insects. In New Mexico, California 

myotis, western pipistrelles, and long-legged myotis 

are classified as Lepidoptera (moths) strategists. 

Pallid bats, long-eared myotis, and fringed myotis are 

classified as Coleoptera (beetles) strategists (ChungMcCoubrey 1996). These species consume arthropods 

in addition to moths and beetles, including Orthoptera 

(grasshoppers), Hymenoptera (bees/wasps), Diptera 

(flies), Homoptera (leafhoppers), Hemiptera (true 

bugs), and Isoptera (termites) (Chung-McCoubrey 

1996, Fagerstone and Ramey 1996). Diet composition 

reflects seasonal peaks of different arthropod species 

(Chung-McCoubrey 1996) and probably varies with 

habitat type.

Conservation—Bats have many natural enemies, 

and consequently large numbers of bats die while still 

young. Predators include great horned owls, some 

species of hawks, peregrine falcons, raccoons, house 

cats, and snakes. Bats can also be caught on barbedwire fences, fall from roosts, or die if their roost site 

(for example, cave) is flooded. Some 40 percent of 

the bat species in the United States and Canada are 

endangered or candidates for such status. Even small 

disturbances in their habitats can seriously threaten 

bat survival. Agricultural insecticides are responsible 

for killing bats in great numbers. When bats consume 

chemical-laden insects, the bats become poisoned and 

die (Chung-MacCoubrey 1996).

Desert Grasslands__________________

Desert grassland ecosystems impose many restrictions on its residents. Precipitation is scarce 

and unpredictable, and temperatures and wind 

velocities fluctuate greatly. Cover from weather and 

predators is scarce. Wildlife species in the harsh desert 
60 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

environment have adapted morphologically, physiologically, and behaviorally to survive drought, heat, 

and cold.

A well-known and studied area in Socorro County 

in south-central New Mexico is the Sevilleta LongTerm Ecological Research Site (LTER) managed by 

the University of New Mexico. The LTER has a wide 

variety of vegetation types including desert grassland, 

Chihuahuan and Great Basin desert scrub, pinyon-juniper woodland, wetlands, riparian woodland, montane 

forest, and meadows. Scientists at the Sevilleta LTER 

have described the number of mammal species by 

habitat type (table 3-5). A total of 56 mammal species 

have been observed in desert grasslands at the Sevilleta 

LTER and 51 species in desert-scrub. Fewer mammal 

species occupy other habitats at the LTER, suggesting 

that desert grasslands and scrub contribute greatly to 

the mammalian species diversity of the Southwest.

Effects of Grazing on Rodents, and 

Vice Versa

Desert shrub-grasslands in the Southwest are 

primarily used by humans for grazing cattle and 

Table 3-4. Federal status and types of summer roosts by bats in Southwestern grasslands and short-grass prairies (Chung-McCoubery 1996).

Species USFWS Status Type of summer roosts

Species more commonly associated with grasslands

Small-footed myotis (Myotis ciliolabrum) Species of concern Cracks and crevices of cliffs and rocks, 

 abandoned buildings and barns; under rock 

 slabs and loose bark; possibly in caves and 

 mine tunnels.

California myotis (M. californicus) Cliffs, hillsides, rock outcrops, mine shafts, 

 barns, houses, under tree bark and sign 

 boards, amongst desert shrubs, and on the 

 ground.

Cave myotis (M. velifer) Species of concern Primarily caves and tunnels; occasionally 

 buildings, bridges, and under rocks.

Pallid bat (Antrozous pallidus) Rocky outcrops, crevices, mine tunnels, 

 buildings, and under rocks.

Western pipistrelle (Pipistrellus hesperus) Canyon walls, cliffs, and other rock crevices; 

 under rocks, in burrows and buildings.

Mexican free-tailed bat (Tadarida brasiliensis) Caves, mines, bridges; occasionally in 

 buildings.

Species found in grasslands given appropriate habitat

Little brown bat (M. lucifugus) Buildings, hollow trees, natural crevices, mines.

Yuma myotis (M. yumanensis) Species of concern Crevices, mines, caves, buildings.

Fringed myotis (M. thysanodes) Species of concern Caves, mine tunnels, rock crevices, old 

 buildings.

Long-legged myotis (M. volans) Species of concern Abandoned buildings, cracks in ground, cliff 

 face, and other crevices, under loose bark.

Long-eared myotis (M. evotis) Species of concern Tree hollows, loose bark, folds of wood/bark, 

 rock crevices, abandoned buildings, mines.

Hoary bat (Lasiurus cinereus) Foliage of trees and shrubs.

SiIver-haired bat (Lasionycteris noctivagans) Hollow trees, woodpecker holes, under loose 

 bark, and in buildings.

Eastern red bat (Lasiurus borealis) Foliage of trees and shrubs, clump of Spanish 

 moss.

Big brown bat (Eptesicus fuscus) Hollow trees, rock crevices, mine tunnels, 

 caves, buildings: occasionally in cliff swallow 

 nests.

Townsend’s big-eared bat (Plecotus townsendii) Species of concern Caves, mine tunnels, and abandoned buildings.

Spotted bat (Fuderma maculatum) Species of concern Cracks and crevices in rocky cliffs or under 

 loose rocks.

Big free-tailed bat (Nyctinomops macrotis) Species of concern Crevices in rocky cliffs, buildings.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 61

secondarily for farming. These desert grasslands 

are characterized by seasonal bursts of vegetation 

productivity following sporadic rains (Fagerstone 

and Ramey 1996). Vegetation is composed of desert 

shrubs, drought resistant summer-growing perennial 

grasses, and annual plants. Seeds persist after green 

vegetation is gone.

As an example of the effect of grazing on wildlife, 

a look at kangaroo rats is helpful because kangaroo 

rats are a keystone guild. Through seed predation and 

soil disturbance they have major effects on biological 

diversity and biogeochemical processes, facilitating 

the establishment of annuals and shrubs by selectively 

foraging on large seeds, and by seed caching and burrowing activities (Sevilleta LTER: Research 2001). 

Fields and others (1999) showed that kangaroo rats 

have important effects on both species dominance and 

composition for different vegetation patch types and 

may provide a mechanism for small-scale dominance 

patterns at an ecotone.

Merriam kangaroo rats, as a further example, are 

favored by grazing. The range of this species matches 

with the distribution of creosote bush, low humidity and 

rainfall, high summer temperatures, and evaporation 

rates. They avoid sites with dense cover and prefer 

open areas with scattered woody plants and annual 

grasses. Consequently, they tend to inhabit lands that 

are managed on a sustainable basis for cattle grazing 

(Fagerstone and Ramey 1996). On rangelands, the rats 

may do some damage by consuming seeds of desirable 

grazing grasses, but in general, losses attributable 

to them are negligible (Kangaroo Rat 2005c). Sound 

management of grazing on high elevation grasslands 

allows the maintenance of perennial grass with interspersed shrubs, sustains forage for livestock and 

reduces erosion. At lower, drier elevations, however, a 

shrubby cover may be all that can be sustained. When 

grasslands are improperly grazed, there is a gradual 

downward trend in perennial grass density and a 

corresponding increase in Merriam kangaroo rats. A 

corresponding invasion of mesquite is also observed 

at some localities where grass densities are reduced, 

and this vegetation shift is notably accompanied by 

an increase in Merriam kangaroo rat populations 

(Fagerstone and Ramey 1996).

The abundance of plant species whose seeds are 

favored foods of small mammals can be strongly affected by seed predation. Foraging by rodent granivores 

substantially reduces the standing crop of large-seeded 

winter annual plants (Fagerstone and Ramey 1996) 

that other small mammals or grazing animals may 

favor. After removal of rodents, densities of these plants 

increased as much as several thousand times.

Rodents are important to plant population recruitment in desert grasslands. Although rodents consume 

large amounts of seed, their seed caches are a major 

Table 3-5. The number of mammal species found in various habitats in Socorro County, central New Mexico (Parmenter and Van 

Devender 1995).

Habitat type

Desert Montane Riparian

Family scrub grassland Pinyon-juniper forest meadow zone

Shrews (Soricidae) 2 1 2 1 2 1

Bats (Molossidae, Vespertilionidae) 9 10 7 12 12 8

Rabbits (Leporidae) 2 2 1 1 1 1

Squirrels (Sciuridae) 4 6 5 5

Gophers (Geomyidae) 4 4 1 1 1 1

Kangaroo rats, pocket mice (Heteromyidae) 5 5 3

Beaver (Castoridae) 1

Mice, rats (Arvicolidae, Muridae, Cricetidae, Zapodidae) 9 13 10 8 8 6

Porcupine (Erethizontidae) 1 1 1 1 1

Coyote, foxes (Canidae) 3 3 3 2 2 3

Bear (Ursidae) 1 1 1 1 1 1

Raccoon, ringtail (Procyonidae) 1 1 1 1 1 1

Weasels, badger, skunks (Mustelidae) 5 4 5 4 3 4

Cats (Felidae) 2 2 2 2 2 2

Deer, elk (Cervidae) 2 2 2 3 3 2

Sheep (Bovidae) 1 1 1

Pronghorn (Antilocapridae) 1 1 1

Total 51 56 46 43 38 31

Note: SEV = Sevilleta National Wildlife Refuge (mixed-grassland-shrubland-woodland)

BDA = Bosque del Apache National Wildlife Refuge (riparian and wetlands)

MAG = Magdalena Mountains (montane forest and meadows)
62 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

source of plant recruitment (Fagerstone and Ramey 

1996, Pocket Mice 2005). Kangaroo rats cache seeds 

in a centrally located burrow, but they also often store 

seed in scattered caches just below the soil surface. Seed 

caches that are not recovered provide for recruitment of 

new plants. The establishment of small-seeded plants 

and of annuals and perennial grasses is influenced by 

pocket mice (Fagerstone and Ramey 1996), whereas, 

the establishment of large-seeded plants is affected 

by kangaroo rats.

Kangaroo rats have variable effects on range condition of desert or arid grasslands. When a range is in 

good to excellent condition, Merriam kangaroo rats 

have little effect on seed dispersal (Fagerstone and 

Ramey 1996). Good quality grassland is less favorable 

habitat for kangaroo rats because its increased cover 

produces obstacles for kangaroo rats when escaping 

from predators, and because large-seeded vegetation, their preferred food, is replaced by small-seeded 

plants. On grasslands in good to excellent condition, 

the seed-burying habits of heteromyid rodents are 

probably beneficial to grassland condition, given that 

large-seeded perennial grasses and tall shrubby plants 

have been shown to increase on areas where kangaroo 

rats were most abundant (Fagerstone and Ramey 1996). 

During a favorable seed year, and when range condition 

is such that the supply of large-seeded perennial grass 

seed is in excess of the needs of kangaroo rats, much 

more seed is cached by rats than is ever recovered 

(Fagerstone and Ramey 1996). Large seeded species 

have difficulty in germinating and establishing from 

seed on the soil surface. Such species may require 

burial in seed caches for seedling establishment to 

occur (LaTourette and others 1971). Because seed 

buried in the ground is in a more favorable environment for germination and seedling survival than seed 

lying on the ground surface, the rate of plant restocking may be enhanced by the presence of kangaroo 

rats (Fagerstone and Ramey 1996). Rodent caches of 

antelope bitterbrush (Purshia tridentata), snowbrush 

(Ceanothus velutinus), squawcarpet (C. prostratus), 

green rabbitbrush (Chrysothamnus viscidifiorus), 

eheatgrass, and Indian rice grass (Oryzopsis hymenoides) have been reported (Fagerstone and Ramey 1996). 

On recently burned or denuded pinyon-juniper and 

sagebrush sites, kangaroo rat caches created opportunities for germination of species such as bitterbrush 

and snowberry (Symphoricarpos spp.). Rodents also 

transport mycorrhizae associated with range plants 

and therefore may facilitate establishment of plant 

species and their mycorrhizae on denuded range sites 

(Fagerstone and Ramey 1996).

As grassland conditions decrease from fair to poor, 

the vegetation composition changes, woody perennial shrubs increase, and perennial grasses decline 

(Fagerstone and Ramey 1996). Increasing openness 

allows kangaroo rats to see and avoid predators 

(Fagerstone and Ramey 1996). Under these conditions, 

kangaroo rats can be an important factor in accelerating range deterioration. In poor grassland, kangaroo 

rat activities may prevent range recovery.

Reduction of kangaroo rat populations to increase 

forage is justified biologically only where the density 

of perennial grass is low and can be increased by 

grazing management or range improvement practices. 

Kangaroo rat control may also be warranted where 

artificial reseeding of large-seeded plants is hampered 

by kangaroo rats (Fagerstone and Ramey 1996).

Kangaroo rats may prevent grassland succession by 

maintaining sub-climax vegetation. Long-term removal 

of a guild of kangaroo rat species from a Chihuahuan 

Desert ecosystem led to the conversion of the habitat 

from shrubland to grassland (Fagerstone and Ramey 

1996). And 12 years after removal, density of tall perennial and annual grasses increased approximately 

three times and rodent species typical of arid grassland 

colonized, including harvest mice.

Reptiles and Amphibians___________

To this point, this discussion of grassland animals 

has focused on mammals. As a group they are homoeothermic, “warm-blooded.” Their body temperature is 

maintained within a narrow range, regardless of ambient temperature. The source of body heat is metabolic, 

powered by their food they eat. The skin is of great 

importance in conserving or disposing of excess heat 

from the body. Insulating layers of hair and/or fat 

prevent heat loss in cold weather. The sweat glands, 

when present, dissipate heat by evaporative cooling. 

(Mammals 1997).

The discussion now turns to reptiles and amphibians. 

By contrast to mammals, reptiles and amphibians are 

ectotherms, “cold-blooded” animals such as reptiles, 

fish, and amphibians, whose body temperature is 

regulated by their behavior or surroundings. They must 

bask in the sun or find a warm spot to get warm and 

become active, and they must find shade or a cool spot 

to cool off. In cold conditions they become sluggish and 

do not move around much, and some enter a state of 

inactivity or hibernation if it becomes cold for a long 

time (Reptiles 2005, Amphibians 2005).

Savage (1960) described a modern North American 

desert and plains herpetofauna whose boundaries are 

determined by zones of relatively rapid species transitions into different surrounding herpetofaunas. With 

the increasingly xeric climate of the late Miocene, 

the modern, unbroken grasslands began to form by 

coalescence of previously scattered and isolated fragments (Scott 1996). In marked contrast to mammalian 

faunas that experienced massive Pleistocene (1.8 

million to about 10,000 years ago) extinctions, North 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 63

American herpetofaunas have changed remarkably 

little since the Pliocene (5.3 million to 1.8 million years 

ago). Pleistocene herpetofaunas from Western North 

America were composed of most of the same species 

that are found now (Rogers 1982, Parmley 1990). A 

major difference between early herpetofaunas and 

modern ones is the loss of several tortoises of the genus 

Geochelone (Moodie and Van Devender 1979). Based 

on fossil evidence, the Great Plains herpetofauna have 

evolved in situ since at least the Miocene (23.8 million 

to 5.3 million years ago), with only minor east-west 

and north-south shifts that coincide with Pleistocene 

glaciations. This stability has produced a recognizable 

grassland herpetofauna that is relatively uniform 

across the North American plains.

The patterns of herpetofaunal diversity were accomplished mainly by the addition or deletion of species 

from a widespread suite of grassland forms. Most species are wide-ranging, supporting the notion that the 

grassland fauna is fairly homogeneous. For example, 

half (6 of 12) of the reptile and amphibian species found 

in the grasslands of Alberta and almost three-fourths 

(32 of 43) of the Kansas tallgrass species are also found 

in the grasslands of the Chihuahuan Desert, several 

hundred kilometers to the south (Scott 1996).

Reptiles

There are four main groups of reptiles: turtles and 

tortoises; lizards and snakes; crocodiles and alligators; 

and the tuatara. Many spend their time on land, but 

some reptiles spend most of the time in water, such as 

crocodiles, alligators, turtles, some species of snakes, 

and some species of lizards. Reptile species can be 

found in all types of habitats except polar ice and 

tundra (Reptiles 2005).

Most reptiles make nests or dig holes to lay their 

eggs in. Some stay to guard the nest and even facilitate 

the hatchlings start in life. But most female reptiles 

leave the nest once eggs are laid; the hatchlings are 

independent from the start and must find their own 

food and shelter (Reptiles 2005).

Reptiles, like other vertebrates, partition habitats 

according to their food and shelter, although competition between various species may also influence 

their relative abundance. For example, in Hidalgo 

County, southwestern New Mexico, desert spiny lizard 

(Sceloporus magister), side-blotched lizard (Uta stansburiana), tree lizard (Urosaurus ornatus), and western 

whiptail lizard (Aspidoscelis tigris) prefer habitat with 

greater densities of shrubs. Other species, including 

the greater earless (Cophosaurus texanus), Longnose 

Leopard lizard (Gambelia wislizenii), round-tailed 

horned (Phrynosoma modestum), and zebra-tailed 

lizards (Callisaurus draconoides), prefer more open 

areas with few shrubs (Parmenter and Van Devender 

1995).

Habitat, food, and behavioral factors were examined 

to determine how four similar species of whiptail lizards 

could coexist in southeastern Arizona (Parmenter and 

Van Devender 1995). Each species used a slightly different part of the habitat. Little striped whiptail lizards 

(Aspidoscelis inornata) preferred areas dominated by 

mesquite. Desert grassland whiptails (Aspidoscelis 

uniparens) inhabited the ecotone between mesquite 

habitats and Arizona Upland Sonoran desertscrub. 

Two all-female (parthenogenetic) whiptail species were 

found in transition zones.

Parthenogenetic whiptail species typically occupy 

transitional ecotones between the habitats where 

their parent species occur. The desert grassland is an 

evolutionary center for all-female whiptails. Seven 

species are mostly or completely restricted to this habitat. Checkered (A. tesselata), Chihuahuan spotted (A. 

exsanguis), and New Mexican whiptails are common in 

desert grassland in Texas and New Mexico, while desert 

grassland and Sonoran whiptails are more common in 

southeastern Arizona. The Gila spotted whiptail (A. 

flagellicauda) is common in desert grassland-interior 

chaparral mosaics below the Mogollon Rim in central 

Arizona. The plateau whiptail (A. velox) lives in Great 

Basin grasslands on the Colorado Plateau above the 

Mogollon Rim (Parmenter and Van Devender 1995).

Turtles and Tortoises

The order Testudines, popularly known as the turtles, 

includes the tortoises and the terrapins. Testudines 

is an ancient clade dating back to the Triassic period, 

248 million to 206 million years ago, and today are 

represented by more than 200 species. Populations of 

many turtle species have declined, and such endangered 

species as the Green Sea Turtle (Chelonia mydas) have 

become international symbols to environmentalists 

and conservationists (Turtles 2005).

All turtles retain the basic strategy of laying eggs in 

nests, always on land, either buried in sand or hidden 

in vegetation. This tactic of depositing eggs has been a 

major factor in the endangerment of many turtles. The 

eggs are abandoned to the mercy of predators. Juvenile 

turtles often play different ecological roles than their 

larger parents, especially as prey to predators. However, 

turtles in general have relatively long life spans and 

mate repeatedly and have a generalized life history 

strategy of producing many young (Turtles 2005).

Turtles have a wide range of diets and habitats, and 

thus fill a variety of ecological roles. The armored shells 

of turtles may seem impregnable, but still the turtles 

have their predators, including predatory birds, and 

some mammals, such as coyotes. Their tough shells are 

not a suit of invulnerability, though tortoises have been 

known to survive wildfires in grasslands by withdrawing into their shells. That same behavior doubtless 

frustrates many predators (Turtles 2005).
64 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Box Turtles—All North American box turtles 

belong in the Emydidae family of turtles. This large 

family also includes the sliders, map turtles, and pond 

turtles from North America and Asia. Box turtles are 

separated from all the other turtles in this family into 

the genus Terrapene (Cook 1997). The western box 

turtle (Terrapene ornata) inhabits the grasslands of 

the Southwest. This turtle is found as far north and 

east as South Dakota, Michigan, and Indiana, south 

through southeastern Arizona, New Mexico, and Texas 

into northern Mexico (Western Box Turtle 2005a). In 

Arizona, the subspecies known as desert grassland box 

turtle (Terrapene ornata luteola) is common in certain 

areas (Parmenter and Van Devender 1995).

The western box turtle of grasslands are found in 

treeless plains to gentle hills with grass or low bushes 

and sandy soils. Their ranges may have developed along 

side the great herds of grazing animals on the North 

American Prairies (Cook 1997). They occasionally 

inhabit desert habitats (Western Box Turtle 2005a). 

This turtle tends to create shallow burrows in loose 

soils; it will also use mammal burrows and bannertail 

kangaroo rat mounds. These burrows are used to avoid 

temperature extremes and reduce desiccation (Western 

Box Turtle 2005a). Their powerful front legs and strong 

claws are perfectly made for tearing apart manure 

piles in search of dung beetles and grubs. Studies 

have shown that the Ornate box turtle’s numbers are 

reduced when cattle are removed from that turtle’s 

home range (Cook 1997).

The western box turtle is omnivorous, feeding 

on insects, especially beetles, berries, leaves, fruits, 

and sometimes carrion. It reproduces from March 

to November, laying two to eight eggs per clutch. 

Breeding strongly correlates with rainfall (Western 

Box Turtle 2005a).

Western box turtles are locally threatened by 

dangers associated with agriculture and increasing 

urbanization. Roads are major threats: hundreds 

of turtles may be killed by vehicles in a single year 

on certain Interstate Highways, and dozens may be 

run over on secondary roads. Machinery used to till 

farmland and grow crops and applications to improve 

farmlands and ranges can inadvertently injure or kill 

box turtles. For example, in Missouri, this species incurred a high rate of mortality as a result of prescribed 

burning of tallgrass prairie in late October (Western 

Box Turtle 2005b).

Box turtles are popular in the pet trade of Europe 

and Southeast Asia. Excessive exploitation for this 

trade may be a significant threat to box turtles, and 

their visibility on roads also increases their vulnerability to collectors (Western Box Turtle 2005b). They 

are listed by The Convention of International Trade in 

Endangered Species of Fauna and Flora (C.I.T.E.S.) 

as a threatened species. Permits for their export and 

import are required. Many States protect native box 

turtles and do not allow collection. These turtles are a 

long-lived species with low egg/clutch numbers, high 

hatchling mortality rates, and ever shrinking habitat. 

Their survival may depend on active conservation 

and research into their needs and demography (Cook 

1997).

Lizards

Lizards typical of desert grassland include lesser 

earless (Holbrookia maculata), side-blotched (Uta 

stansburiana), southern prairie (Sceloporus undulatus consobrinus), and one or more species of whiptail 

(Aspidoscelis spp.) lizards. In dry, gravelly arroyos, 

greater earless and zebra-tailed lizards are usually 

found. In southeastern Arizona and southwestern New 

Mexico, the Gila monster (Heloderma suspectum) is 

occasionally found in rock outcrops in desert grassland 

(Parmenter and Van Devender 1995).

The Gila monster is one of only two venomous lizards 

known. The other venomous lizard is the Mexican 

beaded lizard (Heloderma horridum) in Mexico and 

Central America. Gila monsters are not aggressive 

or dangerous unless they are picked up and handled. 

They typically inhabit the lower slopes of mountains 

and nearby outwash plains, especially in canyons and 

arroyos where water is at least periodically present. 

In some areas, they frequent irrigated farmlands 

that adjoin those habitat types. Other cover in such 

areas often includes boulders, rock crevices, downed 

vegetation, and litter. Gila monsters dig burrows for 

shelter, or use those made by other animals or formed 

by nature. These shelters are occupied both as winter 

hibernacula and as warm-season retreats from the 

heat. Gila monsters are common to rocky slopes, 

and uncommon to mesquite-dominated bajada in 

the Sulphur Springs Valley of Arizona (Gila Monster 

2005b,c, Mexican Beaded Lizard. 2005).

The diet of Gila monsters includes small mammals, 

snakes, lizards, the eggs of birds and reptiles, and 

invertebrates. They are a diurnal and occasionally 

nocturnal predator. They use their tongue to sample 

the air and substrate for molecules of substances that 

provide them information about the environment. This 

mechanism is apparently the principal method used 

to locate their prey. Coyotes, owls, hawks, and eagles 

may prey upon them, and other reptiles probably eat 

young Gila monsters as well (Gila Monster 2005b,c).

Gila monster populations have been exploited by 

commercial and private collectors, and have suffered 

from habitat destruction due to urbanization and 

agricultural development. They are often killed by 

people who believe they are dangerous and a hazard to 

the public. They are also one of the most commercially 

valuable reptile species in North America. Stringent 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 65

prohibitions against commercial exploitation and 

unnecessary killing are needed. As a result of these 

threats, there is no question that the Gila monster 

is less widespread and less abundant than it was 

formerly. Habitat preservation is important, especially 

the protection of den sites. Gila monsters are protected 

under California and Arizona law, and the species is 

listed as endangered in New Mexico (Gila Monster 

2005a,b,c).

The horned lizards or “horny toads” (Phrynosoma

spp.) are an interesting looking group of ant-eating 

reptiles. They are not like typical long slender lizards 

but are flat and chunky. Round-tailed horned lizard, 

regal horned lizard (P. solare), and Texas (P. cornutum) 

horned lizards are common in desert grasslands from 

western Texas to southeastern Arizona (Parmenter 

and Van Devender 1995).

Snakes

Common nonvenomous snakes of the desert grasslands include the gopher snake or bullsnake (Pituophis 

catenifer, and P. c. sayi), coachwhips (Masticophis 

flagellum cingulum, M. f. lineatulus, M. f. piceus, 

and M. f. testaceus), desert grassland kingsnake 

(Lampropeltis getulus splendida), Great Plains Rat 

snake (Elaphe guttata emoryi), western hognose 

snake (Heterodon nasicus), Trans-Pecos ratsnake (E. 

subocularis), and western hooknosed snake (Gyalopion 

canum). The Mexican vine snake (Oxybelis aeneus), 

green rat snake (E. triaspis), and desert hooknosed 

snake (Gyalopion quadrangulare) are tropical species 

occasionally found in desert grassland in the Atascosa 

and Santa Rita Mountains of Pima and Santa Cruz 

Counties, south-central Arizona (Parmenter and Van 

Devender 1995).

Four venomous snakes widespread in desert 

grassland are the Mohave (Crotalus scutulatus), the 

prairie (C. viridis viridis), western diamondback (C. 

atrox) rattlesnakes, and the desert grassland massasauga (Sistrusrus catenatus edwardsi). The diminutive 

Arizona coral (Micruroides euryxanthus), a member of 

the cobra family (Elapidae), reaches desert grassland 

in southeastern Arizona and southwestern New Mexico 

(Parmenter and Van Devender 1995).

Amphibians

There are about 5,500 known species of amphibians, 

divided into three main groups: slamanders, newts, 

and mudpuppies; caecilians; and frogs and toads. 

They are animals that live part of their lives in water 

and part on land. Amphibians are ectothermic and 

cannot regulate their own body heat as mammals do. 

They depend on heat from sunlight to become warm 

and active. They also cannot cool down on their own, 

so if they get too hot, they have to find shade. In cold 

weather, they tend to be sluggish and do not move 

around much (Amphibians 2005).

Young amphibians do not look like their parents. 

They are generally called larvae, and as they develop, 

they change in body shape, diet, and lifestyle, a process 

called metamorphosis. Frogs are familiar examples. 

After hatching from eggs, they start out as tadpoles 

with gills to breathe underwater and a tail to swim 

with. As they grow, they develop lungs, legs, and 

a different mouth. Their eyes also change position 

and they lose their tails. At this point they are adult 

frogs, which spend most of their time hopping on land 

(Amphibians 2005).

Most amphibians have soft, moist skin that is 

protected by a slippery secretion of mucus. They also 

tend to live in moist places or near water to keep 

their bodies from drying out. Many adult amphibians also have poison-producing glands in their skin, 

which make them taste bad to predators and might 

even poison a predator that bites or swallows them 

(Amphibians 2005).

Some form of permanent or ephemeral water must 

be present to facilitate amphibian reproduction in 

the desert grassland. Livestock water developments 

and ponds are reliable water sources that are readily 

colonized by amphibians. Summer thunderstorms 

routinely fill small playas and pools with water. When 

this happens, there is a surge in amphibian reproduction and populations (Parmenter and Van Devender 

1995).

When we think of frogs, we generally picture what are 

called “true frogs.” These amphibians are members of 

the family Ranidae, containing more than 400 species. 

Frogs from this family can be found on every continent 

except Antarctica. These frogs are characterized by (1) 

bulging eyes, (2) strong, long, webbed hind feet that are 

adapted for leaping and swimming, (3) smooth or slimy 

skin, and (4) eggs in clusters (Frogs and Toads 2005). 

The term toad refers to “true toads,” members of the 

family Bufonidae, containing more than 300 species. True toads can be found worldwide except in 

Australasia, polar regions, Madagascar, and Polynesia. 

These amphibians are characterized by (1) stubby 

bodies with short hind legs for walking instead of 

hopping like true frogs, (2) warty and dry skin, (3) 

paratoid, poison, glands behind the eyes, (4) eggs laid 

in long chains. Some toads (genera Nectophrynoides), 

however, are the only types of frogs and toads to bear 

live young (Frogs and Toads 2005).

The physical distinctions between frogs and toads 

can easily get blurred because sometimes the features 

appear mixed or less obvious, and certain species even 

legitimately fall into both categories. It is not uncommon, for example, to find a warty-skinned frog that is 

not a toad, or even a slimy toad. Even the more invisible morphological features such as cartilage structure 
66 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

has been found to sometimes fit both categories (Frogs 

and Toads 2005).

Desert grassland is not usually thought of as conducive habitat for amphibians. Certain toads, however, 

including Couch’s (Scaphiopus couchi), plains (S. 

bombifrons), and western (S. hammondi) spadefoot 

toads, and green (Bufo debilis), Great Plains (B. cognatus), and southwestern Woodhouse’s (B. woodhousei 

australis) toads, can be quite common. The true frogs, 

such as Chiricahua (Rana chiricahuensis), lowland (R. 

yavapaiensis), and plains (R. blairi) leopard frogs, and 

the introduced bullfrog (R. catesbeiana), are generally 

limited to permanent, often artificially developed water 

sources (Parmenter and Van Devender 1995).

The last group of amphibians that is addressed are 

salamanders. Salamanders are in the Order Caudata. 

These amphibians date back 150 million years ago 

to the Triassic period (Salamanders 2005b). They 

are divided into nine families, with some 400 species 

worldwide. More than half of these are found only in 

the New World, and the eastern and western regions 

of North America are centers of salamander diversity. 

Salamanders are tailed amphibians having four legs 

of more-or-less equal size. Most have vertical creases 

down their sides called costal grooves. They are easily 

distinguished from lizards by a lack of claws and scales 

(Salamanders 2005a).

There are three types of salamanders: totally aquatic, 

semiaquatic, and completely terrestrial (Kaplan 2002). 

Most salamander species are largely terrestrial as 

adults but lay their eggs in or near water. They undergo a gilled, aquatic larval stage before transforming 

into reproductive adults and dispersing to terrestrial 

habitats (Salamanders 2005a). However, some species 

retain their gills through their life. They are able to 

take up oxygen through the skin, and in addition the 

lungless salamander can also take up oxygen through 

the membrane of the mouth. They have mucus-forming 

glands that help to keep them moist, and the glands 

also expel toxic secretions when the animal fears 

danger. Whether aquatic or terrestrial, salamanders 

need moisture for survival and are found in only wet 

or damp environments (Salamanders 2005b).

The majority of salamanders and their larva are 

carnivorous, preying on insects, worms, and other 

small invertebrates. Large adults will eat fish, frogs, 

and other salamanders. Secretive, they are chiefly 

nocturnal, hiding under fallen logs and damp leaf litter 

during the daylight hours. The larvae begin feeding 

immediately after hatching, devouring tiny aquatic 

animals. Likely they perform important ecological 

roles in the communities where they live (Salamanders 

2005a,b, Kaplan 2002).

Occasionally one finds tiger salamanders (Ambystoma 

tigrinum) in desert grassland water developments 

(Parmenter and Van Devender 1995). The tiger 

salamander covers a wide range of areas extending 

nearly coast to coast in North America. There are 

several subspecies within this complex. The tiger 

salamanders are large and robust, reaching average 

total lengths up to 8.5 inches (21.6 cm), though some 

individuals over 12 inches (30.5 cm) long have been 

found. Outside of the breeding season they are seldom 

seen, as they spend most of their time underground, 

often in mammal burrows (Salamanders 2002).

Tiger salamanders living in isolated ponds may 

exhibit a condition known as neoteny, in which the 

animal becomes mature at an earlier stage in life 

than usual. The salamanders reproduce as aquatic 

larvae and may never transform into terrestrial 

adults. Neoteny is a survival mechanism some species in arid climates have evolved to assure they can 

reproduce in stressed conditions (Parmenter and Van 

Devender 1995).

Conclusions_______________________

The biology, ecology, and management of several important grassland terrestrial and aquatic vertebrates 

have been presented. Some of these species, for example 

prairie dogs and gophers, have been and are considered 

pests and nuisances by ranchers, farmers, and many 

others. Voles, kangaroo rats, and other small rodents 

are often considered as little insignificant creatures 

that are not even considered in management planning, 

but as discussed above, each of the species are essential 

to the functioning of grassland ecosystems.

The roles each of the species play within the ecosystems are numerous and diverse. All of the species in 

this chapter are prey for different predators. Prairie dog 

and pocket gopher burrows provide habitat for large 

suites of other species. Prairie dogs are required as food 

for the most endangered species in the United States, 

black-footed ferret. Not only does the digging performed 

by voles, gophers, and other mammals provide habitats 

for many other species, their activities also aerate 

and turn the soil, recycle nutrients, and in so doing, 

expose new enriched soil. The new soil creates beds 

for vegetation that differ from the vegetation in the 

surrounding areas, and this consequently contributes 

to the diversity and heterogeneity of habitats.

Additional examples of species that are major players 

in ecosystems are bats. Bats play an important role in 

pollinating certain plants, and as such they assist in 

maintaining the species that have ecological relationships with those plants they pollinate. As insectivores, 

bats likely play a role in regulating insect populations 

and insect-related ecological processes. Bat guano 

supports entire ecosystems of unique organisms.

In summary, the interconnection and integration of 

the species in this chapter—and additional species, both 

animal and plant, that were not mentioned—indicate 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 67

how crucial it is to manage grasslands and ecosystems, even including the so-called pests, nuisances, 

and obscure little insignificant creatures, in order to 

maintain healthy systems that will provide benefits 

now and in the future.

Acknowledgments_________________

I thank Mark A. Rumble, Fred B. Samson, and John 

Sidle for reviewing this chapter and providing insightful 

comments. I express my appreciation to Deborah M. 

Finch for editing the chapter. Also great gratitude goes 

to Kerri Mich and Doug Paulson for their assistance 

in creating the distribution maps.

References________________________

Amphipians. 2005. Animal Bytes: Amphibians. Available: URL 

http://sandiegozoo.org/animalbytes/a-amphibians.html

Andelt, W.F. and S. Ahmed. 2004. Managing voles in Colorado. 

Colorado State University. Available: URL http://www.ext.coloradostate.edu/pubs/natres/06507.html

Black-Footed Ferret. 1995. Feature Series. U.S. Fish and Wildlife 

Service. Vol. 1, No. 1. 2 p. Available: URL http://www.r6.fws.

gov/feature/ferrets.html

Black-Footed Ferret. 1997. Wyoming Game and Fish Dept. 

Conservation Publication. Vol.13, No. 8. 4 p. Available: URL 

http://gf.state.wy.us/services/publications/wildtimes/ferret.htm

Black-Footed Ferret. 2004. Colorado Division of Wildlife. Available: 

URL http://wildlife. state.co.us/Educationlmammalsguide/

black_footed_ferret.asp

Black-Footed Ferret. 2005a. The Mammals of Texas Available: URL 

http://gf.state.wy.us/services/publications/wildtimes/ferret.htm

Black-Footed Ferret. 2005b. National Parks Conservation 

Association. Available: URL

http://www.npca.org/wildlife_protection/wildlife_facts/ferret.asp

Black-Footed Ferret. 2005c. Wikipedia. Available: URL http://

en.wikipedia.org/wiki/Black-Footed_Ferret

Brand, C.J. 2002. Landscape Ecology of Plague in the American 

Southwest, September l9-20, 2000, Fort Collins, Colorado, 

Proceedings of an American Southwest Workshop: Fleston, 

Virginia, U.S. Geological Survey Information and Technology 

Report 2002-0001, 24 p.

Bryan, H.D. 2005. Voles. The University of Tennessee, Agricultural 

Extension Service. Available: URL http://knox.tennessee.edu/

voles.htm

Burgess, T. 1996. Desert Grasslands, mixed shrub, savanna, shrub, 

or semidesessrt scrub? The dilemma of coexisting growth forms. 

In: The desert grassland. M. P, McClaran and T. R. Van Devender 

(eds). The University of Arizona Press. Tucson, Arizona.

Cato, P. 2005. Deer mouse. San Diego Natural History Museum. 

Available: URL http://www.sdnhm.org/fieldguide/mammals/peroman.html

CBSG. 2004. Black-Footed ferret population management planning workshop. Final report. IUCN/SSC Conservation Breeding 

Specialist Group. Apple Valley, MN. 129 pp.

CDC. 2004. All about hantaviruses. Center for Disease Conrol. 

Available: URL http://www.cdc.gov/ncidod/diseases/hanta/noframes/rodents.html

Chung-McCoubrey, A.L. 1996. Grassland bats and management in 

the Southwest. In: D.M. Finch (ed). 1996 Ecosystem disturbance 

and wildlife conservation in western grasslands – A symposium 

preceedings, September 22-26, 1994; Albuquerque, NM. Gen. Tech. 

Rep. RM-GTR-285. Fort Collins, CO: USDA Forest Service, Rocky 

Mountain Forest and Range Experiment Station. 82 p.

Cook, T. 1997. The Eastern and Western Box Turtle. Available: URL 

http://boxturtlesite.org/eastern.html

Cully, J.F. and E.S. Williams. 2002. Interspecific Comparisons of 

Sylvatic Plague in Prairie Dogs. Page 6 In: Brand, C.J., 2002, 

Landscape Ecology of Plague in the American Southwest, 

September l9—20, 2000, Fort Collins, Colorado, Proceedings of an 

American Southwest Workshop: Fleston, Virginia, U.S. Geological 

Survey. Information and Technology Report 2002—0001, 24 p.

Deer Mouse. 2005. Deer mouse. Animal Fact. Available: URL 

http://wildwnc.org/af/deermouse.html

Desert Cottontail. 2005. Desert cottontail. Desert USA. Available: 

URL http://desertusa.com/mag00/apr/papr/rabbit.hrml

Desert USA. 2005. Botta’s Pocket Gopher. Available: URL http://www.

desertusa.com/mag01/jun/papr/gopher.html

Eastern Cottontail. 2005. Eastern cottontail. Available: URL http://

www.iwr-online.org/kids/Facts/Mammals/cottontail.htm

Fagerstone, K.A. and C.A. Ramey. 1996. Rodents and lagomorphs. 

In:Range Wildlife, Kausmann, P.R., (ed). The Society of Range 

Management, Denver, Colorado. 440 p.

Ferret Facts. 2005. Black-Footed Ferret Recovery Implementation 

Team. Available: URL http://www.blackfootedferret.org/facts.

html

Fields, M.J., D.P. Coffin, and J.R. Goez. 1999. Burrowing activities 

of kangaroo rats and patterns in plant species dominance at a 

shortgrass steppe-desert grassland ecotone. Journal of Vegetative 

Science. 10:123-130.

Findley, J.S.1989. Morphological patterns in rodent communities 

of southwestern North America. Pages 253-263 in D.W. Morris, 

Z. Abramsky, B.J. Fox, and M.R. Willig (ed). Patterns in the 

structure of mammalian communities. The Museum, Texas Tech. 

University Spec. Publ. 266 p.

Forest Preserve District of Cook County. 1973. The Pocket Gopher. 

Nature Bulletin No. 493-A. 2 p. Available: URL http://www.newton.

dep.anl.gov/natbltn/400-499/nb493.htm

Forest Preserve District of Cook County. 1974. Deer mice and whitefooted mice. Nature Bulletin No. 545-A. pp. 2. Available: URL 

http://www.newton.dep.anl.gov/natbltn/500-599/nb545.htm

Frogs and Toads. 2005. Frogs and toads. Available: URL http://allaboutfrogs.org/weird/general/frogtoad.html

Geiger, B. 2002. Mysterious Mounds. The mima mounds. Available: 

URL http://www.homestead.com/wintersteel/Mima_Mounds~ns4.

html

Geomyidae. 2005. Available: URL hppt://www.chez.com/rodent/

Geomyidae/Geomydae.html

Gila Monster. 2005a. NatureServe Explorer Available: URL http://www.

natureserve.org/explorer/servlet/NatureServe?init=Species

Gila Monster. 2005b. BISOM_M. Available: URL http://fwie.fw.vt.

edu/states/nmex_main/species/030135.htm

Gila Monster. 2005c. Arizona Game and Fish Department Heritage 

Data Management System. Available: URL http://www.azgfd.

gov/w_c/edits/documents/Helosuci.d_001.pdf

Gober, P. 2002. Dog 12-Month Finding the Significance of the 

Plague Threat in the U.S. Fish and Wildlife Service: Black-Tailed 

Prairie. Page 13 In Brand, C.J. (ed), 2002, Landscape Ecology 

of Plague in the American Southwest, September l9—20, 2000, 

Fort Collins, Colorado, Proceedings of an American Southwest 

Workshop: Fleston, Virginia, U.S. Geological Survey Information 

and Technology Report 2002—0001, 24 p.

Hatfield-Etchberger, L., B. Stroh, B. Bibles, and R. Etchberger. 2002. 

Investigating Epizootic Plague Transmission: Comparison of Flea/

Host Relationships Within an Established Prairie Dog Colony to 

Adjacent Habitat. Page 8 In Brand, C.J., 2002, Landscape Ecology 

of Plague in the American Southwest, September l9—20, 2000, 

Fort Collins, Colorado, Proceedings of an American Southwest 

Workshop: Fleston, Virginia, U.S. Geological Survey Information 

and Technology Report 2002—0001, 24 p.

Heteromyidae. 2000. Heteromyidae: kangaroo rats and pocket mice. 

Arizona-Sonora Desert Museum. Available: URL http://www.

desertmuseum.org/books/nhsd_heteromyidae.html

Heteromyidae. 2005. Heteromyidae. Wikipedia. Available: URL 

http://en.wikipedia.org/wiki/Hetermomyidae

Hof, J., M. Bevers, D.W. Uresk, 0.L. Sawtbeck. 2002. Population 

dependent dispersal: habitat placement for the black-tailed prairie 

dog. Spatial optimization in ecological applications. New York:

Columbia University Press. 114-124 p.

Howard, V.W. Jr. 1994, Kangaroo rats. Prevention and Control of 

Wildlife Damage – 1994. Cooperative Extension Division. Institute 
68 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

of Agriculture and Natural Resources. University of Nebraska 

– Lincoln. 4 p.

Huntly, N. and R. Inouye. 1988. Pocket gophers in ecosystems: 

patterns and mechanisms. BioScience 38: 786-793.

Hygnstrom, S.E. and D.E. Virchow. 1994. Prairie Dogs. Prevention 

and Control of Wildlife Damage – 1994. Cooperative Extension 

Division, Institute of Agriculture and Natural Resources, 

University of Nebraska – Lincoln, 11 p.

Jones, C. and R.W. Manning. 1996. The mammals. In:Range Wildlife, 

Kausmann, P.R. (ed). The Society of Range Management, Denver, 

Colorado. 440 p.

Kangaroo Rat. 2005a. Available: URL http://www.discverseaz.

com/Wildlife/Kangaroo_rat.html

Kangaroo Rat. 2005b. Kangaroo Rats. DesertUSA. Available: URL 

http://www.desertusa.com/aug96/du_krat.html

Kangaroo Rat. 2005c. Merriam’s Kangaroo Rat. The Mammals 

of Texas – Online Edition. Available: URL http://www.nsrl.ttu.

edu/tmotl/dipomerr.htm

Kaplan, M. 2002. Salamanders and newts. Available: URL http://

www.anapsid.org/sallies.html

Knowles, C. 2002. Status of White-tailed and Gunnison’s Prairie Dogs. 

National Wildlife Federation, Missoula, MT and Environmental 

Defense, Washington, DC. 30 p.

Kotliar, N.B. 2000. Application of the new keystone-species concept 

to prairie dogs: how well does it work? Conservation Biology. 

14:1715-1721.

Lagomorphs. 2005a. Rabbits, Hares, and Pikas. Available: URL 

http://www.bobpickett.org/order_lagomorpha.htm

Lagomorphs. 2005b. Rabbits. Wikipedia. Available: URL http://www.

snowspine.com/beastiary/rabbits.html

Mammals. 1997. Introduction to Mammals. Available: URL http://

www.educationalimages.com/it110007.htm

Mammals. 2005. Deer mouse. eNature. National Wildlife Federation. 

Available: URL http://www.enature.com/fieldguide/showSpeciesFT.asp?fotogID=750andcurPageNum=5andrecnum=MA0093

Meadow Vole. 2005. Meadow vole. Available: URL http://www.fcps.

k12.va.us/StratfordLandingES/Ecology/mpages/meadow_vole.

htm

Mexican Beaded Lizard. 2005. Lizards. Available: URL http://www.

corrystuart.net/child_lizards.html

Moodie, K.B. and T.R. Van Devender. 1979. Extinction and extirpation 

in the herpetofauna of the southern High Plains with emphasis on 

Geochelone wilsoni (Testudinidae). Herpetologica 35:198-206.

Mueggler, W.F. 1967. Voles damage big sagebrush in southwestern 

Montana. S. Range Manage. 20:88-90.

Naylor, V. 1994. Black-Footed Ferret. Badlands National Park. 

Available: URL http://www.northern.edu/natsource/ENDANG1/

Bfferrl.htm

Order Lagomorpha. 2005. Order Lagomorpha: Hares and Rabbits. 

The Mammals of Texas - Online Edition. Available: URL http://

www.nsrl.ttu.edu/tmot1/ordlagom.htm

Parmenter, R.R. and T.R. Van Devender. 1995. Diversty, spectial 

variability, and functional roles of vertebrates in the desert 

grasslands. In: The desert grassland. M.P, McClaran and T.R. Van 

Devender (eds). The University of Arizona Press, Tucson.

Parmley, D. 1990. Late Pleistocene snakes from Fowlkes Cave, 

Culberson County, Texas. Journal of Herpetology 24:266-274.

Pauli, J.N. 2005, Ecological studies of the black-tale prairie dog: 

implications for biology and conservation. Masters Thesis. 

University of Wyoming. Laramie. 77 p.

Payne, T. and W. Caire. 1999. Species diversity of small mammals 

in the tallgrass prairie preserve, Osage County, Oklahoma. Proc. 

Oklahoma Academy of Science. 79:51-59.

Payton, I.J., M. Fenner, and G.L. Williams. 2002. Keystone species: 

the concept and its relevance for conservation management in 

New Zealand. Science for Conservation 203. 29 p.

Pocket Gophers. 2005a. Arizona-Sonora Desert Museum. Available: 

URL http://www. desertmuseum.org/books/nhsd_gophers.html

Pocket Gophers. 2005b. Available: URL http://cecalaveras.ucdavis.

edu/pocket.htm

Pocket Gopher. 2005c. Pocket Gopher. Available: URL http://www.

holoweb.com/cannon/pocket.htm

Pocket Mice. 2005. Hispid Pocket Mouse. The Mammals of Taxas 

– Online Edition. Available: URL http://www.nsrt.ttu.eduu/tmit1/

cheahisp.htm

Preston, S. D. and L. F Jacobs. 2001. Conspecific pilferage but not 

presence affects Merriam’s kangaroo rat cache strategy. Behavioral 

Ecology. 12:517-523.

Reichman, J. 2004. Pocket gopher serves as ‘Ecosystem Engineers.’ 

Ascribe Science News Service. University of California, Santa 

Barbara. Available: URL http://www.highbeam.com

Reptiles. 2005. Animal Bytes: Reptiles. Avaibable: URL http://sandiegozoo.org/animalbytes/a-reptiles.html

Rogers, K.L. 1982. Herpetofaunas of the Courtland Canal and Hall 

Ash local faunas (Pleistocene: Early Kansas) of Jewell Co., Kansas. 

Journal of Herpetology 16:174-177.

Saimon, T.P. and W.P. Gorenzel. 2002. Voles. University of California. 

Available: URL http://www.imp.ucdavis.edu/PKG/PESTNOTES/

pn7439.html

Salamanders. 2002. U.S. Department of the Interior, U.S. Geological 

Survey Northern Prairie Wildlife Research Center, 8711 37th 

St. SE, Jamestown, ND 58401 USA Available: URL http://www.

npwrc.usgs.gov/narcam/idguide/atigrin.htm

Salamanders. 2005a. Order Caudata: Salamanders. Available: URL 

http://vernalpools.rtpi.org/guides/order.htm?order=Caudata

Salamanders. 2005b. Salamanders and newts. Maryland Department 

of Environmental Protection. Available: URL http://www.

montgomerycountymd.gov/deptmpl.asp?url=/content/dep/herps/

Newts.asp#top

Savage, J.M. 1960. Evolution of a peninsular herpetofauna. 

Systematic Zoology 9:184-212.

Saving a Species. 2005. Black-Footed Ferret Recovery Implementation 

Team. Available: URL http://www.blackfootedferret.org/savingcaptive.html

Scott, N.J. Jr. 1996. Evolution and management of the North 

American grassland herpetofauna. In: D.M. Finch (ed). Ecosystem 

disturbance and wildlife conservation in western grasslands – A 

symposium preceedings, September 22-26, 1994; Albuquerque, 

NM. General Technical Report RM-GTR-285. Fort Collins, 

CO: USDA Forest Service, Rocky Mountain Forest and Range 

Experiment Station. 82 p.

Severe, D.S. 1977. Revegetation of black-tailed prairie dog mounds 

on short-grass prairie in Colorado. MS. Thesis, Colorado State 

Univ., Fort Collins. 92 p.

Sevilleta LTER: Data. 1998a. Data: Species : Mammal : Bannertailed Kangaroo Rat - Dipodomys spectabilis Available: URL 

http://sevilleta.unm.edu/data/species/mammals/profile/bannertailed-kanfaroo-rat.html

Sevilleta LTER: Data. 1998b. Data: Species : Mammal : Botta’s Pocket 

Gopher – Thomomys bottae. Available: URL http://sevilleta.unm.

edu/data/species/mammal/sevillet/profile/bottas-pocket-gopher.

html

Sevilleta LTER: Data. 1998c. Data: Species : Mammal : Gunnison 

Prairie Dog. Available: URL http://sevilleta.unm.edu/data/species/mammal/sevillet/profile/gunnison-prairie-dog.html

Sevilleta LTER: Data. 1998d. Data: Species : Mammal : Merriam’s 

Kangaroo Rat - Dipodomys merriami Available: URL http://

sevilleta.unm.edu/data/species/mammal/sevillet/profile/merriams_kangaroo-rat.html

Sevilleta LTER: Research. 2001. Research: Local: Plants: Kangaroo 

Rats and Dominance Available: URL http://sevilleta.unm.edu/

data/contents/SEV114/kratdominance/

Sullivan, L.M. 2005. Pocket Gophers. University of Arizona. 

Available: URL http://www.ag.arizona.edu/urbanipm/rodents/

pocketgophers.html

Tracy, R.L. and G.E. Walsberg. 2002, Kangaroo rats revisited: 

re-evaluating a classic case of desert survival. Oecologia. 133:

449-457.

Turtles. 2005. Testudines: Life History and Ecology Available: URL 

http://www.ucmp.berkeley.edu/anapsids/testudines/testudineslh.

html

Uresk, D.W. and D.D. Paulson. 1988. Estimated carrying capacity for 

cattle competing with prairie dogs and forage utilization in western 

South Dakota. Pages 387-390 In R. C. Szaro, K.E. Severson, and 

D.R. Patton (eds). Proc. symposium management of amphibians, 

reptiles, and small mammals in North America. Fort Collins, 

CO: U.S. Department of Agriculture Rocky Mountain Forest and 

Range Experiment Station. Gen. Tech. Rep. RM-166.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 69

Wagner, D. and L.C. Drickamer. 2004. Abiotic habitat correlates 

of Gunnison’s prairie dog in Arizona. Journal of Wildlife 

Management. 68:188-197.

Western Box Turtle. 2005a. Western box turtle (Terrapene ornata). 

Arizona-Sonora Desert Museum. . Available: URL http://www.

desertmuseum.org/books/nhsd_terrapene.html

Western Box Turtle. 2005b. NatureServe Explorer Available: 

URL http://www.natureserve.org/explorer/servlet/

NatureServe?init=Species

Wiscomb, G.W. and T.A. Messmer. 1988. Pocket Gophers. Resources 

Department of Fisheries and Wildlife Utah State University, 

Logan, Utah. pp. 10 Available: URL http://extension.usu.edu?files/

natrpubs/gopher.html
70 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 71

Status and Conservation: 

Introduction_______________________

In the Southwestern United States, the grassland 

avifauna is collectively composed of a mixture of species found primarily in desert grasslands, shortgrass 

steppe, wet meadows, and alpine tundra (as used here, 

desert grasslands incorporate both arid grasslands 

and desert shrub grasslands). Of these habitats, 

desert grasslands and shortgrass steppe are the 

most extensive and support the greatest number of 

grassland bird species. Desert grasslands are patchily 

distributed across the southern halves of New Mexico 

and Arizona, and shortgrass steppe is a component of 

the Great Plains system that in the Southwest region 

extends across the eastern half of New Mexico into the 

panhandles of Texas and Oklahoma. Alpine tundra and 

particularly wet meadows are limited in geographic 

extent and support relatively few species of grassland 

birds in this region (see chapter 2 for detailed maps 

of the distribution of grassland types). Though their 

geographic extent may vary, all of these grassland 

systems provide habitat for distinctive grassland bird 

species in the Southwest and are therefore worthy of 

management concern.

The grassland bird community of the Southwest 

has been shaped by a variety of forces. The challenge 

of surviving in the arid climate, as well as a history 

of coevolution with the effects of grazing and wildfire, 

are some of the primary factors that have influenced 

the diversity of species found here. The relative lack 

of structural heterogeneity in grassland habitats has 

also played an important role in determining species 

composition, as the lack of shrubs or trees eliminates 

a variety of potential ecological niches for birds to 

exploit. This structural simplicity has resulted in an 

avifauna that tends to be characterized by specialists, 

species that have evolved within those few, specific 

niches available. Wilson’s phalarope, for example, is 

a wetland species that occurs only locally where water 

is available in the grassland landscape. Some species, including burrowing owls, are highly dependent 

upon active prairie dog towns, while others such as 

Baird’s sparrow coevolved with grazing ungulates and 

consequently seek out habitats with a mosaic of grass 

heights. Species such as horned larks are partial to 

grass cover so low that it is almost bare ground, while 

Botteri’s sparrow prefers the towering bunchgrasses 

of big sacaton (Sporobolus wrightii) grasslands.

The end result of these collective processes is an 

avian community that is simple in terms of species 

richness (few different species), but high in numbers 

of individual species. For example, one survey of the 

Pawnee National Grasslands in Colorado found that 

just three birds (horned lark, McCown’s longspur, and 

lark bunting) accounted for 87 percent of all 1,047 

individuals recorded (14 native species total; Knopf 

1996a); similar results have been reported in other 

Michele Merola-Zwartjes

Chapter 4:

Birds of Southwestern Grasslands:

Status, Conservation, and Management
72 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

grassland habitats as well (for example, Kantrud and 

Kologiski 1982). Figure 4-1 graphically illustrates the 

relatively low species diversity of grassland birds in 

the Southwest during the breeding season, based upon 

the results of the Breeding Bird Survey.1 However, it is 

important to note that diversity in this region increases 

markedly during the winter, due to the influx of short 

distance migrants. For example, a comparison of the 

distribution of lark buntings during the breeding season 

and in the winter demonstrates how high numbers 

of these grassland breeding birds concentrate in the 

Southwest during the winter months (fig. 4-2 and 4-3). 

For the lark bunting and numerous other species of 

grassland birds that do not travel all the way to the 

neotropics for the winter, the grassland systems of the 

Southwest provide habitat that is critically important 

for overwinter survivorship.

With the possible exception of rodents, birds generally outnumber all other vertebrate groups in grassland 

ecosystems (C. Bock personal communication 2002, 

Parmenter and Van Devender 1995). The low number 

of species reflects a limited number of ecological niches, 

1

 The Breeding Bird Survey (BBS), begun under the auspices of 

the U.S. Fish and Wildlife Services in 1966, consists of nearly 3,700 

randomly located permanent survey routes located along secondary 

roads throughout the continental United States and Canada; each route 

is surveyed annually during the breeding season (Sauer and others 

1997). The total number of individuals of each species recorded along 

the route is used as an index of relative abundance; long-term trends 

are determined by a route-regression method. Although the BBS has 

been criticized for a variety of factors, such as observer variability (see 

DeSante and George 1994 for a brief review), it nonetheless remains 

highly valuable as the most comprehensive and long-running quantitative survey of any vertebrate group in North America.

Figure 4-1. Species richness of grassland 

birds breeding in the United States and 

Canada. Numbers are average number of 

species detected per route on the Breeding 

Bird Survey, years 1982 to 1996 (from Sauer 

and others 1995).

Observed Number

of Species

<5

6-7

8-9

10-11

12-13

>13

resulting in a relatively specialized avifauna. While this 

specialization has allowed for the success of grassland 

endemics in the Southwest, this success has come at a 

price. Many species of grassland birds are apparently 

now in decline, due largely to the alteration or outright 

loss of the habitats that grassland specialists have 

historically depended upon.

The recent declines in numerous species of grassland 

birds have made them a high priority for conservation 

for both governmental agencies and nongovernmental 

organizations alike. Table 4-1 lists some of the bird 

species that characterize the grassland habitats of the 

Southwest and indicates their conservation status. 

Due to the relatively limited information available 

on birds of wet meadows and alpine tundra in the 

Southwestern region, the major focus of this review 

will be on birds inhabiting the desert grassland and 

shortgrass steppe ecosystems. Information specific to 

Southwestern grasslands is used whenever possible. 

However, as the majority of research on grassland birds 

of the United States has been focused on the habitats 

of the Great Plains region, in many cases data from 

Arizona and/or New Mexico are simply not available. 

Research from other regions may be used when data 

from the Southwest is lacking, and will be so noted.

Population Trends and Status

Concerns over apparent declines in populations of 

North American birds, particularly migratory songbirds, have been growing since the 1960s (for example, 

Aldrich and Robbins 1970, Ambuel and Temple 1982, 

Briggs and Criswell 1979, Carson 1962). At least 

initially, decreasing numbers of neotropical migrants 

in the Eastern U.S. forests were the primary focus 

of these worries (for example, Robbins and others 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 73

1989). More recent analyses, however, suggest that 

the negative population trends of North American 

birds are most apparent in the grassland habitats 

of the United States. Many researchers now believe 

that grassland bird populations have been declining 

across the continent for approximately the last 50 years 

(Askins 1993, Peterjohn and Sauer 1999). Grassland 

birds, it is claimed, have experienced “steeper, more 

consistent, and more geographically widespread 

declines than any other behavioral or ecological guild” 

(Knopf 1994:251). These declines have been largely 

due, it is believed, to the degradation and outright 

loss of grassland habitats (Collar and others 1992, 

DeSante and George 1994, Herkert and others 1996, 

Knopf 1994, Stotz and others 1996, Vickery and others 

1999b). Habitat loss or alteration in these systems 

has occurred as the result of conversion of grasslands 

to agriculture, overgrazing or inappropriate grazing 

Figure 4-2. Distribution and densities of breeding lark buntings in the 

United States and Canada, as mean 

numbers of individuals detected per 

route per year. Data from Breeding 

Bird Surveys averaged over the 

years 1982 to 1996 (Sauer and 

others 2001).

Figure 4-3. Winter distribution and densities of 

lark buntings, based on 

Christmas Bird Count 

data. Counts are average number of birds detected per survey over 

the years 1982 to 1996 

(from Sauer and others 

2001). Like many shortdistance migrants, this 

species concentrates 

in the grasslands of the 

Southwest for the winter 

months.

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Average Count

< 1

1-3

3-10

10-30

30-100

> 100
74 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Common namea

Mountain plover • B • Highest priority 

Long-billed curlew • B Highest priority 

Wilson’s phalarope • B Highest priority 

Common snipe • Y, W 

Golden eagle • • Y 

Northern harrier • • • Y,W • 

Swainson’s hawk • • B • Ferruginous hawk ST-AZ • • Y, W • Highest priority • Prairie falcon • • Y Highest priority 

Northern Aplomado falcon • X • E, SE-AZ, SE-NM

White-tailed ptarmigan • Y Highest priority SE-NM

Lesser prairie chicken C • Y 

Masked bobwhite E, SE-AZ • Y 

Scaled quail • • Y High responsibility 

Short-eared owl • • • • M,W 

Burrowing owl • • B, Y • High responsibility • Common nighthawk • • B High responsibility 

Lesser nighthawk • B High responsibility 

Broad-tailed hummingbird • B High responsibility 

Western kingbird • • B 

Cassin’s kingbird • • B High responsibility 

Scissor-tailed flycatcher • B Highest priority 

Ash-throated flycatcher • • B 

Say’s Phoebe • • B, Y High responsibility 

Horned lark • • • Y 

Chihuahuan raven • • B, Y High responsibility 

Cactus wren • Y 

Mountain bluebird • • • • Y, W Priority 

Loggerhead shrike • • Y • Priority 

Bendire’s thrasher • • B, Y • Highest priority 

American pipit • B 

Sprague’s pipit SC-AZ • W Highest priority 

Common yellowthroat • B 

Grasshopper sparrow • • • B, W • Highest priority • (includes Arizona 

 grasshopper sparrow 

 ST-NM)

Baird’s sparrow ST-AZ, • W • Highest priority • ST-NM

Vesper sparrow • • B, W 

Savannah sparrow • B, W Priority 

Lark sparrow • • B, Y • 

Black-throated sparrow • B, Y • Desert grasslands Shortgrass steppe (Great Plains) Wet meadows Alpine tundra Season of useb USFWS migratory nongame bird of management concernc

NM Partners inFlight priorityrankingdAZ Partners inFlight priorityspeciese

Table 4-1. Partial list of bird species typically found in the various grassland types of Arizona and/or New Mexico, with particular 

emphasis on those species that are of conservation concern. Common names in bold face type indicate that the species is listed 

as endangered or threatened under Federal and/or State law (see superscript for specific designation). This list is not meant to 

be exhaustive but rather presents some of the more characteristic species likely to be found in these vegetation types. Species 

are presented in taxonomic order.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 75

regimes, the introduction of exotic grasses, fire suppression, succession to shrublands, and fragmentation 

(Herkert 1994, Vickery and others 1994, 1999a,b).

In one of the first studies to specifically address 

declines in Western birds, DeSante and George (1994) 

produced a list of 75 native species whose breeding 

populations had declined substantially in at least one 

Western State over the past 100 years. For Western 

birds, destruction of riparian habitats was the probable 

cause of decline for the greatest number of species (16), 

followed closely by destruction of grassland habitats 

(15), shooting (13), overgrazing (9), cowbird parasitism 

(7), logging and clearing of forests (7), and other causes. 

Table 4-2 presents a subset of this comprehensive 

list, highlighting those species of the Western United 

States for which destruction of grassland habitats, 

overgrazing, or cowbird parasitism—in other words, 

those factors most relevant to our discussion of 

Southwestern grasslands—have been implicated in 

the reported population declines.

These same authors also considered data collected 

from the Breeding Bird Survey over a period of 26 years, 

from 1966 to 1991 (DeSante and George 1994). In this 

population trend analysis, Western birds were divided 

into two groups: long-distance migrants that winter in 

the neotropics, and short-distance migrants that winter 

in the temperate zone of North America, primarily the 

Southwestern United States and Mexico. Of the shortdistance migrants that displayed significant population 

trends, 79 percent were declining while only 21 percent 

showed increasing trends. In contrast, long-distance 

migrants did not undergo any significant population 

trends in the Western United States during this same 

Sage sparrow • W • 

Botteri’s sparrow • • B • Highest priority • Cassin’s sparrow • • B, Y • High responsibility • Rufous-winged sparrow • Y • McCown’s longspur • • W Highest priority 

Chestnut-collared longspur • • W 

Dickcissel* • B Highest priority 

Lark bunting • B, Y, W Highest priority 

Bobolink* SE-AZ • B Highest priority 

Western meadowlark • • • Y 

Eastern meadowlark • • • Y Priority 

 (includes southwestern 

 subspecies lilianae)

Red-winged blackbird • Y 

Brewer’s blackbird • Y, W 

Brown-headed cowbird • • B, Y 

Brown-capped rosy finch • Y Highest priority 

a

 Superscripts following name indicate the following Federal Endangered Species Act designations: E = Endangered, PT = Proposed Threatened, 

C = Candidate (USFWS 2002). State designations are as follows: SE = State endangered, ST = State threatened, SC = State candidate (AGFD 

1988; NMDGF 2002). An asterisk * indicates highly localized breeding populations. b

 B = breeding, W = wintering, Y = year-round, M = in migration. X = currently extirpated; accidental sightings may occur. More than one designation 

indicates that season of use differs depending upon geographic location. c

 In Southwest Region (USFWS 1995).

d

 Based on New Mexico Partners in Flight Bird Conservation Plan (NMPIF 2001).

e

 Based on Arizona Partners in Flight Bird Conservation Plan (Latta and others 1999). Note that species in the Arizona plan are not broken down 

into separate ranking categories, as in the New Mexico plan. However, individual priority scores are given for each species in appendices B 

and D of the plan.

Table 4-1. Continued.

Lark bunting. (Photo by Gary Kramer, courtesy of USDA Natural 

Resources Conservation Service)
76 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Table 4-2. Native species of landbirds that have decreased in the Western United States for reasons related to grassland habitat 

destruction, overgrazing, or cowbird parasitism, based on a review of major State-level literature. Listed are species that are 

known to have decreased in the last 100 years. The superscript E

 denotes species that have been extirpated; species with an 

asterisk * are those whose breeding populations have decreased by more than 50 percent. Modified from DeSante and George 

(1994).

Western States where 

Species populations have declined Probable cause

Sage grouse WA*, OR*, CA*, NV*, ID, MT, UT*, NME

 Shooting, overgrazing, destruction of grasslands

Greater prairie-chicken MTE

, CO* Shooting, destruction of grasslands

Lesser prairie-chicken CO*, NM* Shooting, destruction of grasslands

Sharp-tailed grouse WA*, ORE

, CAE, NVE, ID*, MT*, WY*, Shooting, overgrazing, destruction of grasslands

 UT*, CO*, NME

Montezuma quail AZ*, NM* Overgrazing

Northern bobwhite WY, AZE

, NM* Overgrazing, hunting

Scaled quail AZ*, NM* Overgrazing, destruction of grasslands

Burrowing owl CA*, NV*, ID, MT, CO*, AZ, NM* Destruction of grasslands, elimination of fossorial 

 mammals, agricultural development, 

 urbanization

Short-eared owl CA*, NM* Destruction of grasslands

Willow flycatcher CA*, AZ* Destruction of riparian habitat, cowbird parasitism

Horned lark AZ Destruction of grasslands

Bell’s vireo CA*, AZ* Cowbird parasitism, destruction of riparian habitat

Gray vireo CA*, AZ Cowbird parasitism?

Yellow warbler OR, CA*, AZ* Cowbird parasitism, destruction of riparian habitat

Common yellowthroat CA, AZ Drainage of marshes and loss of riparian habitat, 

 cowbird parasitism?

Yellow-breasted chat CA, NV* Destruction of riparian habitat, cowbird parasitism?

Botteri’s sparrow AZ* Overgrazing, destruction of grasslands

Rufous-winged sparrow AZ* Overgrazing

Chipping sparrow WA, OR Unknown (cowbird parasitism?)

Vesper sparrow WA, OR Destruction of grasslands?

Lark bunting CA, NV Destruction of grasslands

Baird’s sparrow AZ* Destruction of grasslands

McCown’s longspur AZ*, NM* Destruction of grasslands

Chestnut-collared longspur AZ, NM* Destruction of grasslands

period. These striking results point up an interesting 

difference between migratory bird population trends in 

the Eastern and Western United States. In the East, 

we have evidence for accelerating declines in forestdwelling, long-distance neotropical migrants (Askins 

and others 1990, Robbins and others 1989, Sauer and 

Droege 1992, Terborgh 1989). The West, by contrast, 

is witnessing declines in short-distance migrants, and 

especially those species associated with grasslands 

and shrublands (DeSante and George 1994). Several 

reasons are suggested for these declines in the West, 

including the destruction of grassland ecosystems, 

overgrazing, and increased levels of cowbird parasitism (DeSante and George 1994). There may also be a 

connection between the declines in several species of 

Western short-distance migrants and the degradation 

of their wintering habitat (again primarily grasslands 

and shrublands) in the Southwestern United States 

and Northwestern Mexico (see table 4-3), although 

researchers warn of the lack of sound information on 

the wintering ecology of western migratory birds (for 

example, Knopf 1994).

Other approaches to analysis of population trends 

from BBS data have produced similar results, in which 

grassland birds in particular appear to be declining 

in numbers. In their analysis of 31 years of BBS 

data, covering the period from 1966 through 1996, 

Peterjohn and Sauer (1999) propose that approximately 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 77

three-quarters (77 percent) of our grassland bird species have been declining in numbers since at least the 

1960s, nearly half of them (48 percent) significantly 

so (table 4-4).

In response to the growing concerns over grassland 

bird numbers, many species are now considered priority targets for conservation by Partners In Flight 

(a State, Federal, and private partnership working 

to conserve birds in the Western Hemisphere) and 

the U.S. Fish and Wildlife Service (see table 4-1). 

The northern aplomado falcon and masked bobwhite 

are officially listed as endangered, and the lesser 

prairie chicken is a candidate for listing under the 

Endangered Species Act. Several other grassland birds 

have been considered for listing as well (for example, 

Baird’s sparrow, mountain plover). The reasons for 

the declines of these grassland birds are myriad, but 

several factors appear to be common across many of 

the species.

The sections that follow present a brief overview 

of some of the most prevalent threats to grassland 

birds in the Southwest today, including various 

agents contributing to habitat loss, alteration and 

fragmentation; the effect of declining prairie 

dog populations; pesticide impacts; cowbird 

parasitism; and the consequences of haying 

and mowing practices.

Habitat Loss, Alteration, and 

Fragmentation_______________

Habitat loss or destruction has had a major 

impact on grassland birds across the continental United States. Much recent media attention 

has focused on the loss of the tallgrass prairie 

in the Eastern United States and central 

Great Plains region. Because of its suitability 

for conversion to agriculture, it is estimated 

that the tallgrass prairie has been reduced to 

less than 2 percent of its original extent, and 

even less than that remains in some States 

(Samson and Knopf 1994). Although they have 

not received the same degree of popular coverage, the desert grasslands of the Southwest 

have also experienced dramatic losses over the 

past century. One study of desert grasslands in 

southern New Mexico suggests that although 

about 75 percent of this region was covered 

in grasslands prior to the late 1800s, by the 

late 1960s only 5 percent grassland coverage remained (York and Dick-Peddie 1969). 

Areas that had formerly been characterized 

as supporting “heavy growth” or “excellent 

stands” of grama grass (Bouteloua spp.) are 

now shrublands dominated by creosotebush 

(Larrea tridentata), mesquite (Prosopis spp.), or juniper 

(Juniperus spp.) (York and Dick-Peddie 1969). Although 

debate continues over the mechanism(s) that may 

have promoted this conversion, there is little question 

that the historic grasslands of the Southwest have 

experienced a rapid and dramatic invasion of woody 

plants and cacti since Euro-American settlement 

(Bahre 1995, Humphrey 1958, York and Dick-Peddie 

1969). In assessing human impacts on the grasslands 

of southeastern Arizona, Bahre (1995:231) concludes 

“the two most dramatic changes in the grasslands are 

the extensive increases in woody shrubs and trees and 

the landscape fragmentation resulting from localized 

urban and rural settlements.”

Although more of the shortgrass prairie remains 

in Western States (approximately 60 percent; Weaver 

and others 1996), these Southwestern grasslands are 

also slowly but steadily being lost to urban sprawl 

and conversion to agriculture, among other causes. 

Dick-Peddie (1993) proposes that the primary reason 

for loss of the shortgrass prairie in eastern New 

Mexico is farming (both dryland and irrigation), and 

secondarily due to urbanization and livestock grazing. 

Studies show that, at least in other grassland regions, 

Table 4-3. Short-distance migrants that winter primarily in grasslands of 

the Southwestern United States or Northwestern Mexico with decreasing population trends during either the Breeding Bird Survey census 

periods 1966 to 1991 (26 years) and 1979 to 1991 (13 years; last half 

of census period) based on the Population Trend Ranks of Carter and 

Barker (1993). D indicates a decreasing trend, I an increasing trend, 

---- indicates a trend that was not significant in either direction. Modified 

from DeSante and George (1994).

26 years of BBS Last 13 years of BBS

1966-1991 1979-1991

Species Trend # of States Trend # of States

Burrowing owl D 5 D 5

Short-eared owl D 1 D 4

Common poorwill D 1 --- 3

Say’s phoebe D 10 --- 10

Horned lark D 11 D 11

Sprague’s pipit D 1 D 1

Loggerhead shrike --- 9 D 9

Chipping sparrow D 10 D 11

Brewer’s sparrow D 7 D 7

Black-chinned sparrow D 1 D 1

Black-throated sparrow D 5 --- 5

Baird’s sparrow* I 1 D 1

Grasshopper sparrow D 6 --- 6

Fox Sparrow D 1 --- 2

Song sparrow D 8 --- 8

White-crowned sparrow D 5 D 7

Eastern meadowlark D 1 D 1

Western meadowlark D 10 --- 11

Brewer’s blackbird D 8 --- 9

* Species that increased during one time period but decreased during the other.
78 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Table 4-4. Continental trend estimates for North American grassland birds, based on Breeding 

Bird Survey data over a 31-year period from 1966 to 1996. Species with significant declines 

are in bold type; ** denotes a p-value of 0.01<p<0.05; *** denotes a p-value of < 0.01. Species in italics are those with decreasing trends that are not statistically significant. N is the 

number of BBS routes for which each species has been recorded. Modified from Peterjohn 

and Sauer (1999).

1966-1996

 Mean # of

Mean % change individuals

Species per year p N per route

Northern harrier -0.6 --- 891 0.49

Ferruginous hawk 5.2 *** 186 0.25

Ring-necked pheasant Introduced -1.0 ** 1206 7.30

Sharp-tailed grouse 0.3 --- 124 0.55

Mountain plover -2.7 ** 33 0.31

Upland sandpiper 1.3 *** 581 2.22

Long-billed curlew -1.4 --- 202 1.45

Short-eared owl -2.8 --- 132 0.21

Horned lark -1.3 *** 1805 27.02

Sedge wren 2.2 ** 307 1.15

Sprague’s pipit -4.7 *** 108 1.41

Dickcissel -1.6 *** 783 16.29

Cassin’s sparrow -2.5 *** 203 16.31

Vesper sparrow -0.8 ** 1462 7.84

Lark bunting -0.9 --- 332 42.97

Savannah sparrow -0.6 ** 1477 8.40

Baird’s sparrow -1.6 --- 115 1.87

Grasshopper sparrow -3.6 *** 1404 3.97

Henslow’s sparrow -8.8 *** 149 0.15

LeConte’s sparrow 1.4 --- 154 0.73

McCown’s longspur 1.1 --- 59 4.57

Chestnut-collared longspur -0.1 --- 145 9.27

Bobolink -1.6 *** 1134 5.35

Eastern meadowlark -2.6 *** 1921 20.29

Western meadowlark -0.6 ** 1480 44.48

some birds may continue to utilize agricultural fields, 

particularly those used for haycropping, as surrogate 

grasslands (Herkert 1994; although useful for forage, 

hay fields have potential negative impacts on nest success; see below). However, as these fields are converted 

to rowcropping—a trend that has been increasing over 

the past 50 years—their habitat value is lost (Herkert 

1994). Haycropping and rowcropping are far more common in the Great Plains region than in the Southwest. 

Agricultural fields in the Southwest are more likely 

to be planted in chile or cotton rather than grains 

and grasslike crops, and therefore are not as likely to 

provide suitable habitat for grassland birds.

Like the Great Plains, the grasslands of the 

Southwest have suffered significant direct losses of 

cover through various agents. This destruction of 

grassland habitats is one of the most prevalent factors 

implicated in the declines of Western grassland birds 

(DeSante and George 1994). Even if the changes are 

not permanent but potentially reversible, grassland 

systems may be altered or degraded to the point that 

they no longer provide suitable habitat for some species of birds, and are thus essentially “lost.” Altered 

fire regimes, shrub encroachment, grazing by domestic 

stock, and introduced grasses are some of the factors that contribute to habitat alteration and loss 

through changes in species composition and habitat 

structure.

Altered Fire Regimes and Shrub 

Encroachment—Fire plays a critically important 

role in most grassland systems. Without periodic fires, 

woody plants begin to take hold and invade grasslands, 

converting them to shrublands or woodlands. The 

grasslands of the Southwest are no exception. Many 

researchers agree that historically fires were both 

common and extensive in the desert grasslands, and 

that these fires were instrumental in maintaining the 

integrity of these systems (Bahre 1991, Humphrey 

1958, but see Dick-Peddie 1993). The exception may 

be grasslands dominated by black grama (Bouteloua
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 79

eriopoda). Black grama suffers severe negative impacts 

from burning, indicating that this species is not fireadapted and probably did not evolve under a history 

of frequent burning (Buffington and Herbel 1965, 

Dick-Peddie 1993).

The natural frequency and extent of grassland 

fires in the Southwest are believed to have declined 

dramatically since Euro-American settlement of the 

region in the late 1800s (Bahre 1991, 1995, Humphrey 

1958). A review of the role of fire in desert grasslands 

reveals that the natural frequency of fire in these 

systems was probably on the order of every 7 to 10 

years (McPherson 1995 and references therein). Fires 

occurring on this cycle are believed to be sufficient to 

prevent the establishment of woody plants by killing 

seeds on the surface and preventing woody plants 

from reaching the age where resprouting is possible 

(McPherson 1995). Although fires eliminate grass 

cover in the short term, in the long term grasses are 

rejuvenated by the occurrence of fire and benefit from 

the elimination of woody plants. The timing of fires 

is also important. Fire in the early summer, when 

the growth of many perennials is just beginning, can 

negatively impact warm season grasses, whereas these 

same grasses are tolerant of fire during the dormant 

season (McPherson 1995).

Although many factors contribute to fire regimes, 

perhaps the most important change that has resulted in 

decreased fire frequency and intensity in the Southwest 

is the lack of fine fuels to carry the fires (Humphrey 

1958, McPherson 1995). Historically, the timing of this 

change corresponded with the widespread increase 

in livestock grazing in the Southwest after 1880. At 

this time, stocking rates reached record levels, and 

overgrazing was actually encouraged to reduce the 

fire hazard and encourage the growth of trees (Bahre 

1991, Leopold 1924, as cited in McPherson 1995). Today, 

fragmentation from roads and suburban developments 

also acts to contain the spread of extensive wildfires 

(Bahre 1995, McPherson 1995). These changes in the 

frequency and intensity of natural fire regimes have 

doubtless contributed to the widespread conversion of 

Southwestern grasslands to shrublands (Archer 1989, 

Brown 1982, Humphrey 1958). Grazing by livestock 

and the eradication of prairie dogs may also play a role 

in increasing woody plant cover, as discussed below.

The net result of the conversion of grasslands to 

shrublands is a loss of habitat for grassland birds such 

as grasshopper sparrows that avoid areas with woody 

plant cover. Changes in fire regimes also impact the 

physical structure and plant species composition of 

grasslands, thereby impacting the potential habitat or 

food resources utilized by grassland birds (McPherson 

1995). Species that use areas of bare ground or short 

grass cover, such as horned larks or lark sparrows, are 

negatively impacted by a reduction in fire frequencies. 

Vesper sparrows and other species that depend upon 

increases in seed-producing plants following fires also 

suffer negative consequences from fewer fires. Some 

grassland birds, such as Cassin’s sparrows, do benefit 

from an increased shrub component in their habitat. 

However, as the invasion of woody plants has now 

become so extensive throughout the grasslands of the 

Southwest, it is the more specialized pure grassland 

endemics that will be suffering the negative impacts 

of this form of habitat loss.

To further complicate matters, McPherson (1995) 

notes that once woody plants come to dominate a 

grassland system, the reintroduction of fire alone is 

insufficient to return it to its original composition. 

The use of herbicides or mechanical controls must be 

introduced to restore grassland habitat after woody 

plants have become established. Also, some exotic 

grasses such as Lehmann lovegrass increase after 

fire, and such grasses provide more fine fuel to carry 

fires than native species of grasses (Cox and others 

1984). This may result in a positive feedback loop, 

in which introduced grasses play a beneficial role in 

terms of increasing fire frequency, yet also have the 

negative consequences of extending the coverage of the 

less-desirable exotic grasses as well as increasing the 

intensity of fire (Anable and others 1992, see “Exotic 

Grasses,” below).

Livestock Grazing—Livestock grazing is the 

predominant land use in the Western States. More 

than 70 percent of the land area in the West (11 States, 

from Montana, Wyoming, Colorado, and New Mexico 

westward) is grazed by livestock, predominantly cattle, 

including wilderness areas, wildlife refuges, National 

Forests, and some National Parks (Fleischner 1994 and 

references therein). Inappropriate livestock grazing 

is widely believed to be one of the greatest sources of 

habitat degradation in the West (for example, Noss 

and Cooperrider 1994), leading to widespread declines 

in the native wildlife of North American grasslands 

(Fleischner 1994). However, the impacts of grazing vary 

widely according to grazing intensity or stocking rates, 

season of grazing, the species of livestock involved, 

and the degree of active management, such as pasture 

rotation (Fleischner 1994, Jones 2000). Although much 

has been written about the overall ecological impacts 

of grazing (for example, Fleischner 1994, Hobbs and 

Huenneke 1992, Noss and Cooperrider 1994, Saab 

and others 1995, Jones 2000), this discussion will only 

briefly review the impacts of grazing with respect to 

grassland birds of the Southwestern States.

The response of grassland birds to livestock grazing varies widely according to the species of bird in 

question, the type of habitat being grazed, and the 

ecological history of the region. In general, birds are 

not directly affected by the presence of livestock per 

se, but rather experience the more indirect effects of 
80 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

grazing impacts on grassland vegetation and ecosystem dynamics (Bock and Webb 1984; the exception is 

ground-nesting birds, which may be directly impacted 

by livestock during the breeding season; see below). 

The shortgrass prairie region was historically grazed 

by large numbers of native herbivores, primarily bison, 

pronghorn, and prairie dogs. As bison and pronghorn 

preferentially graze near prairie dog towns, this 

resulted in concentrating areas of disturbance due to 

heavy grazing and trampling, resulting in a mosaic of 

microhabitats ranging from areas of bare ground to 

dense concentrations of grasses (Coppock and others 

1983, Knopf 1996a, Krueger 1986). Native birds of the 

shortgrass prairie evolved within this grazed mosaic, 

and the various species adapted to different aspects of 

the landscape or evolved to utilize a mixture of microhabitat types (for example, for nesting versus foraging). 

Birds such as mountain plovers or McCown’s longspurs 

will use areas of excessive grazing pressure (Knopf 

and Miller 1994, Warner 1994), while species such 

as long-billed curlews, lark buntings, and Sprague’s 

pipits use shortgrass landscapes that have been heavily grazed but require a mixture of interspersed taller 

grasses for nesting (Bicak and others 1982, Finch and 

others 1987, Kantrud 1981). Although historically 

grazing provided such a mixture of conditions, most 

current livestock grazing regimes do not result in the 

heterogeneous grassland structure that was produced 

by native herbivores. Bison were nomadic and grazed 

over large expanses of the landscape. As bison were 

eliminated and livestock came to dominate the Western 

range in the late 1800s, homesteaders began to fence 

the grasslands, reducing the natural variability in 

the grazing behavior of the livestock and resulting 

in standardized grazing intensities and relatively 

homogeneous structure of the grasslands, thereby 

eliminating the habitat heterogeneity required by 

many grassland birds (Knopf 1996b).

The blue grama (Bouteloua gracilis) and buffalo 

grass (Buchloe dactyloides) of the shortgrass prairie 

coevolved with bison and are adapted to heavy grazing 

pressure; these grasses thrive under such conditions by 

reproducing both sexually and by tillering (Knopf 1994). 

By contrast, the desert grasslands of the Southwest 

probably evolved in the absence of large herds of grazing ungulates. Most evidence points to an absence 

of large herds of bison west of the Rockies (Durrant 

1970, Gustafson 1972, as cited in Mack and Thompson 

1982). The arid grasslands of the Southwest tend to 

be dominated by caespitose bunchgrasses, which are 

highly susceptible to grazing by ungulates and which 

respond in a manner quite distinctive from the grasses 

of the shortgrass prairie (Daubenmire 1970, Dyer 1979, 

Tisdale 1961). Whereas grazed areas in the shortgrass 

prairie tend to be recolonized by predominantly native plants (Mack and Thompson 1982 and references 

therein), the morphological and physiological features 

of bunch grasses render them incapable of recovering 

quickly from grazing. Continuous grazing in desert 

grasslands leads to changes in species composition, 

where bunch grasses are replaced by sod-forming 

grasses or annuals (Brown 1982), or invaded by 

Eurasian weeds (Mack and Thompson 1982, see also 

Milchunas and others 1988). Furthermore, the soils of 

these grasslands that evolved with few native grazers 

are protected by a cryptogamic crust of mosses, lichens, 

and liverworts; this crust can be permanently destroyed 

by the trampling of large ungulates, producing sites 

for the establishment of exotic species (Daubenmire 

1970, Mack and Thompson 1982).

Grazing has also been implicated in the widespread increase in woody plant cover witnessed in 

Southwestern grasslands (Bahre 1995, Dick-Peddie 

1993, Humphrey 1958, York and Dick-Peddie 1969). 

Grazing by livestock encourages invasion by woody 

plants by reducing the fine fuels available for fire 

and by facilitating the dispersal and establishment 

of propagules of woody invaders (Archer 1989, Bock 

and Bock 1987, 1988, Hobbs and Huenneke 1992, 

Humphrey 1987). Livestock may also promote the 

growth of woody vegetation by selectively grazing 

on the more palatable species of grasses and herbs, 

effectively reducing competition within the plant community for the less palatable woody plants (Humphrey 

1958). Many Southwestern grasslands subjected to 

livestock grazing, in combination with drought, have 

thus been degraded and transformed to landscapes 

of desert scrub (Buffington and Herbel 1965, Neilson 

1986, Schlesinger and others 1990). This change in 

vegetation composition and structure greatly affects 

the suitability of the habitat for native grassland birds, 

as well as for other animal species.

Birds that prefer excessively grazed areas, such as 

horned larks, tend to respond positively to grazing 

wherever they are found (Saab and others 1995). For 

other species, it depends on the region of interest. 

Grasshopper sparrows or lark buntings, for example, 

may respond positively to moderate levels of grazing 

in tallgrass prairies, but show negative effects when 

the grazing takes place in shortgrass prairies or desert grasslands (Saab and others 1995 and references 

therein). Ferruginous hawks respond differentially to 

grazing depending on the stage of life cycle in question; 

grazing benefits these hawks by opening up grasslands 

for foraging, but they prefer ungrazed areas with more 

substantial cover for nesting (Saab and others 1995). 

Several species of grassland birds have responded 

negatively to grazing wherever they have been studied; 

these include the common yellowthroat and Botteri’s, 

Cassin’s, savannah, Baird’s, and Henslow’s sparrows 

(Saab and others 1995). Ground-nesting species are 

especially vulnerable in areas that are grazed. Temple 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 81

and others (1999) report that 64 percent of nests 

were lost when cattle were introduced into a pasture 

with good grass cover, and experiments with artificial 

ground nests have similarly demonstrated a high rate 

of nest loss from trampling and nest abandonment 

due to cattle disturbance (Koerth and others 1983, 

Paine and others 1996). Table 4-5 shows the observed 

responses of various species of neotropical migratory 

birds to cattle grazing in Southwestern grasslands.

Exotic Grasses—In Arizona, grasslands that 

have been seeded with Lehmann and Boer lovegrass 

(Eragrostis lehmanniana and E. curvula var. conferta), exotic grasses used for cattle forage, have been 

described as “biologically sterile” (Bock and others 

1986:462). Twenty-six native species (10 plants, five 

birds, three rodents, and eight grasshoppers) were 

found to be significantly more abundant in native 

grasslands; only three native species (one bird, one 

rodent, and one grasshopper) were more common in the 

grasslands dominated by the African lovegrasses. Bock 

and others (1986:462) explain: “Indigenous animals 

appear to have evolved specific dependencies on the 

native flora and/or its associated fauna, insofar as most 

find the exotic grasslands far less inhabitable.”

This is a mounting problem in the West, as exotic 

grasses that are either intentionally introduced or 

invade following disturbance (for example, cheatgrass 

Bromus tectorum and crested wheatgrass Agropyron 

cristatum) may spread extremely quickly and displace 

native grasses (Mack 1981, Marlette and Anderson 

1986). Exotics such as Lehmann lovegrass also 

increase in response to grazing, as livestock tend to 

preferentially forage on the native grasses, reducing 

competition for the lovegrasses (Bahre 1995 and 

references therein). The increase of such exotics may 

have further ramifications as well, as they alter the 

natural fire regimes and lead to further ecological 

changes in the system, as discussed above (for example, 

Anable and others 1992, Cox and others 1990). This 

issue is discussed in more detail in chapter 2 of this 

volume.

Habitat Fragmentation—Fragmentation of 

habitats has been called “one of the greatest threats 

to biodiversity worldwide” (Noss and Cooperrider 

1994:51). There are essentially three components 

to habitat fragmentation. First, any fragmentation 

by processes such as the building of roads or urban 

development necessarily results in the outright loss 

of some portion of the original habitat. The second 

component is fragmentation per se, in which the remaining habitat is reduced to small, isolated patches 

across the landscape. “Edge effects” make up the third 

component of fragmentation. These effects refer to the 

alteration of physical processes and biotic interactions 

Table 4-5. Response of neotropical migratory birds to grazing on grassland types found in the Western United States (semidesert, 

shortgrass, and mixed grassland types). Note that due to a paucity of data from the Southwest, many of the responses reported 

here are based on studies from different geographical regions. Modified from Saab and others (1995).

Species Region Grassland type Grazing intensitya

 Effectb Reference

Northern harrier S. Dakota Mixed grass Moderate - Duebbert and Lokemoen 1977

N. Plains Mixed grass Variable - Kantrud and Kologiski 1982

Ferruginous hawk N. Plains Mixed grass Variable + Kantrud and Kologiski 1982

S. Dakota Mixed grass + Lokemoen and Duebbert 1976

Killdeer Colorado Shortgrass Heavy + Ryder 1980

N. Plains Mixed grass Moderate, heavy + Kantrud and Kologiski 1982

Mountain plover Colorado Shortgrass Heavy + Graul 1975

Colorado Shortgrass Heavy + Ryder 1980

N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

N. Plains Mixed grass Heavy + Kantrud and Kologiski 1982

Long-billed curlew Colorado Shortgrass Heavy + Ryder 1980

N. Plains Mixed grass Moderate, heavy 0 Kantrud and Kologiski 1982

Mourning dove S. Arizona Semidesert Moderate + Bock and others 1984

Colorado Shortgrass Heavy 0 Ryder 1980

N. Plains Mixed grass Heavy Mixed Kantrud and Kologiski 1982

Burrowing owl N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

N. Plains Mixed grass Heavy + Kantrud and Kologiski 1982

Short-eared owl S. Dakota Mixed grass Moderate - Duebbert and Lokemoen 1977

N. Plains Mixed grass Variable - Kantrud and Higgins 1992

Common nighthawk Colorado Shortgrass Heavy + Ryder 1980

N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

N. Plains Mixed grass Heavy + Kantrud and Kologiski 1982
82 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Table 4-5. Continued.

Species Region Grassland type Grazing intensitya

 Effectb Reference

Horned lark S. Arizona Semidesert Moderate + Bock and others 1984

Colorado Shortgrass Heavy + Ryder 1980

Plains Shortgrass Heavy + Wiens 1973

Saskatchewan Mixed/short + Maher 1979

Alberta Mixed grass Heavy + Owens and Myres 1973

S. Dakota Mixed grass Heavy + Wiens 1973

N. Plains Mixed grass Moderate, heavy + Kantrud and Kologiski 1982

Northern mockingbird S. Arizona Semidesert Moderate + Bock and others 1984

Sprague’s pipit N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

N. Plains Mixed grass Heavy - Kantrud and Kologiski 1982

Saskatchewan Mixed/short - Maher 1979

Alberta Mixed grass Heavy - Owens and Myres 1973

Botteri’s sparrow S. Arizona Semidesert Moderate - Webb and Bock 1990

Cassin’s sparrow S. Arizona Semidesert Moderate - Bock and Bock 1988

Brewer’s sparrow S. Arizona Semidesert Moderate + Bock and others 1984

Colorado Shortgrass Heavy - Ryder 1980

N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

N. Plains Mixed grass Heavy - Kantrud and Kologiski 1982

Vesper sparrow S. Arizona Semidesert Moderate 0 Bock and others 1984

Alberta Mixed grass Heavy + Owens and Myres 1973

N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

Lark sparrow S. Arizona Semidesert Moderate - Bock and others 1984

Black-throated sparrow S. Arizona Semidesert Moderate + Bock and others 1984

Lark bunting N. Texas Shortgrass Heavy - Wiens 1973

Colorado Shortgrass Heavy - Ryder 1980

N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

N. Plains Mixed grass Heavy 0 Kantrud and Kologiski 1982

Savannah sparrow Saskatchewan Mixed/short - Maher 1979

Alberta Mixed grass Heavy - Owen and Myres 1973

Baird’s sparrow Saskatchewan Mixed/short - Maher 1979

Alberta Mixed grass Heavy - Owen and Myres 1973

N. Plains Mixed grass Moderate, Heavy - Kantrud and Kologiski 1982

Grasshopper sparrow S. Arizona Semidesert Moderate - Bock and others 1984

Colorado Shortgrass Heavy - Ryder 1980

N. Texas Shortgrass Heavy - Wiens 1973

S. Dakota Mixed grass Heavy - Wiens 1973

N. Plains Mixed grass Moderate, Heavy - Kantrud and Kologiski 1982

Mccown’s longspur Colorado Shortgrass Heavy + Ryder 1980

Saskatchewan Mixed/short + Maher 1979

N. Plains Mixed grass Moderate, Heavy + Kantrud and Kologiski 1982

Chestnut-collared S. Arizona Semidesert Moderate + Bock and Bock 1988

 longspur Colorado Shortgrass Heavy 0 Ryder 1980

Saskatchewan Mixed/short + Maher 1979

Alberta Mixed grass Heavy - Owens and Myres 1973

N. Plains Mixed grass Moderate, Heavy 0 Kantrud and Kologiski 1982

Bobolink N. Plains Mixed grass Moderate, Heavy - Kantrud and Kologiski 1982

Eastern meadowlark S. Arizona Semidesert Moderate 0 Bock et al. 1984

Western meadowlark Colorado Shortgrass Heavy - Ryder 1980

N. Texas Shortgrass Heavy - Wiens 1973

Saskatchewan Mixed/short - Maher 1979

Alberta Mixed grass Heavy 0 Owens and Myres 1973

S. Dakota Mixed grass Heavy - Wiens 1973

N. Plains Mixed grass Moderate 0 Kantrud and Kologiski 1982

N. Plains Mixed grass Heavy - Kantrud and Kologiski 1982

Brown-headed cowbird N. Plains Mixed grass Moderate, heavy 0 Kantrud and Kologiski 1982

a

 Grazing intensity as reported by original authors; where more than one level of intensity is reported, both were tested separately and found to 

have the same effect. b

 Grazing effects on abundance: + = increase, - = decrease, 0 = no effect.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 83

that result from the creation of habitat edges in an 

environment where there formerly were none.

Habitat fragmentation has been most thoroughly 

studied for forest-dwelling birds (for example, Lovejoy 

and others 1986, Robinson and others 1995b), but we 

are slowly accumulating information on the impacts 

of fragmentation on grassland birds (Bock and others 

1999, Herkert 1994, Vickery and others 1994). Many 

grassland nesting species appear to have large area 

requirements, and simply are not found utilizing 

grasslands that are less than a particular area in extent; these are so called “area-sensitive” species. Such 

birds will not nest in habitat fragments below a certain 

threshold size, even if the fragment is large enough to 

hold several average-sized nesting territories and appears to be of suitable quality (Herkert 1994, O’Leary 

and Nyberg 2000, Samson 1980, Vickery and others 

1994). The minimum area requirements vary widely 

among species: eastern meadowlarks may require only 

5 ha, savannah sparrows are not found in fragments 

less than 40 ha, and greater prairie-chickens regularly 

breed in fragments of at least 160 ha (Herkert 1994, 

Samson 1980, Westemeier 1985).

Grassland birds also appear to avoid nesting close 

to habitat edges; such edges may be created by roads, 

treelines, fences, or urban development. Of the eight 

grassland nesting birds they studied, Bock and others (1999) found that five were significantly more 

abundant in interior plots (greater than 200 m from 

the suburban edge) than in edge plots adjacent to 

suburban areas, and two others were more abundant 

in the interior plots, although not significantly so. The 

five species that avoided nesting in the edge are all 

declining grassland species: vesper sparrow, savannah 

sparrow, grasshopper sparrow, bobolink, and western 

meadowlark. In contrast, the edge plots held nearly 

five times the abundance of common and/or exotic 

bird species: robins, starlings, grackles, house finches, 

and house sparrows (Bock and others 1999). This 

demonstrates a common trend witnessed along habitat 

edges: native grassland endemics have declined, while 

introduced and “weedy” or more cosmopolitan species have dramatically increased (Knopf 1994). As an 

example, Yahner (1983) found that of 47 bird species 

observed using shelterbelts in western Minnesota, only 

three were species that were typical of the historical 

grasslands found in that area.

In addition to opening up grasslands to potential 

invasion by opportunistic species, the creation of 

edges results in increased levels of predation and 

parasitism for nesting birds. The addition of trees, 

fences, telephone poles, and other vertical structures 

in a grassland landscape provide cover and perches 

for predators, leading to increased rates of predation 

along edges (Burger and others 1994, Gates and Gysel 

1978, Johnson and Temple 1990, Møller 1989, Ratti and 

Reese 1988, Winter and others 2000). Urban development leads to the introduction of associated “urban” 

predators, such as skunks, raccoons, and cats (Wilcove 

1985). Perches along edges also serve as “lookout 

sites” for brown-headed cowbirds, a critical habitat 

feature required by this nest parasite to successfully 

locate potential host nests in grasslands (Norman and 

Robertson 1975). Levels of cowbird parasitism are 

significantly greater for grassland birds nesting near 

edges, resulting in reduced nest productivity or even 

complete nest failure (Best 1978, Johnson and Temple 

1986, 1990, Wray and others 1982). Dickcissels, for 

example, are known to suffer negative consequences 

from brown-headed cowbird parasitism (Zimmerman 

1983). In areas of urban or surburban development, 

human disturbance may also play a role in the reduced 

density of grassland nesting birds in fragmented edge 

habitats (Knight and Gutzwiller 1995). Meadowlarks 

are an example of grassland breeding birds that are 

sensitive to disturbance at the nest (Lanyon 1995).

Haying and Mowing

The production and harvesting of hay is another 

land use that has increased with the development 

of the livestock industry. The harvesting of wild hay 

was once a common and economically viable practice 

in some Southwestern grasslands (Bahre 1987) but is 

now relatively limited in this region. A brief discussion 

of the effects of haying and mowing is included here 

to address those areas of the Southwest that may still 

engage in these practices.

As native grasslands have disappeared, many 

grassland birds have been forced to utilize “artificial” 

grassland habitats for nesting and for foraging; lands 

managed to produce forage for livestock are often 

attractive to grassland nesting birds (Sample 1989, 

as cited in Temple and others 1999). In hayfields 

that are mowed, birds are rarely successful in raising young. Birds are attracted to these hayfields and 

initiate nesting, but particularly due to trends toward 

earlier, more frequent mowing (Ratti and Scott 1991, 

Rodenhouse and others 1993, Ryan 1986), nests are 

destroyed or territories abandoned before the young 

birds fledge (Beintema and Muskens 1987, Bollinger 

and others 1990, Bryan and Best 1994, Frawley and 

Best 1991, Warner and Etter 1989). If forage crops are 

being raised, nests may be trampled or abandoned by 

the parents due to excessive disturbance when cattle 

densities are too high (Koerth and others 1983, Paine 

and others 1996). Pastures or hayfields thus often tend 

to act as population “sinks,” because breeding birds are 

attracted to them but are doomed to failure in their 

reproductive efforts (Temple and others 1999).

The impact of haying and mowing on grassland nesting birds is determined by the timing and frequency 
84 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

of such practices, and upon the habitat preference 

of the species in question. Bobolinks and savannah 

sparrows are negatively impacted in areas that are 

mowed (Bollinger and others 1990, Warner 1992), 

whereas those that prefer areas of short vegetation 

for nesting, such as horned larks or vesper sparrows, 

may increase as long as the mowing does not coincide 

with the nesting cycle (Laursen 1981).

Hayfields are typically cut at least once per year, with 

two to four cuts over the season being common. Harvests 

during May and June, the peak breeding season for 

most grassland nesting birds, are particularly harmful. 

In one study in New York, the production of bobolink 

fledglings in a hayfield was estimated to be reduced by 

40 percent by mowing practices (Bollinger and others 

1990). Mowing is increasingly being considered one of 

the most probable factors contributing to the decline 

of grassland nesting birds, at least in some areas 

(Bollinger and others 1990, Frawley and Best 1991).

Direct Mortality and Indirect Effects of 

Pesticides

As more of the nation’s grasslands are converted to 

agriculture, increasing numbers of our grassland birds 

that use agricultural fields are exposed to potentially 

harmful chemicals. Concerns over the possibly lethal 

effects of pesticides first came to light in the early 

1960s, when populations of birds such as brown pelicans 

and peregrine falcons were plummeting as a result of 

eggshell thinning caused by DDE, a byproduct of the 

popular organochlorine pesticide DDT (Anderson and 

Hickey 1972). Although the use of DDT was banned in 

the United States in 1972, as were most chlorinated 

insecticides, many of these compounds continue to 

be used in Latin America and South America, and 

elevated contaminant burdens continue to be found in 

North American birds that migrate to the neotropics 

for the winter (DeWeese and others 1986, Henny and 

others 1982, Johnston 1975, White and others 1981). 

In addition, “hot spots” of contamination persist. In 

New Mexico and Texas, for example, contamination 

from DDE residues continues to negatively impact 

reproduction in birds (White and Krynitsky 1986). 

Today most of the pesticides used in the United States 

are organophosphates and carbamates (Szmedra 1991). 

Despite the near extinctions of some species of birds 

from DDT, pesticide usage has increased dramatically 

since the 1960s. In 1964, 366 million lb of pesticides 

were used in agricultural applications in the United 

States; by 1997, that figure had more than doubled to 

770 million lb a year (Aspelin and Grube 1999). Table 

4-6 lists the leading pesticides causing avian mortalities in the United States.

Pesticides may affect birds in many ways. Acute 

mortality is, of course, the most obvious negative 

impact (Grue and others 1983). The pesticide carbofuran, for example, is estimated to kill between 1 

and 2 million birds a year in the United States (EPA 

1989). The toxicity of a pesticide may vary according 

to its form; birds have been found to be particularly 

susceptible to the ingestion of insecticide-treated seed 

and insecticide granules (Pimentel and others 1992). 

Direct mortality is often difficult to assess, however, 

as the bodies of birds that have died from poisoning 

or other causes are generally not found. Studies have 

shown that, in general, most bird carcasses (62 to 92 

percent) are scavenged and disappear within 24 hours 

of death (Balcomb 1986, Wobeser and Wobeser 1992). 

In addition, birds that have been poisoned often do not 

die immediately but will move from the site of poisoning to take cover (Vyas 1999 and references therein); 

therefore, the deaths—if noticed—are not necessarily 

associated with any particular pesticide application 

event. However, there are many documented incidents 

in which grassland birds have been found accidentally 

poisoned by pesticides (for example, Grue and others 

1983, Hill and Fleming 1982, McLeod 1967). Horned 

larks—one of the most common grassland birds in the 

United States—have been found killed by both carbamate pesticides (Stone 1979) and organophosphorous 

compounds (Beason 1995, DeWeese and others 1983). 

Although acute mortality has not been observed, some 

researchers have expressed concern over the possible 

negative impacts of the endangered aplomado falcon 

preying on bird species in the Lower Rio Grande 

Valley that have elevated levels of DDE and other 

contaminants (Mora and others 1997). And of course, 

the impacts of pesticides on migratory birds are not 

restricted to events that occur in the United States. 

From 1995 to 1996, nearly 6,000 wintering Swainson’s 

hawks perished in Argentina as the result of ingesting monocrotophos, an organophosphate insecticide 

used to control grasshoppers, a favorite prey item of 

the hawk (DiSilvestro 1996). In addition, some North 

American grassland birds are intentionally poisoned 

while on their wintering grounds in South America, 

where they are considered to be agricultural pests (for 

example, dickcissels; Basili and Temple 1999).

Far exceeding the direct effects of acute mortality 

from pesticide exposure are the sublethal effects that 

are more likely to contribute to long-term population 

declines in grassland birds and other species (Grue 

and others 1997). Avian exposure to organophosphates and carbamate insecticides have been found 

to produce a variety of physiological and behavioral 

deficiencies, including decreased body weight (Grue 

and Shipley 1984), lethargic behavior (Hart 1993), 

reduced territorial maintenance (Busby and others 

1990), decreased clutch size (Grue and others 1997 and 

references therein), reduced parental attentiveness 

(Busby and others 1990), decreased nestling growth 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 85

rates (Patnode and White 1991), increased postfledging 

mortality (Hooper and others 1990), reduced return 

rates to breeding grounds (Millikin and Smith 1990), 

and an inability to thermoregulate properly, resulting 

in hypothermia (Grue and others 1997). There are also 

more indirect effects on bird populations, such as a 

reduction in prey abundance for insectivorous species 

(DeWeese and others 1979, Hunter and Witham 1985, 

Moulding 1976).

Although pesticide usage on agricultural fields 

may be contributing to the observed declines in North 

American grassland birds, we do not currently have 

adequate data to say so definitively (Gard and Hooper 

1995). However, one estimate proposes that as many as 

67 million birds are killed by pesticides on farmlands 

in the United States every year (Pimentel and others 

1992); that estimate recently increased to 72 million 

annually (Pimentel 2001). Birds are more sensitive 

to contaminants than other wildlife (Grue and others 

1983), and as the breeding season for most birds is 

during the time of peak insect abundance to provide 

food for their chicks, it also unfortunately coincides 

with the time of greatest pesticide usage (Gard and 

Hooper 1995). Some crops, such as apples and cotton, 

may be sprayed as many as 20 times per growing season 

(Pimentel and others 1991). From these facts and the 

known negative impacts of pesticides on birds, both 

in terms of acute mortality and long-term sublethal 

effects, we can infer that grassland birds nesting in 

agricultural areas are at high risk. More research 

is badly needed to more exactly assess any possible 

link between agricultural pesticide use and declines 

in grassland birds.

Loss of Prairie Dog Colonies

The prairie dog is considered a “keystone” species, 

a species that has a large overall effect on a community or ecosystem disproportionate to its abundance 

(Kotliar and others 1999, Power and others 1996). The 

activities of these burrowing animals have a dramatic 

impact on the patch dynamics and ecosystem function 

of the Western grasslands that they inhabit. Prairie 

dog disturbances impact the physical and chemical 

properties of the soil, alter vegetational structure, affect 

plant species composition, and improve the nutrient 

value of plants growing in the vicinity of their colonies 

(O’Meilia and others 1982 and references therein, 

Whicker and Detling 1988). Active prairie dog towns 

contribute to increased biological diversity by supporting a different complement of species compared to 

areas unoccupied by prairie dogs (Agnew and others 

1986, Mellink and Madrigal 1993, O’Meilia and others 

1982). Furthermore, several vertebrate species are 

considered highly dependent upon prairie dogs either 

as prey or for the habitat provided by their colonies, 

including the endangered black-footed ferret Mustela 

nigripes (see Kotliar and others 1999 for an excellent 

review).

Once a dominant force in the grasslands of the 

Western United States, the ecological impact of the 

prairie dog on these systems has nearly been extinguished. Up until the early 1900s, prairie dog colonies 

were estimated to cover hundreds of millions of acres 

Table 4-6. Leading pesticides causing avian mortalities in the United States. Fifty-five percent of all incidents reported are associated with two insecticides, Carbofuran and Diazinon. Based on the Ecological Incident Information System (EIIS) as of 1999, 

as reported by the American Bird Conservancy (2001). See also Mastrota (1999).

# of avian

Active incidents in EIIS # of

ingredient Trade name Total Probablea carcasses Common uses Notes

Carbofuran Furadan 352 241 12,341 Grapes, corn, Most granular uses cancelled in 

 alfalfa 1991

Diazinon Diazinon, Spectracide 267 165 4,434 Lawns and turf Uses cancelled on golf courses 

 in 1989

Chlordane Chlordane, Termide 70 Termiticide Cancelled in 1987

Fenthion Baytex, Baycide 58 47 5,545 Avicide, mosquito Use as avicide cancelled in 

 control 1999

Chlorpyrifos Dursban, Lorsban 57 37 Termiticide, lawn

 and turf

Brodifacoum Havoc, Talon 47 47 Rodent control

Parathion Parathion, Folidol 45 2,457 Small grains,

 sunflower, alfalfa

Famphur Warbex 31 Livestock

a

 includes those incidents that were convincingly linked to pesticide use, and not linked to pesticide misuse.
86 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

of shortgrass prairie and desert grasslands west of the 

Great Plains (Anderson and others 1986). Today prairie 

dogs are estimated to persist on a mere 2 percent of 

their former range (Anderson and others 1986, Miller 

and others 1994). One species, the Utah prairie dog 

(Cynomys parvidens) is listed as threatened, and the 

black-tailed prairie dog C. ludovicianus, formerly the 

most abundant and widespread of the five species of 

prairie dogs in North America, has been considered 

for listing (USFWS 2000). The population numbers of 

the black-tailed prairie dog are estimated to have been 

reduced by 98 percent, and the species may occupy 

as little as 0.5 percent of its original range (Mac and 

others 1998, as cited in USFWS 2000). Although the 

conversion of native prairie habitat to other land uses 

may have contributed to some degree, undoubtedly the 

greatest single factor in the loss of prairie dogs has 

been the concerted effort by both Federal and State 

government agencies to exterminate these animals 

for the benefit of the livestock industry (Mulhern and 

Knowles 1996, Parmenter and Van Devender 1995).

The campaign to eradicate prairie dogs from Western 

grasslands began in earnest following the release of a 

Department of Agriculture report suggesting that the 

presence of prairie dogs may reduce range productivity 

by 50 to 75 percent (Merriam 1902). The U.S. Biological 

Survey responded with a massive poisoning campaign 

under the auspices of its Predator and Rodent Control 

program. Aiming to reduce competition with livestock, 

millions of acres of prairie dog colonies were poisoned, 

and shooting of prairie dogs was encouraged across 

their range (Bell 1921, Mulhern and Knowles 1996, 

Parmenter and Van Devender 1995, Van Pelt 1999). 

Fear of sylvatic plague buoyed these efforts after the 

bacterium was discovered in black-tailed prairie dogs in 

Texas in the 1940s (Cully 1989, Mulhern and Knowles 

1996). In some States, annual extermination of prairie 

dogs on State and privately owned lands was a legal 

requirement. Nebraska, for example, only recently 

repealed this mandate in 1995 (Van Pelt 1999).

The black-tailed prairie dog and the Gunnison’s 

prairie dog C. gunnisoni are the two species that 

inhabit the grasslands of Arizona and New Mexico. 

Described as occurring in “immense colonies” in Arizona 

in 1885 (Mearns 1907), the Arizona prairie dog C.l. 

arizonensis, a subspecies of the black-tailed prairie 

dog, was largely extirpated from that State by 1938; 

one small single colony survived until 1960 (Van Pelt 

1999). In New Mexico, the range of the black-tailed 

prairie dog has been reduced by at least 25 percent 

(Hubbard and Schmitt 1984). In the Animas Valley, for 

example, biologists from the Museum of Southwestern 

Biology did not observe one single prairie dog between 

the years 1955 and 1972 (Findley 1987). Yet in 1908 

Vernon Bailey had described this same area as an 

almost continuous prairie dog town for its length and 

breadth, estimating that over 6 million prairie dogs 

inhabited the valley (Bailey 1932).

As a keystone species, the elimination of the prairie 

dog has had far-reaching repercussions, including 

declines in several species of grassland birds. In a 

critical review of the evidence, Kotliar and others 

(1999) propose that nine vertebrate species may be 

considered to be truly dependent upon prairie dogs. 

Five of these nine species are grassland birds: the 

mountain plover, burrowing owl, ferruginous hawk, 

golden eagle, and horned lark (the others are the 

black-footed ferret, swift fox Vulpes velox, deer mouse 

Peromyscus maniculatus, and northern grasshopper 

mouse Onychomys leucogaster). For all of these species 

except the horned lark, evidence shows that local or 

landscape-level declines have closely corresponded 

with declines in prairie dog populations (Kotliar and 

others 1999 and references therein). For these birds, 

prairie dogs may serve as primary prey items (ferruginous hawks and golden eagles) or their activities may 

provide critical habitat (nest burrows for burrowing 

owls, preferred areas of short vegetation and/or bare 

ground for mountain plovers and horned larks). The 

decline of prairie dog populations was one of the key 

factors mentioned in the proposal to list the mountain 

plover as a threatened species (USFWS 1999a) and has 

been clearly linked with local declines in burrowing 

owls and ferruginous hawks (Cully 1991, Desmond 

and others 2000).

Many other species of grassland birds have been 

reported in association with prairie dog colonies as 

well. In the largest complex of prairie dog colonies 

remaining in Mexico, Manzano-Fischer and others 

(1999) report that 31 percent of the birds observed on 

the colonies are grassland specialists; this includes 

a new record for wintering mountain plovers in the 

State of Chihuahua. Numerous avian species have 

been recorded as prairie dog associates, well beyond 

the five birds named earlier (for example, Agnew and 

others 1986, Sharps and Uresk 1990), but at present 

there is not enough evidence to determine the degree 

of their dependence and the potential impact of the 

elimination of prairie dogs on their populations (Kotliar 

and others 1999).

Notwithstanding the drastic declines already witnessed in prairie dog numbers and the evidence of a 

cascade effect on other species, prairie dogs today are 

still widely considered to be vermin and enjoy little in 

the way of legal safeguards (Van Pelt 1999). This is in 

spite of more recent evidence that the level of competition between prairie dogs and livestock is more likely 

on the order of 4 to 7 percent (Uresk and Paulson 1988, 

as cited in Miller and others 1994), and that there is 

no significant difference in the market weight of steers 

whether they graze in conjunction with prairie dogs 

or not (O’Meilia and others 1982), although it should 

be noted that the statistically insignificant weight 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 87

difference did result in an economic loss. Forage on prairie dog colonies is of higher quality, and the increased 

nutritional value of herbage on colonies may act to 

offset any decrease in biomass as a result of clipping 

by prairie dogs (Holland and Detling 1990, O’Meilia 

and others 1982 and references therein). Bison, elk, 

pronghorn, and livestock preferentially graze on prairie 

dog colonies, presumably because of the increased value 

and palatability of the herbage there (Coppock and others 1983, Knowles 1986, Krueger 1986, Wydeven and 

Dahlgren 1985). Prairie dogs also play an important 

role in inhibiting the growth of woody invaders such 

as mesquite Prosopis spp. (Koford 1958, Weltzin and 

others 1997). The elimination of prairie dogs may be 

partially responsible for the widespread encroachment 

of mesquite into Southwestern grasslands in recent 

years (Parmenter and Van Devender 1995).

Following their comprehensive review, Kotliar and 

others (1999:186) conclude prairie dogs are “crucial to 

the structure and function of native prairie systems.” 

Further elimination of prairie dog colonies will almost 

certainly have negative effects on many species of 

grassland birds as well as grassland ecosystems as a 

whole in the Southwest. Not only are scientists today 

stressing the importance of preserving remaining prairie dog colonies (for example, Miller and others 1994), 

some are going further and calling for the reintroduction of prairie dogs to restore ecosystem function (for 

example, Manzano-Fischer and others 1999).

Cowbird Parasitism

Cowbirds of the genus Molothrus are brood parasites; 

these birds do not raise their own young but instead lay 

their eggs in the nests of host species. The brown-headed 

cowbird Molothrus ater is the species commonly found 

throughout the United States, although recently the 

bronzed cowbird M. aeneus has begun expanding its 

range northward from Mexico into Texas, New Mexico, 

and Arizona. As grassland birds are one of the guilds 

least likely to be impacted by cowbird parasitism, 

this discussion will be relatively brief and will focus 

on the brown-headed cowbird, as we have the most 

information on that species.

The brown-headed cowbird is native to the shortgrass prairie west of the Mississippi (Friedmann 1929, 

Mayfield 1965). This species was originally known as 

the “buffalo bird” because it was found in association 

with the great herds of bison on the plains. Cowbirds 

forage for grain in short grass and on bare ground, and 

they follow herd animals to catch the insects stirred 

up by their motion through the grass. Several aspects 

to the life history of the brown-headed cowbird (hereafter just “cowbird”) make this species a particularly 

formidable threat to many bird species. The cowbird 

is a generalist parasite; that is, it does not specialize 

on any particular host species. Cowbirds are now 

known to parasitize at least 240 species of birds in 

North America (Friedmann and Kiff 1985). In addition, 

female cowbirds may lay from 30 to 40 eggs a year; 

some females in captivity have produced nearly 80 

eggs in a single season (Robinson and others 1995b), 

thus giving each individual the capacity to parasitize 

numerous host nests. Cowbirds are not restricted by 

specific habitat needs because they can utilize foraging 

areas quite distant and different from the areas that 

they use to search for host nests, allowing them to be 

found in a tremendous range of breeding habitats. 

Finally, cowbirds are highly successful by producing 

eggs with a short incubation period. Their eggs will 

hatch before that of the host species, and as cowbirds 

generally parasitize species smaller than themselves 

and cowbird nestlings have extremely high growth 

rates, the young cowbirds soon outcompete their 

nestmates, leading to decreased success of the host 

or even complete nest failure (Robinson and others 

1995b). Additionally, most females will remove a host 

egg before laying their own, further decreasing the 

reproductive success of the host (Friedmann 1963, 

Sealy 1992, Weatherhead 1989). If there are two or 

more cowbird eggs in a nest, even a species that might 

normally be capable of raising one cowbird chick in 

addition to its own will most likely fail to raise any 

of its own nestlings to maturity (Robinson and others 

1995b). High levels of cowbird parasitism have been 

associated with the eradication of entire populations 

of some bird species, such as the black-capped vireo 

(Gryzbowski 1991).

Birds that evolved in the historical range of the 

cowbird in the Midwest have for the most part developed the ability to recognize and eject cowbird eggs, 

or to abandon parasitized nests (Rothstein 1975, 

1977, 1982, Scott 1977). Other birds, such as the 

Southwestern willow flycatcher, have not (Robinson 

and others 1995b). Cowbirds have become a problem 

because human alteration of the environment has 

allowed them to vastly expand their range, giving 

them access to any number of naïve host species. The 

combination of human settlements, clearing of forests, 

and introduction of livestock has allowed the cowbird 

population in North America to virtually explode since 

the early 1900s (fig. 4-4).

Although grassland birds face many challenges, 

parasitism by cowbirds is currently one of the lesser 

threats. In general, ground-nesting grassland species are only rarely parasitized. Out of 98 neotropical 

migratory birds, Robinson and others (1995a) classify 

only two grassland nesting species as common cowbird 

hosts and five as uncommon hosts. The highest rates 

of nest parasitism occur in areas with an open subcanopy, which presumably provides a vantage point 

for female cowbirds searching for nests (Robinson and 

others 1995b). Because true grassland habitats lack 
88 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Figure 4-4. The percent of Audubon Christmas bird counts 

reporting cowbirds, from 1900 to 1980 (r = 0.96, p < 0.01) (from 

Brittingham and Temple 1983).

such potential perches, cowbirds are hampered by the 

inability to locate and monitor host nests on the ground. 

However, this fact also points to the importance of 

minimizing edge habitats in grasslands and illustrates 

how the introduction of vertical structures such as 

fences or telephone poles can alter this situation. If 

woody corridors or edges are available within a grassland habitat, rates of cowbird parasitism on grassland 

birds nesting close to that edge increase dramatically 

(Johnson and Temple 1990). To keep levels of cowbird 

parasitism on grassland birds low, the creation of woody 

edge habitats such as treelines or the establishment 

of livestock corrals and other structures must be kept 

to an absolute minimum within the remaining native 

grassland habitat.

Management for Grassland Birds in 

the Southwest______________

General Recommendations and Resources

Most of the published literature on managing 

for grassland birds has focused on the prairies of 

the Midwest and therefore has dealt primarily with 

birds inhabiting mixed-grass and tallgrass habitats. 

Unfortunately there has been little guidance available 

for managing birds that inhabit the more arid 

grasslands of the Southwest, but a few recent 

publications should serve as excellent resources 

for land managers in this region. For those working primarily in shortgrass steppe, the Rocky 

Mountain Bird Observatory has published a 

landowner’s guide to best management practices 

for birds of the shortgrass prairie (Gillihan 

and others 2001); the Observatory also hosts 

a website that provides numerous resources 

for conservation of shortgrass prairie birds: 

http://www.rmbo.org. The Bird Conservation 

Plans produced by Partners in Flight of both 

Arizona and New Mexico provide management 

recommendations for all of the major habitat 

types in these States, and the plans should be 

a highly valuable resource for land managers 

in these States (Latta and others 1999, NMPIF 

2001). Much of what follows in this introduction is based upon the information from these 

three sources. In addition, Partners in Flight 

has recently constructed a Web site that is an 

outstanding resource for best management 

practices to benefit birds in a variety of ecosystems, including grasslands. It is available online at 

http://www.partnersinflight.org/pubs/BMPs.htm.

In general, it is important to keep in mind that 

desert grasslands are even more arid and fragile than 

shortgrass prairie, and that desert grasslands are 

much more sensitive to livestock pressures because 

they did not evolve in concert with large numbers of 

native grazers, as did the flora of the shortgrass prairie. 

Desert grasslands should therefore be treated more 

conservatively than the grasslands of the shortgrass 

steppe. Management practices that have been developed based on climatic conditions in other regions must 

be adapted according to the local soil, precipitation, 

and plant composition characteristics, with careful, 

ongoing monitoring of management practices to ensure 

the continued viability of the system. Fire, for example, 

is a common management tool in grassland systems. 

However, as already noted, historically fires were more 

infrequent in desert grasslands than in mixed-grass or 

tallgrass prairies and may have negative impacts on 

grasslands composed primarily of black grama (see also 

chapter 3 of this volume). The U.S. Geological Survey in 

Tucson, AZ, the U.S. Natural Resources Conservation 

Service, U.S. Bureau of Land Management, and 

land stewards for The Nature Conservancy are good 

resources for information on the recommended frequency and timing of fires for the grasslands in your 

area.

In managing for grassland birds, it is necessary to 

first determine which species are currently present on 

the land in question, as well as which species could 

potentially inhabit the property if the right conditions 

1900 1920 1940 1960 1980

Year (1900-1980)

80

60

40

20

0

% Counts Reporting Cowbirds

• •• ••

•••

••

• •• ••

• •

•

••

•

••

• • • ••

•

• ••

• •

• • •

• • •

• • •

• •

•

• • •

• •

•

• • • •

• • • •

•• • •

• ••••••• ••• • • •• • ••
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 89

were available. For assistance in this regard, wildlife 

biologists from the USDA Forest Service and USDI 

Fish and Wildlife Service should be able to help, and 

your local representatives of Partners in Flight or 

the National Audubon Society are also good sources 

of information.

A common confounding factor is that management 

practices for one species may conflict with those of 

another species that is present or could potentially occur in the area. For example, management to improve 

habitat for Sprague’s pipits, which avoid woody vegetation, would be at odds with management for Cassin’s 

sparrows or loggerhead shrikes, both of which require 

some degree of shrub cover. In these cases a choice must 

be made. If the property is extensive enough to divide 

into large separate management units, one may choose 

to manage for both species separately on different parts 

of the landscape. In these circumstances, researching 

the minimum area requirements for the species in 

question is important, if such information is available. 

Alternatively, one could opt to manage the land for the 

species that is of greater conservation concern in the 

area (Gillihan and Hutchings 2000). Again, the Arizona 

and New Mexico Partners in Flight Bird Conservation 

Plans are good resources for this information, as the 

plans clearly outline groups of bird species that will 

benefit from similar management practices and highlight those species of conservation concern.

Some general management recommendations for 

grassland habitats of the Southwestern United States 

include:

• Whenever possible, preserve native grasslands 

and the plant and animal species found there; 

expansive areas of contiguous grasslands are 

especially important for conservation.

• As a general rule, managers should strive to maintain grass canopy at 50 to 65 percent coverage and 

keep areas of bare ground to a minimum. Shrub 

cover should not exceed 20 percent. In desert 

grasslands, tall shrubs (including yuccas) of 3 

m (9 ft) or more in height should be encouraged 

for use as nest sites.

• Restore and/or preserve natural water sources 

in grassland systems, such as cienegas, springs, 

playas, lakes, and so forth. For wet meadows, 

restore areas drained for agriculture.

• Minimize use of alpine tundra areas for recreation 

and livestock use to protect these fragile systems 

from trampling and erosion.

• Determine appropriate grazing systems for the 

land according to local climatic conditions and the 

type of habitat preferred by birds of management 

concern in your area. Appropriate systems may 

include deferred rotation or rest rotation; continuous grazing is not recommended. In the Midwest, 

a “bird friendly” rotation system was developed 

that includes setting aside approximately onethird of the total pasture (a minimum of 8 ha) as 

a “refuge” for grassland nesting birds. This area 

is excluded from grazing and other disturbances 

for 6 weeks during the peak breeding season, to 

allow birds time to complete their nesting cycle 

(Bartelt 1997).

• Set aside large areas of land from grazing. 

Permanent grazing exclosures of at least 1,000 ha 

are recommended to create a mosaic of habitats 

within more heavily grazed areas and to provide 

appropriate habitat for grassland birds that are 

intolerant of grazing (Saab and others 1995). If 

permanent exclusion is not possible, the area 

should be rested from grazing for a minimum of 

25 to 50 years (Bock and others 1993).

• Limit forage utilization to 35 to 40 percent in 

years of good rainfall and reduce stocking rates 

in drought years (Paulesen and Ares 1962, as 

cited in Loftin and others 2000).

• Manage grassland parcels as large units rather 

than as many small ones, as many grassland bird 

species exhibit a preference for extensive areas 

of grasslands.

• Schedule management activities (haying, burning, and such) to avoid the breeding season (early 

spring through mid-July for most species). High 

stocking rates should also be avoided during this 

time, when nests of ground-nesting birds have a 

high likelihood of being trampled.

• Mow or burn uncultivated areas in rotation, to 

create a patchwork of habitat types available 

for birds (working in large units of land). Use 

a flush bar on mowing devices and mow slowly 

in a back-and-forth pattern to give birds time to 

escape (see Gillihan and others 2001).

• If using fire as a management tool, research the 

natural (that is, historical) frequency of fires for 

the area in question and time burns to avoid the 

nesting and fledging times of local grassland 

birds.

• Use Integrated Pest Management Practices whenever possible; if chemical controls are necessary, 

use those that degrade rapidly.

• Preserve and/or restore colonies of small rodents, 

such as prairie dogs, that serve as important 

food sources for species of grassland birds such 

as golden eagles and ferruginous hawks and 

provide habitat for numerous other species such 

as burrowing owls and mountain plovers.

• Restore and reseed grasslands with native species of grasses.

• In all cases, bird populations should be monitored 

and management practices altered accordingly, 

if need be, to maintain or increase populations 

of the target species.
90 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

A good starting point for gathering information on 

many general grassland management issues in the 

Southwest is The Future of Arid Grasslands: Identifying 

Issues, Seeking Solutions, available from the USDA 

Forest Service (Tellman and others 1998).

Management Reviews for Select Species of 

Grassland Birds

This section contains some brief reviews and specific 

management recommendations for selected representative species of Southwestern grassland birds. These 

particular species represent various examples of the 

broad variety of birds typical of the grasslands of the 

Southwest, including both resident and migratory 

species and birds of various ecological guilds and 

grassland habitat preferences, with an emphasis on 

species believed to be in decline. These summaries are 

not intended to represent an exhaustive accounting 

of all Southwestern grassland birds; such a review 

is beyond the scope of this document. Furthermore, 

some typical grassland species of the Southwest with 

limited distributions, such as the northern aplomado 

falcon and masked bobwhite, have not been included. 

Although these species certainly rank highly in terms 

of conservation concern, both being endangered, they 

were excluded from the discussion here because of 

the limited relevance of the information for most 

land managers in the region (northern aplomado 

falcons having been extirpated from New Mexico and 

Arizona, and the masked bobwhite occurring only on 

the Buenos Aires National Wildlife Refuge in Arizona). 

In addition, these species already have comprehensive 

recovery plans in place, managers can access through 

the USDI Fish and Wildlife Service’s Environmental 

Conservation Online System at http://ecos.fws.gov.

Additional management recommendations for 

other grassland bird species not covered here may 

be found in the Partners in Flight Bird Conservation 

Plans for New Mexico and Arizona and the Rocky 

Mountain Bird Observatory’s publication Sharing 

Your Land with Shortgrass Prairie Birds, referred to 

in the preceding section. Another potential resource is 

Effects of Management Practices on Grassland Birds, 

a series of species accounts compiled by the Northern 

Prairie Wildlife Research Center (Johnson and 

Igl 2001). Although their emphasis is on grassland 

birds of the northern Great Plains, management 

recommendations for many species that also occur in 

the Southwest may be found in these comprehensive 

accounts available online at http://www.npwrc.usgs.

gov/resource/literatr/grasbird/grasbird.htm. See also 

the best management practices for grassland birds 

available from the Partners in Flight Web site at 

http://www.partnersinflight.org/pubs/BMPs.htm.

All of the species accounts that follow end with a list 

of associated species. The birds listed are those that 

may typically be found in habitats similar to that of the 

species detailed and that may also benefit from some 

of the recommended habitat management practices. 

However, not all of the species listed will necessarily 

respond in the same manner. If management for a 

particular group of species is the goal, the manager 

should further research the recommendations for each 

of those species individually. In the interest of saving 

space, multiple references that are cited within a single 

review document are cited as “and references therein” 

rather than listing each citation individually.

Ferruginous Hawk Buteo regalis

Distribution and Population Trends—During 

the breeding season, ferruginous hawks are found in 

appropriate grassland or shrubland habitats across the 

Intermountain West and Western Great Plains from 

Southwestern Canada south to northern Arizona, New 

Mexico, and Texas (fig. 4-5). In the central portion of 

its range, including Colorado, eastern Utah, northern 

Arizona, and New Mexico, this hawk is a year-round 

resident. In New Mexico, the ferruginous hawk is a 

regular resident on the Eastern Plains and the Plains 

of San Agustin; migratory individuals and wintering 

populations are also found across the State (Hubbard 

1978). In Arizona, the ferruginous hawk is a year-round 

resident in the northern half of the State, with many 

migratory birds moving into the southern part of the 

State to winter (Glinski 1998).

In the winter, the more northerly breeding populations migrate southward. Wintering ferruginous hawks 

are most common in California, Colorado, Arizona, 

New Mexico, and northern Texas; smaller numbers are 

found throughout their range from southern Wyoming 

and Nebraska southward into Baja California and 

north-central Mexico (fig. 4-6).

For the most part, this species appears to still occupy 

most of its historic range; the primary exception is a 

reported range contraction in south-central Canada in 

the early 1900s due to conversion of native grassland 

habitats to agricultural uses (Bechard and Schmutz 

1995). Numbers of wintering birds are reported to have 

increased in California and in the eastern portion of the 

hawk’s range since the 1980s due to loss of wintering 

habitat in the Great Plains (Bechard and Schmutz 

1995). Bechard and Schmutz (1995, and references 

therein) state that the ferruginous hawk is considered 

to be declining in several areas, and recent declines 

have been documented in northern Utah and eastern 

Nevada (Olendorff 1993). Populations in Arizona 

also appear to be decreasing in size as well as range 

(Glinski 1998).

There is much conflicting information on the 

population status of this species. A petition to list 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 91

the ferruginous hawk under the Federal Endangered 

Species Act in 1991 (Ure and others 1991) was denied 

(USFWS 1992). In contrast to the declines proposed by 

several researchers (cited above), Peterjohn and Sauer 

(1999), in their analysis of Breeding Bird Survey (BBS) 

data from 1966 to 1999, reported that the ferruginous 

hawk was one of only three species of grassland birds 

demonstrating a positive population trend. During 

this period, BBS data reflects an annual increase in 

ferruginous hawk numbers of 5.2 percent (p < 0.01, n 

= 186); the rate of increase was even greater in more 

recent years, demonstrating a 7.2 percent annual 

increase from 1980 to 1996 (0.01 < p < 0.05, n = 170). 

Some of the inconsistencies come, no doubt, from the 

small numbers of raptors that tend to be detected on 

BBS routes and the resulting tenuous validity of any 

Figure 4-6. Winter distribution and densities 

of ferruginous hawks, 

based on Christmas Bird 

Count data. Counts are 

average number of birds 

detected per survey over 

the years 1982-1996 

(from Sauer and others 

2001).

Figure 4-5. Distribution and densities of breeding ferruginous hawks 

in the United States and Canada, 

as mean numbers of individuals 

detected per route per year. Data 

averaged from Breeding Bird Surveys over the years 1982 to 1996 

(Sauer and others 2001).

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and above

None counted

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100
92 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

statistical analysis attempted to identify trends based 

on that data (Sauer and others 2001). Furthermore, 

ferruginous hawk populations are known to fluctuate 

in concert with cycles in prey abundance (Dechant 

and others 2001a and references therein), so it may 

be natural for their numbers to oscillate over time 

(Bechard and Schmutz 1995). Regardless of population 

fluctuations, the continental population of ferruginous 

hawks is relatively small and therefore worthy of 

conservation concern; a 1993 estimate placed the 

number of individuals in North America at somewhere 

between 6,000 and 11,000 birds (Olendorff 1993). The 

ferruginous hawk is a USFWS Migratory Nongame Bird 

of Management Concern, is designated as a Highest 

Priority species by New Mexico Partners in Flight, and 

is a priority species in the Arizona Bird Conservation 

Plan. The ferruginous hawk is considered Threatened 

by the State of Arizona (table 4-1).

Preferred Habitat and Nest Site 

Characteristics—Preferred breeding habitat includes flat, rolling grasslands, deserts, and shrubsteppe 

regions; ferruginous hawks avoid high elevations, 

interior forests, and narrow canyons (Bechard and 

Schmutz 1995). Ferruginous hawks will also use 

woodland edges (Olendorff 1993). Although they utilize 

pastures and croplands to some extent for nesting and 

foraging, these hawks tend to avoid areas where more 

than 50 percent of the landscape has been converted to 

such uses (Dechant and others 2001a and references 

therein). They are also sensitive to human disturbance 

and prefer to nest far from human activities (Dechant 

and others 2001a and references therein).

One of the key criteria for breeding habitat is an 

abundance of prey. Ferruginous hawks specialize on 

small mammals, and their diet is largely based on prairie 

dogs (Cynonmys spp.), ground squirrels (Spermophilus

spp.), and rabbits (Lepus spp.) (Bechard and Schmutz 

1995 and references therein). These hawks will avoid 

areas of dense vegetation that interfere with their 

ability to see prey (Dechant and others 2001a).

Winter habitat is similar to that used in the breeding 

season—extensive, open areas of grassland, desert, or 

shrubland—except that in the winter there is an even 

greater dependence of this hawk upon prairie dogs and 

ground squirrels, and they tend to congregate in the 

vicinity of prairie dog towns (Bechard and Schmutz 

1995, Berry and others 1998, Plumpton and Andersen 

1998). In Arizona and New Mexico, pocket gophers 

(Thomomys spp.) are also important prey (Bechard 

and Schmutz 1995). Wintering ferruginous hawks in 

Colorado avoided landscapes that were more than 5 to 

7 percent urbanized (Berry and others 1998).

Ferruginous hawk nests are large, bulky structures that may be added to and used year after year 

(Bechard and Schmutz 1995). Most often ferruginous 

hawks will nest in a lone tree or artificial structure, 

such as a utility pole, but they will also nest directly 

on the ground (Bechard and Schmutz 1995). Trees 

appear to be the preferred site if they are available 

(Bent 1937, Bechard and Schmutz 1995). In a sample 

of more than 2,000 ferruginous hawk nests, Olendorff 

(1993) found that 49 percent were placed in trees or 

shrubs, 21 percent were on cliffs, 12 percent on utility 

structures, and 10 percent on ground outcrops. When 

nesting on the ground, nests are usually placed on a 

slope, knoll, ridgecrest, or rock pinnacle rather than 

on level ground (Palmer 1988). In New Mexico, nests 

are often found near the edge of open grasslands in 

juniper savannas or pinyon-juniper woodlands (Hawks 

Aloft 1998).

Timing of Breeding and Migration—In the 

Southwest, breeding ferruginous hawks are believed to 

be resident year-round (Bechard and Schmutz 1995). 

Wintering individuals from more northerly breeding populations begin migrating south from August 

through October (Schmutz and Fyfe 1987), and return 

to their breeding grounds from February through April 

(Lokemoen and Duebbert 1976).

Breeding pairs begin forming in February or March, 

and in the Southwest most nest building activity occurs 

in March (Bechard and Schmutz 1995). Ferruginous 

hawks generally raise only a single brood a year 

(Bechard and Schmutz 1995). Incubation is estimated 

to be 32 to 33 days (Palmer 1988), and the young hawks 

typically leave the nest between 38 and 50 days after 

hatching (Bechard and Schmutz 1995). Ferruginous 

hawks are highly sensitive to disturbance during the 

breeding season and are particularly likely to abandon 

nests in the early stages of nesting, during the incubation period (Dechant and others 2001a and references 

therein). Reproductive success is closely linked with 

the local abundance of major prey species (Schmutz 

and Hungle 1989, Woffinden and Murphy 1989).

Area Requirements— In a study of several 

breeding areas, the average nearest neighbor distance between ferruginous hawk nests was 13.4 km 

(Olendorff 1993). Home range sizes vary widely, from 

3 to 8 km2 in the Great Basin (Janes 1985) up to 90 

km2 in Washington (Leary and others 1998). In Idaho, 

Wakely (1978) estimated that approximately 22 km2

may be required to support one pair of ferruginous 

hawks. Wintering ferruginous hawks in Colorado used 

average home ranges of between 2.3 km2 and 4.7 km2; 

habitat suitability was highly dependent upon the 

presence of black-tailed prairie dog colonies (Plumpton 

and Andersen 1998).

Management Issues and Recommendations—

Ferruginous hawks will avoid areas that have been 

largely converted to agriculture but appear to be 

relatively tolerant of grazing (for example, Kantrud 

and Kologiski 1982, Wakely 1978). A key management 

recommendation is therefore the protection of large 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 93

tracts of native grassland from conversion to cropland 

or urbanization (Dechant and others 2001a).

Ranching operations appear to be largely consistent with ferruginous hawk conservation (Bechard 

and Schmutz 1995). The provision of abundant prey 

populations is also a critical consideration for this 

hawk (Dechant and others 2001a). Any land uses 

should be managed so as to maintain large numbers 

of small mammals, particularly ground squirrels and 

prairie dogs. Reintroduction of prairie dogs into areas 

formerly occupied will also benefit the ferruginous 

hawk, as well as other bird species that occur in association with prairie dogs (for example, burrowing 

owls; Latta and others 1999). Although it has been 

proposed that grazing may benefit ferruginous hawks 

by decreasing vegetative cover and thereby increasing 

the visibility of potential prey (Wakely 1978), Bock 

and others (1993) warn that reduction of ground 

cover in shrubsteppe habitats may result in decreased 

numbers of small mammals, therefore lowering the 

quality of the habitat for foraging ferruginous hawks. 

Increasing quality habitat for small mammals may 

also be achieved through creating a mosaic of habitats 

when converting land; for example, by leaving patches 

of untreated areas when clearing sagebrush steppe 

for conversion to grassland (Howard and Wolfe 1976). 

Attempts to eliminate populations of small mammals 

will negatively impact ferruginous hawks by removing 

their prey base. It is recommended that such control 

be exercised only to reduce high points in the cyclic 

populations of small mammals rather than trying to 

eliminate them altogether (Olendorff 1993). In addition, use of poisons such as strychnine on rodents may 

secondarily poison ferruginous hawks (Bechard and 

Schmutz 1995).

The provision of suitable nesting substrates is often 

successful in encouraging the reoccupation of formerly 

utilized areas, such as where woodlands have been 

converted to grasslands or natural nest areas have 

been lost through other means. This may be achieved 

by leaving individual trees or patches of trees when 

converting wooded areas to grasslands (Olendorff 

1993), by protecting existing trees or nest structures 

such as old utility poles through fencing or by other 

means (Olendorff 1993), or by the provision of artificial 

nest platforms. Schmutz and others (1984) describe 

platform designs that have been used successfully in 

Canada. Scattered trees and utility poles may also be 

useful as hunting perches.

One of the most important management tools for 

this species is to avoid disturbance at the nest sites 

from roughly March through July or August (Bechard 

and Schmutz 1995, Howard and Wolfe 1976). As these 

hawks are most likely to abandon their nests during 

the incubation period, protection of these nests from 

March through May is of the utmost importance. It is 

recommended that any treatments of the land, such as 

plowing, burning, chaining, or discing, be performed 

during the nonnesting season (Olendorff 1993). 

Olendorff (1993) suggests creating buffer zones of 0.25 

km radius around active ferruginous hawk nests for any 

brief disturbances, 0.5 km for intermittent activities, 

0.8 km for more prolonged activities, and approximately 

1.0 km for construction or other activities that will 

continue over several months or more.

Associated Species—Golden eagle, northern 

harrier, American kestrel, prairie falcon, scaled quail, 

mountain plover, burrowing owl (if burrows present), 

mourning dove, common nighthawk, ladder-backed 

woodpecker, Say’s phoebe, horned lark, loggerhead 

shrike, Bendire’s thrasher, vesper sparrow, lark 

sparrow, Eastern meadowlark, Western meadowlark, common raven (Latta and others 1999, NMPIF 

2001).

Lesser Prairie-Chicken Tympanuchus 

pallidicinctus

Distribution and Population Trends—The 

range of the lesser prairie-chicken has always been 

restricted to five States within the Southern Great 

Plains: Kansas, Oklahoma, Colorado, Texas, and New 

Mexico. In New Mexico, the greatest concentration of 

lesser prairie-chickens occurs within just 21 percent of 

its former range within the State, in southern Roosevelt, 

northern Lea, and eastern Chaves Counties (Bailey 

and Williams 2000). Small and/or isolated populations 

may also be found in Curry, extreme southern Quay, 

southern Guadalupe, central DeBaca, and northern 

Roosevelt Counties, an area comprising approximately 

23 percent of its former distribution; the lesser prairie-chicken is presumed to have been extirpated from 

the remaining 56 percent of its historical range in the 

northeastern and southeastern portions of the State 

(Bailey and Williams 2000).

The estimated current distribution of lesser prairiechickens is presented in figure 4-7. Although technically 

still found throughout this range, since the 1900s the 

numbers of lesser prairie-chickens have been drastically reduced, and the populations within this range 

have become highly fragmented and isolated from one 

another (Giesen 1998). Overall the range covered by 

lesser prairie-chickens within the boundaries of their 

historical distribution is estimated to have declined 

by 92 percent since the 1800s; between the years 1963 

and 1980 this species disappeared from 78 percent of 

its remaining range (Taylor and Guthery 1980). The 

extirpation of the lesser prairie-chicken from much 

of its former habitat reflects an estimated 97 percent 

decrease in the population of this species since the 1800s 

(Crawford 1980, Taylor and Guthery 1980). Formerly a 

highly abundant bird, in 1914 Walter Colvin reported 
94 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Lesser prairie chicken. (Photo by Gary Kramer, USDA Natural 

Resources Conservation Service)

seeing “flocks of fifty to five hundred . . . there were 

from thirty-five hundred to four thousand chickens in 

this one field, a sight never to be forgotten” (Colvin 

1914, as cited in Bent 1932:283). Yet within only a few 

years, Colvin found that the numbers of lesser prairiechickens had already noticeably diminished in this 

same area; he proposed that it was not the result of 

hunting, as hunters were so few in the area, but that 

rather it was “due largely to the cutting up of this vast 

wilderness into small farms . . . with the advancement 

of civilization the flocks scatter and become depleted” 

(Colvin 1914, as cited in Bent 1932:284).

The lesser prairie-chicken has continued its precipitous decline since that time, an apparent victim 

of habitat loss and fragmentation due to recurrent 

droughts exacerbated by overgrazing of rangelands, 

conversion of rangelands to croplands, chemical control 

of sand sage and shinnery oak; hunting may have 

played some role after populations had already been 

reduced to vulnerable levels (Bent 1932, Crawford 

1980, Hoffman 1963, Jackson and DeArment 1963, 

Mote and others 1999, Taylor and Guthery 1980). In 

New Mexico, dramatic reductions in the population 

were noticeable as early as 1926 (Ligon 1927). The 

historical population of lesser prairie-chickens in New 

Mexico was estimated at 125,000, by 1968, and that 

number is estimated to have fallen to 8,000 to 10,000 

birds (Sands 1968). Additional factors cited in the 

decline of the lesser prairie-chicken in New Mexico 

include oil and gas development and small population 

effects, such as inbreeding depression and increased 

vulnerability to stochastic environmental events that 

often occur in reduced, isolated populations (Bailey and 

Williams 2000). In addition, Bailey and others (2000) 

found that within the remaining core distribution 

of the lesser prairie-chicken in New Mexico, only 4 

percent of the potential nesting habitat could be rated 

as in good condition; the remainder was only fair (16 

percent) or poor (80 percent). The quality of nesting 

habitat is therefore also a likely limiting factor for 

lesser prairie-chickens in New Mexico.

The lesser prairie-chicken is listed as a threatened 

species in Colorado (Giesen 2000), and it has been 

recommended for listing as such in New Mexico 

(Bailey 1999). In 1995 it was proposed that the lesser 

prairie-chicken be listed as threatened under the 

Endangered Species Act. In 1998, the USDI Fish and 

Wildlife Service issued a “warranted but precluded” 

finding, indicating that although the data support 

the listing, other priorities currently take precedence 

(USFWS 1998). All attempts at translocating lesser 

prairie-chickens for reintroduction or augmentation of 

existing populations have failed to date (Giesen 1998 

and references therein, Horton 2000). Despite the 

recognition that it is likely to become an endangered 

species within the foreseeable future (the definition of 

“threatened”), the States of Kansas and Texas continue 

to allow hunting of the lesser prairie-chicken during 

limited open seasons (Giesen 1998).

Preferred Habitat and Nest Placement 

Characteristics—The original distribution of the 

lesser prairie-chicken coincided with the distribution of 

sand sage-bluestem (Artemisia filifolia-Schizachyrium

spp.) and shinnery oak-bluestem (Quercus havardii-Schizachyrium spp.) vegetation types. In the 

southwestern portion of its range (New Mexico, Texas, 

Oklahoma), the lesser prairie-chicken is found in sandshinnery oak grasslands dominated by sand bluestem 

Andropogon hallii, little bluestem Schizachyrium 

scoparium, sand dropseed Sporobolus cryptandrus, 

three-awn Aristida spp., and blue grama; sideoats 

grama B. curtipendula, hairy grama B. hirsuta and 

buffalograss may also occur (Copelin 1963, Giesen 

1998 and references therein).

Figure 4-7. Approximate current range boundaries of the lesser 

prairie chicken (from National Geographic Society 1999).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 95

The lesser prairie-chicken is not migratory and 

requires a complex of habitats within a relatively 

limited area to meet its needs across seasons. It is a 

lekking species, meaning that males congregate on 

special display grounds called “leks” to perform their 

communal courtship displays; females attend leks to 

select a mate but otherwise are entirely responsible 

for raising the young on their own. The lek forms the 

center of all other habitats for the lesser prairie-chicken; 

most birds will spend their entire life within a 5 km 

radius of the lek site (Applegate and Riley 1998). Lek 

sites are characterized by short, sparse vegetation that 

allows for excellent visibility, and the sites are often 

on elevated knolls or ridges (Copelin 1963, Jones 1963, 

Taylor and Guthery 1980). Disturbed areas created 

by human activities, such as oil pads, roads, burning, 

or herbicide treatment, may be used as display sites 

(Giesen 1998 and references therein). Usually, though, 

populations have traditional lek sites that may be 

used continuously for decades (Copelin 1963, Giesen 

1998); males display a high degree of fidelity to their 

traditional display grounds (Campbell 1972).

Because hens usually nest within 3 km of the lek 

site, they require appropriate habitat for nesting and 

brood-rearing within this range (Giesen 1994). High 

quality nesting habitat is characterized by dense 

clumps of tall grasses with scattered forbs and shrubs. 

Tall bunchgrasses (43 to 81 cm height) in dense clumps 

(1 to 3.3 m diameter) are typical in quality habitats; 

nests may be placed either at the base of such grasses or 

under sand sage or shinnery oak (Applegate and Riley 

1998, Giesen 1998). Good nesting habitat is composed 

of approximately 65 percent tallgrass cover and 30 

percent shrubs, with some additional forbs (Riley and 

others 1992). In New Mexico, female prairie-chickens 

used bluestem grasses most frequently as nesting cover. 

Greater nest success was associated with higher basal 

composition of sand bluestem (greater than 25 percent) 

and vegetation at a mean height of 66.6 cm (as opposed 

to a mean of 34.9 cm at unsuccessful nests; Riley and 

others 1992). Nest success is positively correlated with 

the height, density, and abundance of residual grasses 

(dead, standing vegetation from the previous year); 

nesting habitat is therefore particularly vulnerable to 

degradation through overgrazing, which diminishes 

the height and density of residual grasses (Bailey and 

others 2000, Giesen 1998).

Lesser prairie-chicken chicks are highly precocious 

and leave the nest within hours of hatching (Giesen 

1998). The chicks cannot move through extremely 

dense grasses, and so broods are reared in habitats 

that are more open at the ground level than those used 

for nesting (Applegate and Riley 1998). Unlike adults, 

chicks are almost entirely insectivorous and require 

habitats that will supply an abundance of grasshoppers, 

treehoppers, and beetles, their preferred prey items 

(Applegate and Riley 1998, Davis and others 1979). 

Good brood-rearing habitat is composed of 40 to 45 

percent grasses with an equal quantity of shrubs and 

the remainder composed of forbs; bare ground coverage should be about 60 percent, with the rest live and 

residual plants to provide food and cover (Riley and 

Davis 1993). The requirement for more open vegetation and increased forbs makes brood habitat less 

vulnerable to disturbance. In fact, moderate levels of 

grazing, burning, or other disturbances may improve 

dense grassland habitats by creating an open mosaic 

of vegetation that is more suitable for foraging by 

prairie-chicken adults and chicks alike (Applegate 

and Riley 1998, Bailey and others 2000, Davis and 

others 1979).

The diet of the lesser prairie-chicken changes with 

the season, and consequently so does its foraging 

habitat. In the spring and summer, birds in New Mexico 

feed on approximately 55 percent animal foods, 23 

percent vegetative material (leaves and flowers), and 

22 percent mast and seeds (Davis and others 1979). 

In the fall, the vegetable portion of the diet increases 

to include 43 percent seeds, 39 percent leaves and 

flowers, and 15 percent insects; winter birds subsist 

primarily upon the acorns of shinnery oak (69 percent 

of diet) and the seeds of wild buckwheat (Eriogonum

spp.) (Riley and others 1993a). Shinnery oaks are a 

particularly important food source for this species in 

the fall and winter, when they provide more than 50 

percent of the bird’s diet in the form of either acorns 

or other vegetative matter (Riley and others 1993a). 

Good foraging habitat in the fall and winter comprises 

approximately 60 to 65 percent grasses and 35 to 40 

percent shinnery oak or sand sage (Riley and others 

1993b). In addition, lesser prairie-chickens commonly make use of cultivated grain fields in the winter 

where these are available (Applegate and Riley 1998, 

Campbell 1972, Giesen 1998). The availability of free 

water does not appear to be a requirement at any time 

of year (Giesen 1998).

Timing of Breeding and Migration—Hens attend leks from late March through May. In Colorado, 

New Mexico, and Texas, the peak of breeding activity 

is usually during the second and third weeks of April. 

Nests are initiated in mid-April through late May, and 

hatching peaks from late May to mid-June. If the first 

nest fails, the hens may renest, extending the hatching 

period through early July. Broods of chicks break up 

at 12 to 15 weeks of age, coinciding with fall dispersal 

(Giesen 1998).

Lesser prairie-chickens are not migratory; their 

seasonal movements and home ranges are restricted to 

the suitable habitats adjacent to their lek sites (Giesen 

1998). Males may be present at lek sites from January 

through June and then again from September through 
96 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

November; most display activity is restricted to midFebruary through early May (Giesen 1998).

Area Requirements—In New Mexico, the home 

range of hens during the nesting season ranges from 

8.5 to 92 ha; outside of the nesting season, home 

ranges may range from 62 to 240 ha (Giesen 1998 and 

references therein). Home range requirements tend to 

increase in drought years due to the reduction in cover 

and increased scarcity of food (Copelin 1963, Merchant 

1982, as cited in Giesen 1998). An area estimated at at 

least 32 km2 and composed of a minimum of 63 percent 

good quality shrub/grassland habitat is required to 

support a population of lesser prairie-chickens over 

the long term (Mote and others 1999).

Management Issues and Recommendations—

Lesser prairie-chickens require extensive areas of 

sand sage or shinnery oak grasslands with a mosaic 

of habitats that provide adequate nesting habitat, 

brood habitat, and fall-winter habitat all within range 

of a central lek site (Applegate and Riley 1998, Davis 

and others 1979). In addition to maintaining and 

preserving high quality habitats where they still occur, 

such a mosaic of habitats may be created through the 

controlled use of proper grazing, prescribed fire, and 

brush control techniques in sand sage or shinnery oak 

grassland landscapes.

Excessive livestock grazing of grasslands, particularly during drought years, is highly detrimental to 

successful reproduction in lesser prairie-chickens 

(Merchant 1982, as cited in Giesen 1998). Long 

recognized as one of the primary threats to the lesser 

prairie-chicken (Bent 1932), intensive long-term grazing can alter plant species composition, reduce the 

abundance of preferred tall grass species, and dramatically reduce the residual vegetative cover required 

by this species (Applegate and Riley 1998, Bailey and 

Williams 2000, Bailey and others 2000).

To improve or maintain optimum nesting cover, 

grazing utilization levels should be less than 25 to 

35 percent of the annual growth of forage species 

(Holochek and others 1989, Riley and others 1992, 

1993b). Deferred and rest-rotation systems may also be 

used if high quality nesting and brood rearing habitat 

with residual grass cover is available in deferred and 

rested pastures (Applegate and Riley 1998). Cattle 

exclosures may be used in areas likely to be used for 

nesting; these areas should receive little if any grazing pressure (Taylor and Guthery 1980). The negative 

impacts of grazing on prairie-chicken habitat during 

drought years may be mitigated by promptly reducing livestock numbers during low precipitation years 

(Bailey and Williams 2000).

Shrubs are an important component of lesser 

prairie-chicken habitat, providing shade, cover from 

predators, and food (Cannon and Knopf 1981, Davis 

and others 1979, Giesen 1998, Taylor and Guthery 

1980). Brush control programs that have resulted in the 

eradication of sand sage and/or shinnery oak have 

negatively impacted lesser prairie-chickens (Cannon 

and Knopf 1981, Haukos and Smith 1989, Jackson and 

DeArment 1963). If more than 40 to 50 percent of the 

landscape coverage is shinnery oak, limited reduction 

of the shrub may benefit the lesser prairie-chicken, 

provided that subsequent management allows for an 

increase in tall bunchgrasses (Davis and others 1979, 

Doerr and Guthery 1980, Mote and others 1999). 

However, any reduction in shinnery oak simultaneously 

reduces an important source of winter food (Jackson 

and DeArment 1963), and as this species reproduces 

vegetatively, any eradication is essentially permanent 

(Bailey and Williams 2000). If adequate food and cover 

remain, chemical control of shinnery oak and sand sage 

should not have a negative impact (Donaldson 1969, 

as cited in Giesen 1998, Olawsky and Smith 1991).

If shrub control is desirable, no more than 50 to 

70 percent of the shrub cover should be eliminated 

to maintain appropriate habitat for lesser prairiechickens (Doerr and Guthery 1980). Applying chemical 

controls so as to create a mosaic of treated and untreated areas is one option (Olawsky and Smith 1991). 

Pesticide (herbicide and insecticide) use should always 

be kept to a minimum because it not only reduces the 

shrubs and forbs that support the insects that are an 

important food source for the prairie-chickens, but it 

may also directly reduce insect populations (Applegate 

and Riley 1998).

Prescribed burning can be used to improve habitat 

by increasing vegetative growth and insect abundance; 

burns should be limited to 20 to 33 percent of the management unit to ensure the preservation of residual 

nesting cover (Bidwell and others 1995, cited in Mote 

and others 1999). Burned areas may not recover to 

the point of providing adequate cover for nesting until 

2 to 3 years after the fire (Boyd and Bidwell 2001). 

Applegate and Riley (1998) suggest burning in the late 

winter or early spring every 3 to 4 years over 20 to 33 

percent of rangeland to rejuvenate grasses.

Burns, herbicides, or mowing may also be used to 

create artificial leks of short, sparse vegetation in 

areas of extensive but relatively homogeneous habitat 

where natural lek sites are lacking. Lek sites should be 

placed in elevated areas where possible and should be 

placed at least 1.2 km apart (Taylor 1980). Applegate 

and Riley (1998) suggest that a habitat complex should 

contain at least six and preferably 10 or more lek sites 

with a distance between leks of about 2 km. During 

the breeding season (April through July), activities 

should be restricted within a 3 km radius of any lek 

site, whether natural or artificially created, in order 

to minimize disturbance of nests and chicks (Giesen 

1998, USFWS 1998).

As fragmentation and isolation of populations is one 

of the greatest threats to the lesser prairie-chicken, 

extensive, contiguous expanses of native shrub-grass-
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 97

land habitats should be preserved whenever possible 

to maintain connectivity between the remaining 

populations. In areas dominated by trees or exotic 

grasses, lesser prairie-chickens would benefit from 

the conversion of these areas to native shinnery oak, 

sand sage, and native, warm season grasses (Applegate 

and Riley 1998). The provision of small grainfields 

within the general area may help to sustain lesser 

prairie-chickens in the winter by augmenting their 

food supply (Applegate and Riley 1998).

Associated Species—Swainson’s hawk, ferruginous hawk, scaled quail, northern bobwhite, mourning 

dove, common nighthawk, scissor-tailed flycatcher, 

western kingbird, loggerhead shrike, Chihuahuan 

raven, northern mockingbird, Cassin’s sparrow, 

lark bunting, chestnut-collared longspur (C. Rustay, 

personal communication 2002, S.O. Williams, personal 

communication 2002).

Scaled Quail Callipepla squamata

Distribution and Population Trends—The 

scaled quail or “cottontop” is a common bird in the desert 

grasslands and sparse scrublands of the Southwest. 

Although primarily a Mexican species, the range of the 

scaled quail extends northward from Central Mexico 

into southwestern Kansas, southeastern Colorado, the 

Oklahoma Panhandle and southwestern Oklahoma, 

western and southern Texas, and throughout most 

of New Mexico and southeastern Arizona (American 

Ornithologists’ Union 1998). In Arizona, the scaled 

quail is found in southeastern and south-central areas, 

west to the Baboquivari Mountains, north to the Gila 

River Valley and to the Arizona New Mexico border. The 

species also occurs locally in the upper Colorado River 

drainage, possibly through introductions (Phillips 

and others 1964). In New Mexico, the scaled quail is 

found almost Statewide below elevations of approximately 2,100 m (Hubbard 1978). The scaled quail is 

a nonmigratory resident; the breeding and wintering 

distributions and densities of this species in the United 

States are shown in figures 4-8 and 4-9.

Recent literature on the scaled quail refers to declines 

in the abundance of this species over its range in the 

United States (for example, Schemnitz 1994). Inspection 

of Breeding Bird Survey data, however, reveals that the 

reported surveywide decline (-3.7 percent annually, 

p = 0.05, n = 147) appears to be primarily driven by 

significant declines in Texas. During the years 1966 to 

2000, there were no significant trends in the numbers 

of scaled quail in any other State (Sauer and others 

2001). Natural variability in the abundance of scaled 

quail makes population trend analysis a challenge for 

this species. Researchers refer to the characteristic 

“boom and bust” cycles of scaled quail populations, when 

periods of abundance are interrupted by widespread 

reproductive failure in the species (Payne and Bryant 

1994, Schemnitz 1994). The cycling of scaled quail 

populations is thought to be in part driven by patterns 

in rainfall, with widespread reproductive failure occurring as the result of drought and an inadequate supply 

of succulent green vegetation (Campbell and others 

1973, Schemnitz 1994). High mortality may also follow severe winters with deep snows (Schemnitz 1994). 

Several authors have speculated that excessive grazing 

by livestock has degraded habitat for the scaled quail 

by altering the composition and structure of the plant 

communities used for both food and cover, leading to 

at least localized declines (Brennan 1993, as cited in 

Schemnitz 1994, Ligon 1937, Schemnitz 1994). The 

scaled quail is a High Responsibility species for New 

Mexico Partners in Flight (NMPIF 2001).

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Figure 4-8. Distribution and densities of breeding 

scaled quail in the United States and Canada, 

as mean numbers of individuals detected per 

route per year. Data averaged from Breeding 

Bird Surveys over the years 1982 to 1996 (Sauer 

and others 2001).
98 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Scaled quail.

Preferred Habitat and Nest Site Selection—

There have been surprisingly few studies of habitat 

requirements for the scaled quail, especially considering its status as a gamebird. In general, desert 

grasslands with scattered shrubs and seed-producing 

forbs appear to provide high quality habitat for scaled 

quail. Breeding and nonbreeding habitat are believed 

to be essentially the same (Schemnitz 1994). In New 

Mexico, scaled quail are common in areas dominated by 

bunchgrasses such as mesa dropseed Sporobolus flexuosus with a scattered shrub component (Saiwana and 

others 1998). Bunchgrasses tend to have a relatively 

large amount of bare ground in the interstitial spaces, 

providing for easy passage by quail that typically walk 

rather than fly for casual movements. Such openings 

between the grasses are particularly important for the 

movements of chicks. Probably for the same reason, 

scaled quail are generally not found in grasslands 

of dense, stoloniferous grasses such as black grama 

(Saiwana and others 1998, Smith and others 1996). In 

northern New Mexico, scaled quail are found in blue 

grama grasslands with scattered shrubs (Zwartjes and 

others 2005). In southern Arizona, a combination of 

low-growing grasses, forbs, and shrubs with a ground 

cover between 10 and 50 percent has been described 

as optimum habitat for the scaled quail (Goodwin and 

Hungerford 1977).

The species composition and structure of vegetation 

are key components of scaled quail habitat for many 

reasons. For one, unlike most birds, green foliage 

actually makes up a significant percentage of the 

diet. This is particularly true in the critical winter 

months when herbage comprises 30 percent of the 

food intake (Schemnitz 1994). In addition, the seeds 

of forbs and woody plants make up most of the scaled 

quail’s diet (insects are also an important diet item, 

but grass seeds are relatively unimportant; Medina 

1988, Schemnitz 1994). Certain plants are therefore 

key to providing favored food items. Mesquite seeds, 

for example, are considered important food items for 

scaled quail (Best and Smartt 1985, Davis and others 

1975). In New Mexico, seeds of snakeweed Gutierrezia 

sarothrae are a staple during the winter months (Davis 

and others 1975).

Besides providing the required food items, plants 

play a role for scaled quail in terms of providing cover 

for nesting, roosting, escaping from predators, and for 

shade and rest (Schemnitz 1994). Because scaled quail 

use a variety of plants for many different reasons, they 

are most often found in habitats that provide diversity in terms of plant species composition, structure, 

and density (for example, Saiwana and others 1998, 

Schemnitz 1994). According to Schemnitz (1994), 

quality habitat for the scaled quail includes a diverse 

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100

Figure 4-9. Winter distribution and densities 

of scaled quail, based on Christmas Bird 

Count data. Counts are average number of 

birds detected per survey over the years 1982 

to 1996 (from Sauer and others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 99

mix of grasses and forbs, accompanied by scattered 

shrubs such as yucca (Yucca spp.), fourwing saltbush 

(Atriplex canescens), littleleaf and/or skunkbush sumac 

(Rhus microphylla and R. trilobata), and various cacti 

(Opuntia spp.).

Scaled quail roost on the ground and therefore 

require roosting cover. In New Mexico, Sawyer 

(1973) found that scaled quail preferred grasses for 

roosting cover when they were available. Stormer 

(1984) proposes that overall 35 percent shrub cover 

mixed with 45 percent ground cover of grasses 0.1 to 

0.4 m tall should be adequate, if at least 1 percent 

of the shrubs provide suitable roosting cover (such 

as yucca about 0.4 m tall). Shrubs and grasses are 

also important for providing overhead cover for resting, shade, and protection from avian predators. In 

southern Arizona, mesquite and wolfberry (Lycium

spp.) were important for these purposes (Goodwin and 

Hungerford 1977); mesquite provided important loafing and escape cover in southern New Mexico as well 

(Sawyer 1973). Scaled quail will avoid areas that are 

overly shrubby, however, presumably because such a 

condition interferes with escape flight (Sawyer 1973, 

Schemnitz 1994, Wallmo 1957). If natural cover is 

not available, scaled quail will readily make use of 

artificial cover such as lumber piles, old machinery, 

and so forth for both resting and escape purposes 

(for example, Schemnitz 1994).

The ground nest of the scaled quail is a shallow scrape sparsely lined with grasses, most often 

sheltered under a shrub or other plant (Baicich and 

Harrison 1997). In southern New Mexico, scaled quail 

nests were found well concealed beneath mesquite 

or small soapweed yuccas (Y. glauca) (Sawyer 1973). 

Scaled quail nests have also been found in scrap piles 

(Schemnitz 1994).

Area Requirements—Little information is available on the area requirements of scaled quail. In 

the nonbreeding season (September through March 

or April) scaled quail form large coveys. Average 

home range sizes during the winter vary greatly: In 

Oklahoma, mean winter home range size was 21.1 ha, 

whereas in western Texas it was 145.7 ha (Schemnitz 

1994 and references therein).

Timing of Breeding and Migration—Pair formation usually begins in mid-March. The earliest nests 

with eggs are documented on April 15, and the latest 

clutch hatched on September 22 in New Mexico (Jensen 

1925). The incubation period is 22 to 23 days (Schemnitz 

1994). True second broods are rare, but renesting after 

nest failure is common (Schemnitz 1994). In New 

Mexico, most chicks hatch between June and August 

(Campbell 1968). Reproductive success appears to be 

positively correlated with rainfall during the breeding 

season (Campbell 1968, Wallmo and Uzell 1958). This 

relationship is probably based upon the effects of rain 

on vegetative growth and also on the supply of insects 

and seeds for successfully raising chicks. Vitamin A 

is important to quail reproduction, and vitamin A 

levels increase in green plants as a function of rainfall 

(Hungerford 1964, Lehmann 1953, Wallmo 1956, as 

cited in Campbell 1968).

The scaled quail is a resident nonmigrant. 

Movements between summer and winter ranges have 

been reported as less than 4 km in distance (Schemnitz 

1961).

Management Issues and Recommendations—

Scaled quail prefer an intermediate seral stage of 

grassland that contains sufficient shrubs for roosting 

and resting, enough herbaceous vegetation to provide 

seeds and insects, and good cover of bunchgrasses 

(for example, Saiwana and others 1998). Light to 

moderate grazing may be beneficial for scaled quail 

by maintaining this preferred habitat (Campbell and 

others 1973, Medina 1988). However, grazing that 

promotes high shrub density will have a negative 

impact on scaled quail, as will heavy levels of grazing 

that reduces the vegetative cover this species requires 

for food, loafing cover, and nesting cover (Schemnitz 

1994). Heavy grazing that results in the elimination 

of bunchgrasses favored by this species will also 

have a detrimental effect. Saiwana (1990, as cited 

in Schemnitz 1994) suggests that moderate grazing 

levels utilizing 30 to 40 percent of grasses should 

provide good food and cover conditions for scaled quail. 

Grazing levels should be reduced in drought years to 

maintain quality habitat (Holochek and others 1989). 

During drought conditions in Texas, habitat quality 

was better and scaled quail abundance greater in 

pastures that were subjected to high intensity, short 

duration grazing when compared to pastures that were 

grazed continuously year-round (Campbell-Kissock 

and others 1984).

In grassland areas lacking sufficient natural occurrences of loafing, nesting, or escape cover, provision 

of human-made cover may enhance habitat for scaled 

quail (Schemnitz 1994 and references therein). In New 

Mexico, Campbell (1952) suggested placing brush 

piles near water sources to improve habitat. Although 

water is attractive to scaled quail, it is considered the 

least important feature of the habitat for this species 

(Snyder 1967, as cited in Schemnitz 1994) and scaled 

quail are often found far from permanent water sources 

(for example, Brown 1989).

Because scaled quail use woody plants for many 

purposes, overly aggressive shrub control efforts 

can have detrimental impacts on their populations. 

Thinning of shrubs, however, can improve habitat 

by providing the more open, scattered woody cover 

preferred by scaled quail, particularly if such thinning 

results in increased coverage of understory grasses 

and forbs (Sawyer 1973).
100 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Associated Species—Ferruginous hawk, longbilled curlew, scissor-tailed flycatcher, Bendire’s 

thrasher, Cassin’s sparrow (NMPIF 2001).

Mountain Plover Charadrius montanus

Distribution and Population Trends—Despite 

its common name, the mountain plover is not found 

in montane regions at all but is most commonly a 

denizen of flat, arid expanses of shortgrass prairie. 

In 1929, Bent (1962) described the breeding range of 

the mountain plover as extending along the eastern 

edge of the Rockies, from New Mexico into southern 

Canada, and eastward into the Dakotas down through 

Nebraska, Kansas, and Oklahoma into Texas. Within 

this range, the species was considered a common breeding resident and was abundant enough to be a common 

game bird for market hunters (Coues 1874, as cited in 

Graul and Webster 1976, Sandoz 1954). Concerns over 

the apparent decrease in numbers of mountain plovers 

were expressed as early as 1915. While admitting that 

some of this decline in abundance may have been attributable to market hunting, Cooke (1915) speculated 

that the major threat to the species was the loss of 

suitable breeding habitat to agriculture and livestock. 

In 1957, Laun (as cited in Graul and Webster 1976) 

conjectured that the majority of the breeding population of mountain plovers was restricted primarily to 

Montana, Wyoming, and Colorado. These States still 

support the greatest numbers of breeding mountain 

plovers, with additional populations in northeastern 

to west-central New Mexico and the Oklahoma and 

Texas Panhandle regions (fig. 4-10). There is an isolated population in the Davis Mountains of Texas, and 

evidence of breeding mountain plovers has recently 

been reported in both Utah and Nuevo Leon, Mexico 

(Ellison-Manning and White 2001, Knopf 1996c and 

references therein, Knopf and Rupert 1999). Mountain 

plovers were not documented nesting in Canada until 

1981 (Wallis and Wershler 1981), and there is only 

one documented breeding record for Arizona, from 

1996 (McCarthey and Corman 1996). Currently, the 

two primary strongholds for breeding populations of 

mountain plovers are the Pawnee National Grassland 

in Colorado (Graul and Webster 1976) and the Charles 

M. Russell National Wildlife Refuge in Montana (Knopf 

and Miller 1994). Between them, these two areas are 

believed to support approximately half of the present 

continental population of nesting mountain plovers 

(Knopf and Miller 1994).

Mountain plovers winter in north-central California 

south to Baja and the northern mainland of Mexico, 

as well as southern Arizona and the southern coast of 

Texas (Knopf 1996c and references therein). By far the 

greatest concentration of wintering birds appears to be 

in the Sacramento, San Joaquin, and Imperial Valleys 

of central California (Knopf and Rupert 1995).

The mountain plover evolved in a grassland environment strongly influenced by the disturbance activities 

of vast numbers of dominant herbivores, including 

bison (Bison bison), prairie dogs (Cynomys spp.), and 

pronghorn (Antilocapra americana). The large-scale 

disappearance of these prairie animals, along with early 

market hunting practices and increased agricultural 

operations, have apparently contributed to the marked 

contraction witnessed in the breeding range of this 

species, as well as its significant decline in numbers. 

Knopf (1996c) estimates that the breeding population 

of mountain plovers may have decreased by nearly 

two-thirds between the years 1966 and 1993. As of 

1995, the estimated North American population of 

mountain plovers totaled 8,000 to 10,000 birds (Knopf 

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Figure 4-10. Distribution and densities of 

breeding mountain plovers in the United States 

and Canada, as mean numbers of individuals 

detected per route per year. Data from Breeding 

Bird Surveys averaged over the years 1982 to 

1996 (Sauer and others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 101

1996c). This is down from an estimated 214,200 to 

319,200 breeding birds in 1975 (Graul and Webster 

1976), although one of the authors acknowledges that 

estimate may have been off by as much as an order of 

magnitude (Knopf 1996c).

The mountain plover is a USFWS Migratory 

Nongame Bird of Management Concern and is a species 

designated Highest Priority by New Mexico Partners 

in Flight (table 4-1).

Preferred Habitat and Nest Placement 

Characteristics—The mountain plover both breeds 

and winters in open, flat arid country. In the breeding 

season, mountain plovers are most closely associated 

with expansive areas of short, sparse grasslands—most 

often blue grama Bouteloua gracilis and buffalo grass 

Buchloe dactyloides—with significant areas of bare 

ground (Finzel 1964, Graul 1975, Graul and Webster 

1976, Knopf 1996c; Wallis and Wershler 1981). 

Vegetation height in preferred habitats tends to be approximately 10 cm (Graul 1975, Kantrud and Kologiski 

1982, Parrish and others 1993, Wallis and Wershler 

1981), and most often the area is flat, with less than 2 

or 3 percent slope (Graul 1975, Knopf 1996c, Parrish 

and others 1993, Shackford 1991). Although sites used 

are primarily grasslands of short stature, mountain 

plovers will tolerate sparse amounts of taller vegetation, such as occasional shrubs or scattered clumps of 

cacti and bunchgrasses (Graul 1975).

On the Pawnee National Grassland in Colorado, 

an area of shortgrass prairie, sites used for nesting 

were composed of 68 percent grass cover (blue grama 

and buffalo grass) and 32 percent bare ground; prickly 

pear (Opuntia spp.) occurred on only 7 percent of the 

plots (Knopf and Miller 1994). By contrast, control 

plots ranged from 85 to 88 percent grass cover, and 

the number of plots with prickly pear on them ranged 

from 22 to 33 percent. When nesting in mixed grass 

prairie or sparse, semidesert shrublands, mountain 

plovers tend to use highly disturbed areas that offer 

shorter vegetation and more bare ground (Knopf 

1996c). At the Charles M. Russell National Wildlife 

Refuge in Montana, mountain plovers nest in prairie 

dog colonies, in areas of approximately 27 percent bare 

ground (Olson and Edge 1985). Nest sites in blue grama 

grasslands in southern Canada ranged from 45 to 55 

percent bare soil (Wallis and Wershler 1981) and were 

nearly 72 percent bare ground in Wyoming (Parrish 

and others 1993). This plover’s strong preference for 

extensive areas of bare ground led Knopf and Miller 

(1994) to propose that 30 percent bare ground may not 

necessarily represent an optimum, but is more likely 

a minimum habitat requirement.

Their liking for bare ground explains why nesting 

mountain plovers are so frequently associated with 

highly disturbed areas, such as prairie dog towns 

(Ellison-Manning and White 2001, Knowles and 

Knowles 1984, Knowles and others 1982, Olson and 

Edge 1985, Olson-Edge and Edge 1987) and areas that 

have been subjected to heavy grazing (Dechant and 

others 2001b, Graul and Webster 1976, Kantrud and 

Kologiski 1982, Knowles and Knowles 1984, Knowles 

and others 1982, Wallis and Wershler 1981). Intensive 

summer grazing appears to create particularly favorable habitat for the mountain plover (Graul and 

Webster 1976, Wallis and Wershler 1981). Mountain 

plovers are associated with an intensity of grazing 

that is described by Knopf (1996a: 141) as “heavy 

grazing pressure to the point of excessive surface 

disturbance.”

The nest of the mountain plover is a simple scrape in 

the ground, usually placed in a flat area of short grass 

(blue grama, sometimes mixed with buffalo grass), 

or in an area of entirely bare ground (Knopf 1996c). 

Mountain plovers also demonstrate a clear tendency 

to construct their nests near some conspicuous object 

in the environment; usually this is a pile of dried cow 

manure or a rock (Graul 1975, Knopf and Miller 1994, 

Olson and Edge 1985, Wallis and Wershler 1981). 

Mountain plovers are also found nesting in cultivated 

fields, either on plowed, barren ground, in early, lowgrowing croplands, or in stubble fields (Shackford 1991, 

Shackford and others 1999), although productivity 

in these habitats is probably low due to agricultural 

operations (see discussion below). After the chicks 

hatch, broods are immediately moved to areas with 

greater densities of forbs or with structures such as 

fence posts or watering tanks that serve to provide 

shade (Graul 1975).

In migration and during the winter, mountain 

plovers occupy habitats that are generally similar to 

those used in the breeding range. Particularly on the 

winter range, these plovers spend much of their time 

on plowed fields, but this is more likely due to the 

dominance of these habitats in the landscape than to 

choice; mountain plovers will preferentially use heavily 

grazed native grasslands or burned fields if they are 

available (Knopf and Rupert 1995). In New Mexico, 

mountain plovers are often located on commercial sod 

farms in the winter (Knopf 1996c). Alkali flats are used 

when available (Knopf and Rupert 1995).

Area Requirements—At the Pawnee National 

Grassland site in Colorado, Knopf and Rupert (1996) 

studied the minimum area required by mountain plovers to successfully raise chicks. They found that the 

absolute minimum area required was 28 ha, but the 

average was closer to 57 ha (range 28 to 91ha). However, 

it should be noted that these plovers raised their chicks 

in broadly overlapping territories; therefore, a suitable 

area of habitat may provide the potential for two or 

three individuals to produce successful broods.

In areas where mountain plovers depend on prairie 

dog colonies to provide the appropriate habitat, this 
102 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

species demonstrates a clear preference for large colonies. In Montana, the average size of prairie dog towns 

this plover used was 57.5 ha (Knowles and Knowles 

1984); towns less than 10 ha were considered to be of 

marginal habitat quality (Dechant and others 2001b).

Timing of Breeding and Migration—Mountain 

plovers arrive on the breeding grounds from early 

March to mid-April and depart for their winter 

range from early July through August (Knopf 1996c). 

Mountain plovers become more gregarious in migration and in winter and may start to be seen forming 

small flocks in June or July on the breeding grounds 

(Graul 1975).

In Colorado, the earliest clutches were laid in mid 

April, and the latest clutches were started in mid-June 

(Graul 1975). Although only one brood is raised per 

individual per season, mountain plovers may increase 

their productivity by double-clutching. The female first 

lays a complete clutch that the male incubates, and 

then lays a second complete clutch that she incubates 

herself (Graul 1975, Knopf 1996c). Mountain plovers 

may renest after the loss of a clutch or brood if it is 

early in the season, usually before June (Knopf 1996c). 

Incubation is 29 days, and chicks are led away from 

the nest within hours of hatching—generally as soon 

as they are dry (Graul 1975).

Management Issues and Recommendations—

One of the greatest problems facing mountain plovers 

today is their affinity for bare ground, and therefore 

their tendency to nest on plowed or cultivated fields 

(Knopf 1996c, Shackford and others 1999). Much 

of the grassland historically used by these plovers 

for breeding has been converted to agriculture, and 

these areas often end up serving as population sinks 

for the species. Early in the season, when the ground 

is bare or vegetation is short, plovers will nest in 

these fields, only to have their nests destroyed by 

farm equipment working the fields later in the spring 

(Knopf and Rupert 1999). They will also abandon their 

nests once the vegetation grows too tall, at about 20 

cm in height (Knopf and Rupert 1999). Knopf and 

Rupert (1999) speculate that the decline in productivity of mountain plovers now relegated to nesting 

in agricultural fields may largely explain their more 

recent population declines. To reduce nest and chick 

losses, they advise not preparing fields for later sowing 

months in advance, so as not to attract nesting plovers 

early in the season when their nests are certain to be 

destroyed by later activities. Protection of extensive 

areas of native grasslands from conversion to tillage 

is an obvious and important strategy for protection of 

this and other grassland bird species (Dechant and 

others 2001b and references therein).

Management for the mountain plover hinges on 

providing extensive areas of flat grassland dominated 

by short, sparse vegetation, which may be achieved 

through a variety of means (Dechant and others 

2001b). Recommendations to provide the short vegetation required by mountain plovers include grazing at 

heavy intensities in summer or late winter (Wallis 

and Wershler 1981). Heavy grazing is particularly 

important to provide short vegetation in mixed-grass 

landscapes (Knowles and others 1982, Wallis and 

Wershler 1981). The wisdom of intensively grazing 

desert grasslands to manage for mountain plovers is 

questionable and in need of further evaluation.

To more closely mimic the historic conditions experienced by mountain plovers, Wallis and Wershler (1981) 

suggest that grazing pressures be varied to provide 

a mosaic of areas that are intensively grazed, lightly 

grazed, and not grazed at all. This approach should 

provide more structural heterogeneity than is present 

in the evenly grazed grasslands we see today (Knopf 

and Rupert 1999); grasslands of such even structure do 

not provide appropriate habitat for mountain plovers 

(NMPIF 2001).

In the Southwest, historically the disturbed nest 

sites used by mountain plovers in desert grasslands 

would have been provided by the activities of prairie 

dogs, not bison (Mack and Thompson 1982). As prairie 

dog towns have declined in landscape coverage by 98 

percent since 1900 (Miller and others 1994, Parmenter 

and Van Devender 1995), this has no doubt had a 

negative impact on mountain plover populations in the 

Southwest. Protection of existing prairie dog colonies 

is therefore an important management strategy for 

this species (NMPIF 2001).

Prescribed burning can also be used to maintain 

mountain plover breeding habitat (Dechant and others 2001b, Knopf 1996c). In shortgrass prairie, early 

spring burns are used to attract mountain plovers 

(NatureServe 2001), a strategy that has thus far been 

successful (Knopf 1996c).

Mountain plovers do not appear to be adversely 

impacted by oil and gas extraction activities, and 

may even be attracted to the disturbed, open areas of 

bare ground created around oil well pads (Day 1994, 

Dechant and others 2001b). It is advisable to limit any 

disturbance in the area during the breeding season, 

however (April through July; Ball 1996, as cited in 

Dechant and others 2001b).

Associated Species—Burrowing owl (if burrows 

present), horned lark, vesper sparrow (NMPIF 2001).

Long-Billed Curlew Numenius americanus

Distribution and Population Trends—The 

long-billed curlew was once common as a breeding bird 

throughout the Western United States as far east as 

Illinois and Wisconsin, and possibly Ohio (Bent 1962), 

and large migratory and wintering populations were 

found along the Eastern shores. In the 1800s, fall 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 103

market hunting took a heavy toll on wintering curlew 

populations along the Atlantic shore (Forbush 1912), 

and the long-billed curlew quickly became “only a 

rare straggler anywhere on the Atlantic coast” (Bent 

1962:98). Curlews began to disappear simultaneously 

on their more easterly breeding grounds as human 

populations expanded westward and native prairies 

were lost to agricultural and livestock uses (Bent 1962, 

Pampush and Anthony 1993 and references therein, 

Sugden 1933). As early as 1873, the last breeding 

curlew was reported in Illinois (Bent 1962). The breeding range of the long-billed curlew is now restricted 

to localized areas in the grasslands of the Western 

United States and Southwestern Canada (fig. 4-11), 

but within these areas populations are not exhibiting any significant declines (Sauer and others 2001). 

In the Southwest, the greatest numbers of breeding 

long-billed curlews are found in the shortgrass plains 

of eastern and northeastern New Mexico, extending 

into the Panhandles of Texas and Oklahoma. Although 

this species does not breed regularly in Arizona, there 

is one recent record of long-billed curlews nesting just 

west of Springerville, with some indication that this 

may not have been an isolated occurrence (T. Corman, 

personal communication 2002). The long-billed curlew 

is occasionally found in the southern part of Arizona 

as a wintering migrant (American Ornithologists’ 

Union 1998). Wintering curlews are also found in 

good numbers in eastern New Mexico and western 

Texas, particularly in the playa lakes region, although 

the greatest concentrations of winter birds are along 

the Gulf Coast of Texas and in the central valley 

of California (fig. 4-12). The long-billed curlew is a 

Highest Priority species for New Mexico Partners in 

Flight (table 4-1).

Preferred Habitat and Nest Placement 

Characteristics—Preferred habitat for the longbilled curlew in the breeding season consists of open, 

extensive areas of level to gently sloping grasslands 

such as shortgrass prairie or recently grazed or burned 

mixed-grass prairie (Dechant and others 2001c and 

references therein), often in the vicinity of a water 

source (Cochran and Anderson 1987, McCallum and 

others 1977). Vegetation is generally less than 10 cm 

tall with relatively little barren ground (Allen 1980, 

Bicak and others 1982, Cochran and Anderson 1987, 

Paton and Dalton 1994). A study at Great Salt Lake, 

Utah, reported a mean vegetation height ranging from 

4.9 to 6.5 cm (Paton and Dalton 1994), and Redmond 

(1986) found that plant growth exceeding a height of 

12 cm in the prebreeding season forced adults to forage away from their breeding territories and resulted 

in delayed egg laying. Curlews avoid areas with tall, 

dense shrubs and weedy vegetation (Pampush and 

Anthony 1993).

Although relatively short, even grasses characterize 

the breeding territory of the curlew, the nest itself is 

generally placed within a microsite of taller, denser 

grasses (Cochran and Anderson 1987, Pampush and 

Anthony 1993) . In Utah, vegetation within 3 to 6 m 

of the nest was taller than that in the surrounding 

area, and there was less bare ground within 3 m of the 

nest itself (Paton and Dalton 1994). Although curlews 

nest on dry prairie or upland areas (for example, 

Silloway 1900), these sites are usually in proximity to 

wet meadows or some other water source (Johnsgard 

1980). The proximity of mixed-grass uplands to wet 

meadows was considered the most important criterion 

for nest-site selection in Nebraska (Bicak 1977, as cited 

in Dechant and others 2001c), and dependence upon 

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Figure 4-11. Distribution and densities of breeding long-billed curlews in the United States 

and Canada, as mean numbers of individuals 

detected per route per year. Data averaged from 

Breeding Bird Surveys over the years 1982 to 

1996 (Sauer and others 2001).
104 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

moist grasslands or proximity to wetlands or some 

other water source (stock ponds, irrigation facilities) is 

seen as a key characteristic of habitat selection across 

the curlew’s breeding range (Cochran and Anderson 

1987, Ligon 1961, McCallum and others 1977, Paton 

and Dalton 1994). In some cases, however, curlews 

may be found nesting far from water because curlews 

tend to return to the same location to nest each year, 

regardless of whether the water source that originally 

attracted them to the area is still available or not 

(McCallum and others 1977).

The nest of the long-billed curlew is a simple scrape 

or hollow, thinly lined with grasses or other materials, 

although occasionally they will build a substantial platform of grasses to hold the eggs (Bent 1962). Curlews 

appear to choose nest sites near some conspicuous 

component of the environment, such as a shrub, mound 

of dirt, rocks, or discarded metal can, and nests are 

frequently reported next to piles of horse manure or 

cow dung (Dechant and others 2001c and references 

therein). Nests are also often placed on small hummocks 

or raised areas, presumably to aid with predator detection and/or to avoid moisture or flooding when nesting 

in wet areas (Cochran and Anderson 1987).

Long-billed curlews only rarely utilize agricultural 

lands such as hay fields, croplands, fallow fields, or 

stubble fields for nesting (Bent 1962, Cochran and 

Anderson 1987, McCallum and others 1977, Renaud 

1980, Salt and Wilk 1958, Shackford 1994), although 

these areas are used for foraging during incubation, and 

birds will move into these areas with their young to loaf 

and feed once the young have left the nest (Johnsgard 

1980, Salt and Wilk 1958). In the breeding season and 

winter, curlews forage primarily in upland grasslands, 

often in prairie dog colonies, if available, feeding on 

insects, worms, caterpillars, and occasinally berries 

or the eggs and nestlings of other birds (Dechant and 

others 2001c, Ehrlich and others 1988, Terres 1991). 

Their ability to detect prey may be one reason that 

curlews prefer short vegetation. Even though prey 

density may be higher in areas of tall grass, prey 

capture rates have been observed to be higher in short 

grass (Bicak 1983, Bicak and others 1982).

During migration and winter, curlews are often 

found in small groups, foraging in areas of shortgrass 

similar to that used for breeding. In these seasons they 

will also use grasslands that are even more sparse in 

terms of vegetation height and cover; however, they 

consistently eschew thick or brushy vegetation.

Area Requirements—Studies in Idaho found that 

long-billed curlew density during the breeding season 

was positively correlated with the amount of area 

providing suitable habitat, defined as vegetation less 

than 10 cm in height (Bicak and others 1982). In the 

most densely populated areas, territory sizes averaged 

14 ha, with an unoccupied buffer zone of 300 to 500 

m around the edge of suitable habitat (Redmond and 

others 1981). In Washington State, territories were 

larger (20 ha) in flat, more homogenous grassland 

vegetation as compared to areas with more rolling hills 

and shrubs near the nesting sites (6 to 8 ha territory 

size) (Allen 1980).

The New Mexico Partners in Flight Bird Conservation 

Plan recommends a minimum of 2 ha per breeding pair 

and notes that nests can be as little as 229 m apart 

(NMPIF 2001).

Timing of Breeding and Migration—Long-billed 

curlews arrive on the breeding grounds quite early. In 

New Mexico most birds arrive in March, territories 

are established and nests initiated by April, and nests 

with eggs may be found from May through June (S.O. 

Williams, personal communication 2002). In some 

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100

Figure 4-12. Winter distribution and densities of long-billed curlews, based on Christmas Bird Count data. Counts are average 

number of birds detected per survey over 

the years 1982 to 1996 (from Sauer and 

others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 105

areas, eggs have been found in nests as late as July 

(Bent 1962, Redmond 1986). Incubation takes from 

27 to 30 days, and the precocial young fledge in 32 to 

45 days (Ehrlich and others 1988). In New Mexico, 

curlews generally begin to depart on their fall migration in July (S.O. Williams, personal communication 

2002). Birds that have not successfully raised broods 

generally are the first to leave (Allen 1980, Paton and 

Dalton 1994). Long-billed curlews are single-brooded, 

and as a rule will not renest if the first attempt fails 

(Allen 1980), therefore the success of the first nest 

attempt is critical to the productivity in this species. 

Curlews tend to return to the same location to nest 

every year, and individual birds may even reuse the 

same territories (Allen 1980, McCallum and others 

1977, Redmond and Jenni 1982, 1986).

Management Issues and Recommendations—

Studies in western North America have consistently 

demonstrated that short vegetation (grasses less than 

10 cm) is one of the key components in habitat mangagement for the long-billed curlew (Allen 1980, Bicak 

and others 1982, Cochran and Anderson 1987, Paton 

and Dalton 1994). Because of this preference, there 

are several measures that can be taken to enhance 

habitat suitability for the curlew. Actions that clear 

out areas of tall, dense residual vegetation prior to the 

onset of the nesting period, particularly following years 

of high rainfall, are especially important. Prescribed 

burning helps to clear out woody vegetation, increase 

habitat openness, and maintain grasses at a shorter 

height. Redmond and Jenni (1986) noted a 30 percent 

increase in the density of nesting curlews in western 

Idaho following a fall range fire. In New Mexico, late 

summer burning may improve grassland condition for 

the following breeding season (NMPIF 2001).

Mowing or haying also has the potential to maintain 

the short grass preferred by curlews, but in utilizing 

this method on the breeding grounds timing is critically important. Mowing must be accomplished so that 

short vegetation is available early in the season, when 

curlews first arrive to establish territories, and must 

be carried out when there is no danger of damaging 

active nests (essentially avoiding the months of April 

through July).

Grazing is another avenue for providing short 

grassland vegetation, although again timing is an 

important factor. Grazing must be accomplished such 

that short vegetation is available to the curlews prior 

to their arrival on the breeding grounds (mid-to-late 

March) (Bicak and others 1982, Cochran and Anderson 

1987). Moderate grazing in known nesting areas just 

prior to nest initiation activities (usually March) may 

be beneficial (NMPIF 2001). Grazing during the nesting season should be avoided to prevent trampling of 

nests and young; nest studies in Wyoming and Idaho 

have demonstrated that curlew nests may be lost to 

trampling by livestock, and that nests in areas that 

were grazed during the incubation period had lower 

success rates than nests in ungrazed areas (Cochran 

and Anderson 1987, Redmond and Jenni 1986). Sheep 

appear to cause more damage through trampling than 

cattle (Sugden 1933). In South Dakota, curlews were 

observed to utilize unoccupied pastures, or pastures 

occupied by cattle, but did not use pastures with sheep 

present (Timken 1969).

Conversion of native prairie to agriculture has 

been indicated as one of the primary reasons for the 

extirpation of the long-billed curlew from many of its 

former breeding grounds (Bent 1962, Jewett 1936, 

Sugden 1933, Wickersham 1902). Hence, prevention of 

further losses of shortgrass prairie to agriculture is an 

important management goal, as is shortgrass steppe 

restoration. Maintenance of large areas of shortgrass 

is also important. Some researchers suggest that areas of suitable habitat need to be at least three times 

larger than the average long-billed curlew territory 

(14 ha) since curlews generally do not occupy a large 

buffer strip surrounding suitable habitat, bringing the 

minimum habitat area to 42 ha (Dechant and others 

2001c, Redmond and others 1981).

Because long-billed curlews often show a preference 

for moist grasslands or at least proximity to wetlands 

(Ligon 1961, Dechant and others 2001c), any action 

that would result in a lowering of the water table, 

otherwise dry out the soil substrate, or eliminate 

available water sources should be avoided.

Associated Species—American kestrel, prairie 

falcon, scaled quail, western kingbird, lark bunting, 

vesper sparrow, Say’s phoebe, western meadowlark 

(NMPIF 2001).

Burrowing Owl Athene cunicularia

Distribution and Population Trends—In the 

United States, breeding burrowing owls may be found 

in areas of open shortgrass prairie, desert grasslands, 

agricultural lands, and urban environments such as 

airports and vacant lots throughout the West (fig. 4-

13; there is a separate subspecies of burrowing owl in 

Florida). Northerly breeding populations are migratory 

and will vacate the breeding grounds as far south as 

central California, New Mexico, Arizona, and northern 

Texas (fig. 4-14). California, New Mexico, and Arizona 

are considered important wintering areas for this species in the United States (James and Ethier 1989). In 

the Southwest, populations in the southern portions of 

New Mexico and Arizona are resident year-round.

In the United States, the burrowing owl is considered 

either endangered or a species of special concern in 

several States (Haug and others 1993 and references 

therein). Although reported as declining across its 

range (Haug and others 1993), Breeding Bird Survey 
106 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

data do not show a significant downward trend over the 

years 1966 to 2000 (-1.6 percent annually, p = 0.56, n 

= 288; Sauer and others 2001). The greatest cause for 

concern is the Canadian population, which has shown 

dramatic and significant declines on the order of 11.6 

percent a year (p = 0.02, n = 8) over the same 35 years 

(Sauer and others 2001). The burrowing owl has been 

extirpated from British Columbia, and the species is 

listed as threatened by the Committee on the Status 

of Endangered Wildlife in Canada (Haug and others 

1993). This species is a USFWS Migratory Nongame 

Bird of Management Concern, is a High Responsibility 

species for New Mexico Partners in Flight, and is a 

priority species in the Arizona Partners in Flight Bird 

Conservation Plan (table 4-1).

In areas where numbers of burrowing owls are decreasing, these declines have been attributed primarily 

to habitat loss and fragmentation through agricultural and urban land conversion and extermination 

of burrowing mammal colonies (Bent 1938, Dundas 

and Jensen 1995, Haug and others 1993, Rodriguez 

Estrella and others 1998). Other factors cited in declines 

include a combination of pesticide usage, increased 

predation, and vehicle collisions (Clayton and Schmutz 

1999, Haug and Didiuk 1991, James and Espie 1997, 

Sheffield 1997, Wellicome 1997). Vehicle collisions are 

a major source of mortality, particularly in fragmented 

habitats, as these owls will sit or hunt along roads at 

night, or attempt to feed on roadkill prey items (Bent 

1938, Haug and Oliphant 1987, Haug and others 1993, 

Konrad and Gilmer 1984). Pesticides used in the vicinity of owl burrows or used for control of prairie dogs or 

ground squirrels may result in direct mortality to the 

owls or indirect mortality through the loss of their prey 

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Figure 4-13. Distribution and densities of 

breeding burrowing owls in the United States 

and Canada, as mean numbers of individuals 

detected per route per year. Data averaged from 

Breeding Bird Surveys over the years 1982 to 

1996 (Sauer and others 2001).

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100

Figure 4-14. Winter distribution and densities 

of burrowing owls in the United States and 

Canada, based on Christmas Bird Count data. 

Counts are average number of birds detected 

per survey over the years 1982 to 1996 (from 

Sauer and others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 107

base (Baril 1993, Berkey and others 1993, James and 

Fox 1987, James and others 1990, Wellicome 1997). 

These small owls are already vulnerable to predators, 

and in fragmented habitats predation rates may be 

particularly high. Burrowing owls fall prey to a variety 

of predators, especially badgers, but also to domestic 

cats, opossums, weasels, skunks, and dogs (Bent 1938, 

Butts 1973, Green 1983 as cited in Haug and others 

1993, Haug 1985).

Perhaps the greatest threat to this owl is the lost 

of nest burrows, the most critical component of its 

required breeding habitat (Haug 1985, Martin 1973, 

Thomsen 1971, Wedgwood 1978). Programs dedicated 

to the eradication of burrowing mammals such as 

prairie dogs have effectively eliminated most of the 

prime breeding habitat for burrowing owls, and this 

eradication is widely believed to be the primary reason 

for their decline (Butts 1973, Butts and Lewis 1982, 

Desmond and Savidge 1998, Faanes and Lingle 1995, 

Grant 1965). Activities such as intensive agriculture 

and urbanization have also been harmful in destroying 

the burrows used by these owls, as well as eliminating 

foraging habitat (Bent 1938, Faanes and Lingle 1995, 

Haug 1985, Konrad and Gilmer 1984).

Preferred Habitat and Nest Placement 

Characteristics—Burrowing owls are found in 

treeless, dry, open flat areas of shortgrass such as 

shortgrass prairie, deserts, or shrubsteppe, and are 

usually associated with colonies of burrowing mammals, most often prairie dogs or ground squirrels (for 

example, Bent 1938, Desmond and others 1995, Faanes 

and Lingle 1995, Konrad and Gilmer 1984, Plumpton 

and Lutz 1993). Burrowing owls may occasionally be 

found using old dens and burrows of badgers, marmots, 

skunks, armadillos, kangaroo rats, tortoises, foxes, and 

coyotes (Dechant and others 2001d, Haug and others 

1993). Vegetation is short and sparse, and there is 

often a fair amount of bare ground. The availability of 

appropriate burrows for nesting is the critical limiting 

habitat factor for this owl, as the Western burrowing 

owl does not excavate its own burrows (Haug 1985, 

Martin 1973, Thomsen 1971). These owls will also 

occasionally use natural cavities in rocks if mammal 

burrows are not available (Gleason and Johnson 1985, 

Rich 1986). Burrowing owls often line their burrows 

with manure, speculated to be a predator avoidance 

mechanism by masking the smell of the burrow (Green 

and Anthony 1989, Martin 1973).

Burrowing owls appear to prefer areas of native 

grassland (for example, Clayton and Schmutz 1999), 

but also use pastures, fields, road right-of-ways, and a 

variety of urban habitats such as golf courses, cemeteries, and airports as long as the required nesting burrows 

are present (Botelho and Arrowood 1996, Dechant and 

others 2001d, Haug and others 1993). These owls are 

occasionally found nesting in croplands, but most of 

these nests probably fail when the land is cultivated 

(Dechant and others 2001d and references therein).

Burrowing owls prefer areas of short vegetation for 

nesting, most often areas of grasses approximately 

10 cm in height (Butts 1973, Butts and Lewis 1982, 

Clayton and Schmutz 1999, Plumpton and Lutz 1993). 

Burrowing owls may be more closely associated with 

black-tailed prairie dogs than with white-tailed prairie 

dogs because the former keep the vegetation shorter 

and more open around their colonies (Martin 1983). 

Although these areas of short stature are used for 

nesting, burrowing owls often use nearby areas of taller 

vegetation for foraging, as they hunt for insects and 

small mammals, their primary prey items (Dechant and 

others 2001d, Haug and others 1993). Prey species for 

this owl tend to be more abundant in areas with taller 

vegetation (Wellicome 1994, Wellicome and Haug 1995), 

although vegetation more than 1m tall appears to be 

too tall for them to forage successfully (Dechant and 

others 2001d). Burrowing owls also appear to benefit 

from the presence of scattered observation perches 

such as shrubs or fence posts for hunting and detection of potential predators, although this practice is 

not as common in areas of quite short vegetation (for 

example, less than 5 to 8 cm; Grant 1965, Green and 

Anthony 1989, Haug and others 1993).

Large, densely populated prairie dog towns appear 

to provide the best nesting habitat for burrowing owls. 

Owls in larger colonies (some 35 ha) have higher rates 

of nesting success, experience lower rates of predation, 

and are more likely to return to nest at the same site in 

subsequent years than are owls that use smaller colonies 

or colonies with fewer prairie dogs (Butts 1973, Desmond 

and Savidge 1996, 1998, 1999). The abundance and 

density of nesting burrowing owls increase as a function 

of the number of active prairie dog burrows in a colony 

(Desmond and others 2000). Burrowing owls are only 

infrequently found in abandoned prairie dog colonies 

(Bent 1938, MacCracken and others 1984) as the structural integrity of the burrows deteriorates rapidly once 

the prairie dogs are absent, and the area is soon invaded 

by taller, denser vegetation (Butts 1973, Grant 1965). 

Burrowing owl. (Photo © Courtney Conway, used with permission)
108 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

In Oklahoma, colonies became unsuitable for occupation by burrowing owls between 1 and 3 years after 

prairie dogs were removed (Butts 1973).

Area Requirements—Burrowing owls tend to 

cluster their nests within prairie dog colonies (Butts 

1973, Desmond and Savidge 1996, Desmond and others 2000). Colonies of some 35 ha appear to provide 

adequate space requirements for these owls; mean 

nearest neighbor distances in these larger colonies 

was 125 m in Nebraska (Desmond 1991, Desmond and 

Savidge 1996, Desmond and others 1995). In Oregon, 

successful nests were generally more than 110 m apart 

(Green and Anthony 1989), and in Colorado mean 

nearest neighbor distances within prairie dog colonies 

were 101 m (Plumpton 1992, as cited in Dechant and 

others 2001d). Minimum nesting area requirements 

are estimated to range from 3.5 ha (Thompson 1984, 

as cited in Dechant and others 2001d) to 7.3 ha (Grant 

1965). Foraging area requirements are considerably 

greater, although there are few quantitative studies. In 

Saskatchewan, mean foraging areas ranged from 35 ha 

(Sissons and others 1998) to 241 ha (Haug 1985, Haug 

and Oliphant 1990). As might be expected, foraging 

areas tend to be larger in areas with lower densities 

of potential prey (Haug 1985).

Timing of Migration and Breeding—Little is 

known of the migration and wintering ecology of this 

species. Northern migratory populations are believed 

to migrate southward in September and October and 

to return to their breeding grounds in March and 

April (Haug and others 1993). In the Southwest, more 

southerly populations in New Mexico and Arizona are 

believed to be permanent residents.

In New Mexico, egg laying commences in mid to 

late March (Martin 1973). Burrowing owls normally 

produce only one brood per season but will attempt to 

renest if the first nest fails early in the breeding season 

(Butts 1973, Wedgwood 1976). Incubation is 28 to 30 

days, and the young leave the nest at about 44 days 

(Haug and others 1993). After hatching, the parents 

may divide the brood and move them between several 

nearby satellite burrows, possibly to reduce the risk 

of predation or to avoid nest parasites (Butts 1973, 

Butts and Lewis 1982, Desmond and Savidge 1998, 

1999, Konrad and Gilmer 1984, Plumpton and Lutz 

1993). For the most part, these owls may be able to 

nest successfully in areas subject to moderate levels 

of human disturbance (for example, on golf courses, 

college campuses; Botelho and Arrowood 1996, Haug 

and others 1993); they are most vulnerable during 

the egg laying and incubation stages (Olenick 1990, 

as cited in Haug and others 1993).

Although some migratory birds may return and 

use the same burrow in following years, more often 

burrowing owls will return to the same traditional 

nesting areas but not use the same burrow (Dechant 

and others 2001d and references therein). Resident 

birds use and maintain their burrows year-round 

(Haug and others 1993).

Management Issues and Recommendations—

Eradication of fossorial mammal colonies quickly 

eliminates breeding populations of burrowing owls 

(Butts and Lewis 1982, Haug and others 1993). The 

protection, maintenance, and restoration of burrowing 

mammal populations, especially black-tailed prairie 

dogs in the Southwest, is one of the most important 

steps that can be taken to conserve this owl (Dechant 

and others 2001d, Haug and others 1993, Latta and 

others 1999, NatureServe 2001). Protection or expansion of large prairie dog colonies (approximatley 35 

ha) may be especially beneficial (Dechant and others 

2001d, Desmond and others 1995). If control measures 

against burrowing mammals are deemed absolutely 

necessary, these should be timed to avoid the breeding 

activities of burrowing owls (Butts 1973).

Traditional nesting sites should be preserved 

whenever possible. Where natural burrows are not 

available, artificial burrows may be installed to encourage occupation by burrowing owls; such burrows have 

been used successfully in the past (Collins and Landry 

1977, Olenick 1990 as cited in Haug 1993, Thomson 

1988 as cited in Dechant and others 2001d, Trulio 

1997). Some researchers suggest providing horse or 

cow manure to burrowing owls if it is not otherwise 

available (Green and Anthony 1997).

Because burrowing owls are sensitive to the effects 

of habitat fragmentation, large, contiguous areas of 

native grasslands, especially treeless shortgrass plains, 

should be conserved to the extent possible (Clayton and 

Schmutz 1999, Warnock 1997, Warnock and James 

1997, Wellicome and Haug 1995). The short vegetation required by the owls may be maintained through 

grazing, mowing, or periodic burning (Dechant and 

others 2001d). Although burrowing owls prefer to nest 

in grasslands that have been heavily grazed (Butts 

1973, Bock and others 1993, James and Seabloom 

1968, Kantrud and Kologiski 1982, MacCracken and 

others 1985, Wedgwood 1976), it may be important to 

provide a mosaic of grassland habitats that includes 

some areas of taller grass that support a greater abundance of prey for the owls while foraging (Dechant and 

others 2001d, NatureServe 2001). Extensive areas of 

overgrazed grassland may not provide the prey base 

required to support these owls (Dechant and others 

2001d); rotational grazing will help to maintain adequate prey populations (Wellicome and others 1997). 

Note, however, that although measures such as grazing, 

mowing, or burning will provide the grasses of short 

stature used by these owls, these methods alone are 

not sufficient to provide the necessary habitat in the 

absence of fossorial mammals to construct and maintain 

the burrows used by the owls.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 109

Burrowing owls are sensitive to pesticide use. 

Spraying carbofuran within 50 m of a nest burrow 

caused a 54 percent reduction in the number of 

young per nest (James and Fox 1987). In Oklahoma, 

application of sodium fluoroacetate to a prairie dog 

colony resulted in the eradication of burrowing owls 

nesting at the site (Butts 1973 as cited in Dechant and 

others 2001d). In addition to being potentially toxic 

to the owl itself, the use of pesticides may indirectly 

impact the owls by reducing their prey base (James 

and Fox 1987, James and others 1990). If the use of 

pesticides is absolutely necessary in an area where 

burrowing owls are nesting, those compounds with the 

lowest possible toxicity to nontarget animals should 

be used (Fox and others 1989, James and Fox 1987). 

Pesticides should not be sprayed within 600 m of nest 

burrows during the breeding season (Dechant and 

others 2001d and references therein).

Associated Species—Ferruginous hawk, golden 

eagle, prairie falcon, mountain plover, horned lark, 

common raven, loggerhead shrike, lark sparrow, blackthroated sparrow, sage sparrow, McCown’s longspur, 

eastern meadowlark, western meadowlark (Latta and 

others 1999, NMPIF 2001).

Loggerhead Shrike Lanius ludovicianus

Distribution and Population Trends—The loggerhead shrike is a breeding bird across most of the 

United States and south-central Canada; although 

it once occurred in the New England States, it is no 

longer found there. Historically, the core of its range 

is thought to have been deserts, shrub steppes, and 

southern savannas (Cade and Woods 1994), but clearing of forests in central and eastern North America 

in the 1800s allowed this species to greatly expand 

its range (Yosef 1996). However, this expanded range 

again contracted beginning in the latter half of the 

20th century due to loss of appropriate habitat through 

succession, urbanization, and other causes (Yosef 1996). 

The current breeding distribution of the loggerhead 

shrike is shown in figure 4-15.

Loggerhead shrikes are migratory in the northern 

portion of their breeding range. In the more southerly 

area of its range (approximately northern California 

eastward across central Colorado, Kansas, and other 

Central States to northern Virginia, south to Chiapas, 

Mexico), the shrike is a permanent resident. The 

abundance and distribution of wintering loggerhead 

shrikes is shown in figure 4-16.

Although numerous grassland birds have been in 

decline across the United States in recent years, the loggerhead shrike is one of the few species that has shown 

consistently negative population trends across every 

region. Between the years 1966 and 2000, Breeding 

Bird Survey data indicate significant annual declines 

averaging 4.3 percent (p = 0.00, n = 407) in the East, 

3.0 percent (p = 0.00, n = 5.81) in the Central States, 

and 4.0 percent (p = 0.00, n = 421) in the West (Sauer 

and others 2001). In the Southwest, Arizona populations have declined 4.7 percent annually (p = 0.01, n 

= 50), and New Mexico populations have decreased on 

average by 5.7 percent a year (p = 0.00, n = 53) during 

the same period.

Much speculation exists on the underlying causes 

of these declines, but there is little hard evidence to 

support any one theory. In general it is thought that 

changes in land use practices and use of pesticides may 

be largely responsible (Yosef 1996). Graber and others 

(1973) suggested that early declines in Illinois, from 

roughly 1900 to the 1950s, were due to habitat loss, 

and then more rapid declines between 1957 and 1965 

may have been a response to pesticide use during that 

period. Because of its carnivorous habits and especially 

its dependence upon grasshoppers, viewed as pests 

by most in the agricultural business, shrikes may be 

exposed to high levels of toxic chemicals. Shrikes have 

been found to have elevated levels of organochlorines 

in their tissues and eggs (Yosef 1996 and references 

therein), and may be ingesting pesticides through 

consumption of prey in areas that have been sprayed 

(Anderson and Duzan 1978, Korschgen 1970). Sharp 

declines in shrike abundance were noted in areas that 

had been treated with the pesticide dieldrin, intended 

to kill the grasshoppers that make up 30 to 75 percent 

of the shrike’s diet (Yosef 1996). Because shrikes frequent shrubs and fencelines along roadways, they are 

also exposed to herbicide treatments used for roadside 

vegetation management (Yosef 1996). Other potential 

causes of decline include habitat loss or degradation 

due to urbanization, shrub removal, surface mining 

Loggerhead shrike. (Photo © Bill Schmoker, www.schmoker.

org/BirdPics, used with permission)
110 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

operations, and overgrazing. The loggerhead shrike is 

listed as a Migratory Nongame Bird of Management 

Concern by the USDI Fish and Wildlife Service and is 

a Priority species for New Mexico Partners in Flight 

(table 4-1).

Preferred Habitat and Nest Placement 

Characteristics—Loggerhead shrikes use the same 

type of habitat year-round, and in areas such as the 

Southwest where they are resident, they are highly 

territorial, and pairs will defend the same territory 

permanently (Yosef 1996). Preferred habitat is open 

areas of short grass interspersed with areas of bare 

ground and widely scattered trees or shrubs (Dechant 

and others 2001e, Yosef 1996). These shrikes are found 

in deserts, prairies, pastures, sagebrush steppe, and 

open woodlands, as well as more suburban areas, golf 

courses, and cemeteries (Dechant and others 2001e, 

Yosef 1996). In any open country, the critical component 

of the habitat is the availability of elevated perches 

for hunting and appropriate shrubs or trees for nest 

sites (Brooks and Temple 1990, Craig 1978, Dechant 

and others 2001e, Yosef 1996, Yosef and Grubb 1994). 

Shrikes demonstrate a particular fondness for dense, 

thorny, woody vegetation, which they use both for 

nesting and for their peculiar habitat of impaling prey 

items for storage (Porter and others 1975, Yosef 1996); 

barbed wire fences are also frequently used for caching prey. Loggerhead shrikes are entirely carnivorous 

and spend much of the day hunting for grasshoppers, 

lizards, amphibians, small mammals, and even some 

small birds in open areas of short grass (Yosef 1996). 

Their hunting style requires elevated perches for 

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100

Figure 4-15. Distribution and densities of 

breeding loggerhead shrikes in the United 

States and Canada, as mean numbers of 

individuals detected per route per year. Data 

averaged from Breeding Bird Surveys over 

the years 1982 to 1996 (Sauer and others 

2001).

Figure 4-16. Winter distribution and densities 

of loggerhead shrikes in the United States 

and Canada, based on Christmas Bird Count 

data. Counts are average number of birds 

detected per survey over the years 1982 to 

1996 (from Sauer and others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 111

sighting their prey; these perches may be shrubs or 

trees but are also often fences, utility lines or poles, 

leading to a high density of loggerhead shrikes along 

roadways (Yosef 1996). Fencerows or windbreaks 

between pastures also serve as popular nest and perch 

sites (Yosef 1996). In shrubsteppe habitats, loggerhead 

shrikes prefer a mosaic of shrubs and openings, with 

little slope and high horizontal and vertical structural 

diversity (Poole 1992, as cited in Dechant and others 

2001e). Although shrikes forage in areas of short grass, 

presumably because of increased prey visibility, it appears that the best habitats also offer nearby patches 

of taller grasses that may harbor a greater reservoir of 

vertebrate prey species (Dechant and others 2001e).

Area Requirements—Territories generally range 

from 4 to 9 ha (Yosef 1996 and references therein); in 

semidesert habitats, territories tend to be larger, from 

10 to 16 ha (NatureServe 2001). Territory size may 

vary both as a function of vegetation and perch density. 

Miller (1951) found that shrikes nesting in barren, 

sparsely vegetated dunes had territories two to three 

times as large as those nesting in more wooded sites. 

On agricultural lands, territory size decreased as the 

number of available perches increased (Yosef 1996).

Timing of Breeding and Migration—Loggerhead 

shrikes are early nesters. Nest construction may begin as early as February; in Arizona, eggs have been 

reported from March through June (Terres 1991, Yosef 

1996). In several regions, the peak of nest initiation 

appears to be in mid-April (Yosef 1996 and references 

therein). The incubation period is 16 days, and the 

young leave the nest 17 to 20 days after hatching (Terres 

1991, Yosef 1996). Loggerhead shrikes usually raise 

only a single brood a year, although they will attempt 

to renest following a failed attempt (Yosef 1996). In 

the Southwest, breeding pairs are resident on their 

territories year-round.

Management Issues and Recommendations—

As nest sites and hunting perches are critical habitat 

elements for the loggerhead shrike, much of the management focus is on providing these key components 

in the grassland and shrubsteppe communities of 

the Southwest. Scattered trees or shrubs, fencerows, 

and shelterbelts should be maintained or planted in 

otherwise open pastures and grasslands to provide 

suitable nest sites and perches (Dechant and others 

2001e, Hands and others 1989, Yosef 1996). Thorny 

shrubs and barbed wire fences are especially valuable components of shrike habitat that should be 

maintained whenever possible (NatureServe 2001). 

Augmenting existing fencerows by adding native, 

thorny trees or shrubs may enhance shrike habitat 

(Hellman 1994, as cited in Dechant and others 2001e). 

Shrikes do not appear to be particularly sensitive 

to human activities, so buildings or roadways near 

nest sites is most likely not an issue for this species 

(NatureServe 2001).

As loggerhead shrikes seem to do best in grasslands 

that provide a mix of short and taller grasses (Gillihan 

and Hutchings 2000), activities such as grazing and 

mowing should be closely controlled to provide areas of 

taller grasses (20 cm or greater) that provide a greater 

abundance of vertebrate prey in addition to expanses 

of shorter vegetation (Dechant and others 2001e). 

When possible, trees and shrubs used for nesting and 

perching should be protected from cattle grazing and 

rubbing by fencing or other means (Dechant and others 

2001e, Yosef 1996).

Whenever possible, use of herbicides or insecticides should be avoided in loggerhead shrike habitat 

(Dechant and others 2001e, Yosef 1996). To maintain 

roadside nesting habitats, Yosef (1996) suggests manual 

trimming or selective removal of shrubs or trees instead 

of frequent mowing or the use of herbicides.

Prescribed burning may be used sparingly to 

maintain shrike habitat by preventing the domination of woody vegetation, but if burning is conducted 

too frequently it may eliminate the trees and shrubs 

the shrike requires for quality habitat (Dechant and 

others 2001e).

Whenever possible, large tracts of native grasslands 

and sagebrush/scrub habitat should be preserved, and 

conversion of prairies to croplands should be avoided. 

To be of adequate size, protected areas should be 

large enough to support several average-sized territories of asymmetrical shape (Dechant and others 

2001e), or about 30 to 48 ha minimum in semidesert 

habitats.

Associated Species—Ferruginous hawk, American 

kestrel, scaled quail, common nighthawk, ladderbacked woodpecker, Say’s phoebe, scissor-tailed flycatcher, Bendire’s thrasher, Cassin’s sparrow, lark 

sparrow, black-throated sparrow, western meadowlark

Horned Lark Eremophila alpestris

Distribution and Population Trends—The 

horned lark is holarctic in distribution—that is, it is 

found throughout most of the northern hemisphere. 

In North America, breeding populations are found 

from the Arctic coast of northern Canada southward 

to Baja California and Central Mexico; portions of 

Central Canada and the Southeastern United States 

are excluded from the breeding range (fig. 4-17). Most 

birds are resident across their breeding ranges, but 

more northerly populations are migratory and will 

move southward in the winter months (fig. 4-18). Birds 

breeding at high elevations (for example, in tundra 

habitats) will move to lower elevations.
112 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

This species is one of the few grassland birds that 

actually increased its range and numbers following Euro-American settlement of North America. 

Historically more restricted to the West, horned 

larks began a rapid expansion into the Midwest and 

Eastern States in the late 1800s. This range expansion 

coincided with the clearing of dense eastern forests 

and replacement of tallgrass prairie with agricultural 

fields, creating the types of open, short-statured herbaceous habitats preferred by this species (Hurley and 

Franks 1976). Horned larks will even colonize quite 

small, localized areas of suitable habitat, such as the 

mowed areas surrounding airplane landing strips; 

this has allowed them to expand into regions that are 

otherwise thickly forested and would not otherwise 

support this species.

Horned lark populations have declined in recent 

years, however, averaging a 2.7 percent negative trend 

annually across the continent (p = 0.00, n = 1,681; 

Sauer and others 2001) and dropping at an even higher 

rate in the Southwest. In New Mexico, Breeding Bird 

Survey data demonstrate an average 5.1 percent annual decrease in horned lark numbers between 1980 

and 2000 (p = 0.00, n = 55); during that same period 

Arizona populations have declined by 6.2 percent a 

year on average (p = 0.001, n = 42; Sauer and others 

2001). Although no precise cause has been pinpointed 

for these decreases, the recent increases in shrub cover 

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100

Figure 4-17. Distribution and densities of 

breeding horned larks in the United States 

and Canada, as mean numbers of individuals 

detected per route per year. Data averaged from 

Breeding Bird Surveys over the years 1982 to 

1996 (Sauer and others 2001).

Figure 4-18. Winter distribution and densities of horned larks in the United States and 

Canada, based on Christmas Bird Count 

data. Counts are average number of birds 

detected per survey over the years 1982 to 

1996 (from Sauer and others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 113

in Southwestern grasslands in response to years of fire 

suppression and overgrazing have almost certainly 

played some role in eliminating suitable habitat for 

the horned lark in this region (Archer 1989, Bahre 

1995, Glendening 1952, Humphrey 1958).

Preferred Habitat and Nest Placement 

Characteristics—The horned lark occurs in habitats covering a great variety of moisture regimes and 

elevations, as long as they share certain common 

characteristics: areas of short, sparse grassland with 

few to no woody plants and preferably at least some 

bare ground (Dinkins and others 2000 and references 

therein). Areas utilized include desert grasslands, 

shortgrass prairies, open, low-growing shrubsteppe 

habitats, and alpine meadows (Cannings and Threlfall 

1981, Rotenberry and Wiens 1980, Verbeek 1967). 

Horned larks also readily make use of agricultural areas 

and are found in croplands, pastures, stubblefields, 

and roadsides (Beason 1995, Dinkins and others 2000 

and references therein).

Numerous studies have reported that grass height 

in nesting habitat is approximately 10 cm (Dinkins 

and others 2000 and references therein). Areas used by 

horned larks in early spring will be abandoned later in 

spring or early summer when the vegetation reaches 

unsuitable heights (approximately 30 to 40 cm; Dinkins 

and others 2000). The highest population densities are 

often correlated with the amount of bare ground at the 

site (Beason 1995). Ranges reported are from 10 to 37 

percent bare ground cover at occupied sites; shrub cover 

most often ranges from 0 to 7 percent (Dinkins and 

others 2000 and references therein). As an example of 

typical breeding habitat in the Western United States, 

in north-central Colorado Creighton (1974, as cited in 

Dinkins and others 2000) reported mean vegetation 

measurements of 65 percent shortgrass cover (blue 

grama and buffalo grass), 2 percent cover of mid-height 

grasses (for example, little bluestem), 6 percent sedge 

cover, 7 percent forb cover, 2 percent cactus cover, 0.8 

percent shrub cover, 17 percent bare ground, 1 percent 

rock cover, and vegetation height of 7.2 cm. Horned 

larks consistently prefer open areas of short, sparse 

vegetation and tend to increase in abundance as the 

amount of forb and shrub cover decreases (Beason 1995 

and references therein, Dinkins and others 2000 and 

references therein); they avoid forests and wetlands 

(Beason 1995, Dinkins and others 2000).

In the short term, horned larks consistently respond 

positively to heavy levels of grazing, most likely due 

to the resultant reduction in vegetation height and 

cover that this species seems to prefer (Anstey and 

others 1995 as cited in Dinkins and others 2000, Bock 

and Bock 1988). Horned larks have also been found to 

use the bare areas of ground created by cattle or bison 

disturbances (Skinner 1975). In Saskatechewan, Maher 

(1973) found horned lark densities to be three times 

greater on grazed as opposed to ungrazed grasslands. In 

assessing grazing regimes classified as light, moderate, 

or heavy, it is the moderate to heavy grazing regimes 

that create habitat most favored by horned larks (for 

example, Bock and Webb 1984, Kantrud and Kologiski 

1982, Porter and Ryder 1974, Ryder 1980). Over the long 

term, however, continuous heavy grazing that leads 

to shrub encroachment will result in the ultimate loss 

of suitable habitat for the horned lark as this species 

avoids areas of woody vegetation.

Although several studies have reported finding 

horned larks to be more common in native pasture than 

in pastures of exotic grasses (Anstey and others 1995 as 

cited in Dinkins and others 2000, Dale and others 1997, 

Johnson and Schwartz 1993, Prescott and Murphy 

1996 as cited in Dinkins and others 2000), some have 

found no difference (Davis and Duncan 1999, Sutter 

and Brigham 1998), and numerous studies have found 

high densities of horned larks in cultivated croplands 

(for example, Best and others 1997, Johnson and Igl 

1995, King and Savidge 1995, Patterson and Best 

1996, Prescott and Murphy 1999). Because they occupy 

areas with few shrubs and low vegetative cover, horned 

larks respond positively to burning. In Arizona, horned 

larks increased in numbers on burned plots for 2 to 3 

years after the burn occurred (Bock and Bock 1992); 

the increased seed set observed in grasslands postburn 

(Bock and others 1976) would be a benefit to horned 

larks in addition to the vegetative changes that result 

from a burn. On mixed-grass prairie in Saskatchewan, 

the greatest densities of horned larks were recorded 2 

years after a late summer burn (Maher 1973).

Nests are built in a small hollow or depression on 

the ground in areas of sparse grasses or barren ground 

(Baicich and Harrison 1997). Although natural depressions may be used, the female usually excavates the 

nest site herself (Beason and Franks 1973, Sutton 

1927), then constructs a nest of fine plant materials 

within the depression. Nests are often placed out of 

the wind adjacent to a tuft of grass, cowpie, or other 

object (Baicich and Harrison 1997, Porter and Ryder 

1974, With and Webb 1993).

Timing of Breeding and Migration—Resident 

horned larks are early breeders; pair formation may 

start as early as January, and nests are generally 

constructed from mid-March to mid-July. Horned larks 

usually raise from two to three broods a season and 

will renest if a nest attempt fails. Incubation generally 

lasts 11 days, and the young leave the nest about 10 

days after hatching (Beason 1995).

In Southwestern grasslands, horned larks are 

permanent residents year-round (with the exception of high-altitude populations, which move to 

lower elevations during the nonbreeding season). 

These grasslands may support additional wintering 
114 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

migratory individuals as well, but where the more 

northerly populations winter is not known.

Area Requirements—Horned larks do not appear 

to be an area-sensitive species. In tallgrass prairies 

fragments in Illinois, Herkert (1991) found them in 

patches less than 10 ha. In shortgrass habitats of the 

West, breeding territory sizes ranged from 0.3 to 1.5 ha 

on lightly grazed pastures (average 0.7 ha; Boyd 1976, 

as cited in Dinkins and others 2000) to 1.0 to 1.7 ha on 

heavily summer and winter-grazed pastures (average 

1.5 ha; Wiens 1970, 1971). In mixed-grass habitats, 

territories in pasture averaged 1.1 ha, and those in 

undisturbed areas averaged 1.6 ha (Wiens 1971).

Management Issues and Recommendations—

The horned lark is tolerant of a wide range of 

disturbances. The most limiting factor for this species 

appears to be a reduction in suitable nesting habitat 

due to increased vegetation height and/or density, and 

particularly increases in woody plants. The horned 

lark therefore generally benefits from management 

practices such as burning, mowing, or grazing to 

reduce woody species and maintain vegetation at the 

preferred short, sparse levels (Dinkins and others 

2000 and references therein). Skinner and others 

(1984) recommend prescribed burning in the spring 

to reduce woody species. As discussed above, horned 

larks respond positively to the reduction in grass height 

accompanying moderate or high levels of grazing. 

However, if grazing begins to promote woody plant 

establishment this will offset the benefits of grazing for 

horned lark habitat. Horned larks are somewhat more 

vulnerable to disturbance when nesting on croplands; 

frequent cultivation of these lands may destroy nests 

or young.

Due to their frequent use of agricultural lands, 

horned larks tend to be exposed to high levels of pesticide use, and there have been many documented cases 

of direct mortality or reduced densities of this species 

in response to such use (see Dinkins and others 2000 

for detailed discussion and references). If pesticide 

applications are required, it is recommended that only 

rapidly degrading chemicals be used, and that they 

are applied at the lowest application rates possible 

(McEwen and others 1972).

Associated Species—Prairie falcon, mountain 

plover, burrowing owl (if burrows present), mourning 

dove, eastern meadowlark, western meadowlark.

Sprague’s Pipit Anthus spragueii (Winter 

Only)

Distribution and Population Trends—

Historically, Sprague’s pipit was considered widespread 

and abundant (Robbins and Dale 1999 and references 

therein). This pipit has entirely disjunct breeding 

and wintering grounds, breeding in northern native 

prairies from southern Alberta, Saskatchewan, and 

Manitoba southward into Eastern Montana, western 

South Dakota, and northwestern Minnesota (fig. 

4-19). Wintering birds are found from grasslands of 

southern Arizona and New Mexico east to Arkansas 

and Louisiana and south to central Mexico (American 

Ornithologists’ Union 1998, Robbins and Dale 1999) 

(fig. 4-20). The species began to decline rapidly as 

Euro-American settlers in the 1800s converted the 

native prairie to agriculture (Robbins and Dale 1999). 

Thompson (1893) reported that as early as 1892, 

Sprague’s pipit had entirely disappeared as a breeding bird from areas of Manitoba where it had been 

considered abundant only 10 years earlier. In the 

northern Great Plains of the United States, Roberts 

(1932, as cited in Bent 1950:53) described Sprague’s 

pipit as being “one of the common birds of the valley” 

(referring to Minnesota’s Red River Valley), but by 1939 

Youngworth (as cited in Bent 1950:54) reported being 

able to “drive now for hundreds of miles in North and 

South Dakota and never hear or see a pipit.”

Sprague’s pipit continues to experience severe 

population declines, attributed to a combination of 

habitat loss through conversion of prairie to agriculture, 

overgrazing of native grasslands, shrub encroachment, 

and the introduction of exotic grasses and other plants 

(Bent 1950, Robbins and Dale 1999, Samson and Knopf 

1994, Stewart 1975). Breeding Bird Survey data for the 

years 1966 through 2000 indicate significant negative 

trends surveywide, with populations declining at a rate 

of 4.7 percent a year (p < 0.01, n = 126, Sauer and others 

2000). Sprague’s pipit has been classified as Threatened 

by the Committee on the Status of Endangered Wildlife 

in Canada, is listed as a Species of Special Concern 

in Arizona, and is a Highest Priority species for New 

Mexico Partners in Flight; it is a candidate for State 

listing in Arizona (table 4-1).

Preferred Habitat—Little information has been 

published on the ecology of wintering Sprague’s pipits. 

This account therefore presents data based on studies of wintering birds when it is available but will of 

necessity be supplemented with information from 

research on breeding pipits when deemed appropriate 

and will be so noted.

Sprague’s pipit is one of the few bird species considered to be truly endemic to North American grasslands, 

in particular the shortgrass and mixed-grass prairies 

(Knopf 1996a, Robbins and Dale 1999). This pipit uses 

Southwestern grasslands solely as a migrant or winter 

resident, but winter habitat preferences of the species 

are virtually unstudied. Although it has generally been 

assumed that the pipit’s winter habitat is similar to 

those observed on the breeding grounds (for example, 

Robbins and Dale 1999), the New Mexico Partners in 

Flight Bird Conservation Plan proposes that wintering 

habitat is somewhat different from summering habitat 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 115

(B. Howe, R. Meyer, personal communication as cited in 

NMPIF 2001). According to this plan, wintering pipits 

use areas with grasses greater than 0.3 m in height 

and with more than 60 percent grass canopy cover. 

Assuming that other aspects of habitat used between 

seasons may be similar, some particulars regarding 

pipit breeding habitat are presented here.

Nesting pipits demonstrate a preference for grasses 

of intermediate height and density with moderate to 

low levels of litter, low visual obstruction, and few or no 

woody plants; even grasslands with low shrub density 

are avoided (Robbins and Dale 1999 and references 

therein). Sutter (1996, 1997, as cited in Dechant and 

others 2001f) reports average vegetation characteristics 

at nest sites in Saskatchewan as 52.7 percent grass 

and sedge cover, 10.5 percent forb and shrub cover, 15.2 

percent litter cover, 16.8 percent bare ground cover, 

55.6 forb contacts per m2, 27.7 cm maximum vegetation 

height, 2.4 cm litter depth, and vegetation density of 

1.1 contacts above 10 cm and 3 contacts below 10 cm. 

Another Saskatchewan study found abundance to be 

positively associated with the cover of narrow-leaved 

grasses approximately 10 cm in height (Anstey and 

others 1995). Vegetation in alkaline meadows along 

lake borders and in dry lake bottoms has also been 

found to support Sprague’s pipits (Saunders 1914, 

Stewart 1975, Wershler and others 1991).

Although preferred breeding habitat descriptions 

range broadly from lush grasslands (for example, 

Wershler and others 1991) to areas of sparse grass 

(for example, Kantrud 1981), one characteristic that is 

mentioned with remarkable consistency is a preference 

for native grasses over exotics. Numerous studies have 

reported significantly greater numbers of Sprague’s 

pipits in native prairie when compared to pastures 

of introduced grasses such as smooth brome Bromus 

inermis or crested wheatgrass Agropyron cristatum

(Dechant and others 2001f and references therein). In 

Saskatchewan, Dale (1990, 1992, as cited in Robbins 

and Dale 1999) found that singing males were two 

to three times more abundant in native grasslands 

than in grasslands dominated by brome. Another 

study reports even more dramatic differences, with 

from four to 25 times as many singing males in native 

grassland when compared to crested wheatgrass at 

one site, and a complete absence of pipits from other 

crested wheatgrass sites in Alberta (Prescott and 

Wagner 1996, as cited in Dinkins and others 2000). 

Pipits are almost entirely absent from croplands during the breeding season (DeSmet and Conrad 1991, 

Owens and Myres 1973). Although not yet documented, 

Robbins and Dale (1999) speculate that conversion of 

native grasslands to croplands or exotic grasses on the 

wintering grounds may have had a significant negative 

impact on Sprague’s pipits.

For the most part, Sprague’s pipits appear to prefer 

undisturbed native grasslands, or if grazed, then only 

lightly to moderately so (Dale and others 1997, Faanes 

1983, Maher 1973, Owens and Myres 1973, Stewart 

1975, but see Kantrud 1981). Sprague’s pipits tend 

to decrease in abundance as a function of increased 

grazing intensity (Dale 1984) and avoid heavily grazed 

grasslands (Anstey and others 1995, Kantrud and 

Kologiski 1982, Wershler and others 1991). Many variables come to play in determining the ultimate impact 

of grazing on pipits, including the timing and intensity 

of grazing, plant species composition, moisture regime, 

and soil type (Robbins and Dale 1999 and references 

therein). In the short term, heavy grazing probably 

reduces the grass height and density below the threshold preferred by pipits. In the long term, overgrazing 

would have a detrimental impact by promoting shrub 

encroachment and invasion by exotic plants, rendering 

grassland habitats unsuitable for Sprague’s pipits. 

BBS Limit

101 and above

31 to 100

11 to 30

4 to 10

2 to 3

One and below

None counted

Figure 4-19. Distribution and densities of breeding Sprague’s pipits in the United States and 

Canada, as mean numbers of individuals detected 

per route per year. Data averaged from Breeding 

Bird Surveys over the years 1982 to 1996 (Sauer 

and others 2001).
116 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Shrub and tree encroachment due to overgrazing on 

the wintering grounds of the pipit has been implicated 

in the loss of suitable habitat for this species (Robbins 

and Dale 1999, Stotz and others 1996).

Migratory and wintering habitat for the Sprague’s 

pipit can include stubble and fallow agricultural 

fields, as well as weedy fields or pastures (American 

Ornithologists’ Union 1998, Robbins and Dale 1999, 

Terres 1991). In the Southwest (Arizona and New 

Mexico), Sprague’s pipits are found wintering primarily 

in extensive areas of well-developed desert grasslands 

(Hubbard 1978, Monson and Phillips 1981, Zwartjes 

and others 2005).

Area Requirements—The greatest densities of 

breeding Sprague’s pipits are found in areas of extensive grasslands, suggesting that the species is area 

sensitive (Dechant and others 2001f and references 

therein). The area requirements of Sprague’s pipits on 

their wintering grounds are as yet undocumented. We 

do know, however, that in the Southwest, wintering 

Sprague’s pipits are known primarily from extensive 

areas of desert grasslands (Zwartjes and others 2005), 

and tend not to occur in areas of appropriate habitat 

less than 1 ha (NMPIF 2001).

Timing of Migration—Sprague’s pipits are generally on their breeding grounds from April through 

September or October (Dechant and others 2001f). 

On the wintering grounds in the Southwest, pipits 

begin arriving in late September and continue through 

November (Phillips and others 1964). In New Mexico, 

the last migrants finish passing through by early 

December (Zwartjes and others 2005). Sprague’s pipits 

may thus be expected in Arizona and New Mexico 

anytime from September through April.

Management Issues and Recommendations—

Given the preference of Sprague’s pipits for grasslands 

devoid of shrubs or trees, periodic burning of grassland 

habitats may be an appropriate management strategy 

(Dechant and others 2001f, NatureServe 2001). In 

Saskatchewan, Sprague’s pipit numbers declined for 

the first 2 or 3 years following a burn but then recovered to densities similar to or greater than those of 

unburned control areas (Maher 1973, Pylypec 1991). 

Madden (1996, as cited in Robbins and Dale 1999) 

found that in North Dakota, pipits were most abundant 

2 to 3 years after the occurrence of fire and up to 7 

years afterward, but that none were present on native 

prairie that had not been burned or grazed in more 

than 8 years. Some researchers suggest grasslands 

should be burned every 2 to 4 years to provide suitable 

habitat for pipits on the breeding grounds (Madden 

1996 as cited in Robbins and Dale 1999, Madden and 

others 1999). The recommended frequency for winter 

habitat, particularly in the desert grasslands of the 

Southwest, must be gauged according to local conditions 

(for example, moisture regimes, species composition of 

grassland). It should be noted that although relatively 

frequent burning may be appropriate in some portions 

of the pipit’s breeding range, studies conducted in drier 

areas report that pipit abundance may remain high in 

undisturbed (unburned) native grasslands for 15 to 32 

years (for example, Dale and others 1997). Thus, the 

required frequency of burning of winter habitat in the 

Southwest is most likely far lower than that suggested 

on many parts of its breeding grounds. Following 

a burn, pipit densities should initially decline as it 

requires a year or two for the vegetation to recover to 

the desired density and height, but burning should be 

beneficial in the long run to prevent the establishment 

of woody plants and clear out excessive litter and dead 

vegetation. Whether achieved by burning or some other 

means, one of the most important management goals 

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100

Figure 4-20. Winter distribution and densities 

of Sprague’s pipits in the United States and 

Canada, based on Christmas Bird Count data. 

Counts are average number of birds detected 

per survey over the years 1982 to 1996 (from 

Sauer and others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 117

for Sprague’s pipit is to maintain grasslands that are 

free of woody vegetation (Dechant and others 2001f 

and references therein).

All published studies on the impacts of grazing 

have been carried out on the breeding grounds (Saab 

and others 1995), but the majority of studies have 

demonstrated that Sprague’s pipits respond negatively 

to heavy grazing levels (Dechant and others 2001f 

and references therein). In New Mexico and Arizona, 

Sprague’s pipits winter primarily in the desert 

grasslands along the southern border of the States. 

As grazing has been shown to have more dramatically negative effects on breeding Sprague’s pipits 

in drier, less densely vegetated grasslands (Robbins 

and Dale 1999), it is likely that pipits wintering in 

the Southwest would be more vulnerable to grazing 

impacts than populations wintering in more mesic 

environments. Although pipits feed primarily on 

arthropods, seeds are also consumed, particularly 

on migration and in the winter months (Robbins 

and Dale 1999, Terres 1991). Grazing that prevents 

the formation of seed heads could thus eliminate a 

potentially important component of the pipit’s winter 

diet. Overall, Sprague’s pipit has proven to be tolerant of many grazing regimes subjectively described 

as light to moderate, depending upon local conditions 

(Dechant and others 2001f, Robbins and Dale 1999). 

Some low level of grazing is most likely acceptable 

for Sprague’s pipit management in the Southwest, 

as long as monitoring of the population indicates 

that the species is maintaining its numbers at the 

site under the regime practiced (Dechant and others 

2001f). On the breeding grounds, deferred rotational 

grazing has been suggested as appropriate (Drilling 

and others 1985).

Because Sprague’s pipit does not nest in the 

Southwest, the usual issues associated with haying 

or mowing (that is, appropriate timing to avoid destruction of nests or young) do not apply in this case. 

Management recommendations for this species on the 

wintering grounds include maximizing the extent of 

appropriate grassland habitats and maintaining native grasslands or restoring haylands and pastures to 

native vegetation as much as possible (Dechant and 

others 2001f and references therein).

Associated Species—Grasshopper sparrow, 

chestnut-collared longspur (NMPIF 2001).

Rufous-Winged Sparrow Aimophila carpalis

Distribution and Population Trends—The 

rufous-winged sparrow, according to Allan Phillips 

(1968:902), is “a bird of exceptional interest . . . the 

most misunderstood bird in the United States.” This 

statement was doubtless based at least in part on 

the highly irruptive nature of the species, leading to 

repeated appearances and then prolonged disappearances within its limited range. Reputed to be one of the 

last two distinct bird species described in the United 

States, the rufous-winged sparrow was first discovered 

in Arizona by Charles Bendire in 1872 at Fort Lowell 

east of Tucson (Phillips 1968). Described as abundant 

in the Tucson area at the time, within a matter of years 

the population had virtually disappeared. Following 

the collection of a single specimen in 1886, the species 

was not seen again until it was reported in 1915 on the 

Papago Indian Reservation (now known as the Tohono 

O’odham Nation) southwest of Tucson (Phillips and 

others 1964). By 1931 the American Ornithologists’ 

Union had designated the rufous-winged sparrow as 

extirpated from Arizona (AOU 1931) and attributed 

the loss of the species to overgrazing (based on Swarth 

1929). In 1936 the species was rediscovered, again 

near Tucson, and has been repeatedly documented as 

present in south-central Arizona ever since although 

its population levels and areas occupied have varied 

widely from year to year (Lowther and others 1999, 

Phillips 1968). There is some speculation that the 

species may never really have been extirpated from 

Arizona, but that its presence was missed due to the 

marked changes in habitat use and dramatic fluctuations in population levels of this species over time (for 

example, Phillips and others 1964). Furthermore, as 

Phillips (1968:903) points out, “no one understood the 

bird’s requirements and everyone looked for it in the 

wrong places.”

In the United States, the rufous-winged sparrow is 

found only in south-central Arizona, being primarily 

a Mexican species (fig. 4-21). This species is known in 

Arizona primarily from the eastern portions of Pinal 

and Pima Counties (around Oracle and Tucson) and the 

western half of Santa Cruz County (near Winkelman), 

south to Nogales and west through the Tohono O’Odham 

Nation and the Sauceda Mountains in Maricopa County 

(Latta and others 1999 and references therein; Lowther 

and others 1999). During irruptions, rufous-winged 

sparrows may be found much farther afield, occurring 

as far east as Sierra Vista and west to Quitobaquito 

(Monson and Phillips 1981). This sparrow appears to 

be more abundant over its range in Mexico, occurring 

relatively commonly in portions of central Sonora and 

locally in the western region of Sinaloa (Lowther and 

others 1999 and references therein).

Encounters with rufous-winged sparrows are so 

infrequent that data on this species are not presented 

for either Breeding Bird Surveys or Christmas Bird 

Counts, making population trend analysis impossible 

(Sauer and others 1996, 2001). Populations are nonetheless considered vulnerable because much of the 

former prime habitat for this species near Tucson is 

now unsuitable due to the expansion of urban areas, 

agriculture, and/or grazing (Lowther and others 1999). 
118 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

The rufous-winged sparrow is one of the priority species included in the Arizona Partners in Flight Bird 

Conservation Plan (Latta and others 1999).

The primary threat to this species is considered to 

be loss of habitat due to grazing and urban development (Lowther and others 1999). Phillips (1968) cited 

overgrazing as the most likely cause of local extirpations 

and drastic population declines witnessed in rufouswinged sparrows in the Tucson area. As evidence, he 

points out that this species survived in isolated pockets 

only in areas that had not been heavily grazed, and 

that the most “flourishing colonies” were formerly 

found in meadows of tobosa (Hilaria mutica), a grass 

that disappears under heavy grazing pressure. Even 

light grazing practices had led to the replacement of 

formerly dominant grasses such as grama (Bouteloua 

spp.) with Aristida (Phillips 1968), and grazing reduced 

the overall grass cover required to provide suitable 

habitat for the sparrow (Phillips and others 1964). 

Phillips (1968:904) further notes that the riparian 

habitats once utilized by rufous-winged sparrows “were 

soon grazed to destruction in Arizona.” Overgrazing was 

also believed to have been the underlying cause of the 

presumptive extirpation of the rufous-winged sparrow 

from Arizona in the early 1900s (Swarth 1929).

Population losses have also been documented 

following the clearing of former habitat for housing development (Anderson 1965), and the Arizona Partners 

in Flight Bird Conservation Plan points out that many 

of the already limited areas in the State that currently 

serve as core habitat for the rufous-winged sparrow 

in Arizona are slated for further development (Latta 

and others 1999).

Brown-headed cowbirds may also be a problem 

for the rufous-winged sparrow, as Bendire (1882, as 

cited in Phillips 1968) reported that “about one-half” 

of the nests he found in a mesquite thicket in 1872 

contained cowbird eggs. Although Phillips (1968) 

found little parasitism of the species in their most 

favored habitat, grassy swales, this is unfortunately 

the habitat type used by rufous-winged sparrows that 

has been most degraded. More recent reports cite a 

lower level of cowbird parasitism than that found by 

Bendire, however. Ohmart (1969, as cited in Lowther 

and others 1999) observed only seven cowbirds reared 

out of 90 rufous-winged sparrow nests during a 4 year 

study, and cowbird eggs were reported in 17 percent 

of nests in a study at the Santa Rita Experimental 

Range (Lowther and others 1999).

Preferred Habitat and Nest Placement 

Characteristics—The preferred habitat of the 

rufous-winged sparrow is generally characterized as 

gently sloping mixed thornscrub grasslands, composed primarily of bunchgrasses with scattered spiny 

shrubs and trees (Lowther and others 1999, Phillips 

1968). Tobosa grass (Hilaria mutica) and false grama 

(Cathestecum brevifolium) are considered essential 

components of optimum habitat, and hackberry (Celtis

spp.), paloverde (Cercidium spp.), and cholla (Opuntia

spp.) are often present (Lowther and others 1999). In 

Arizona this species is typically found below 1,100 m 

in elevation (Lowther and others 1999) in relatively 

flat areas of bunchgrasses and brush, tending to avoid 

steep hillsides (Latta and others 1999).

Phillips (1968) described five types of habitat used 

by the rufous-winged sparrow: (1) grassy swales—wide, 

low channels flooded by desert rains and covered with 

tobosa grass; (2) desert washes—similar to swales but 

more gently sloping with bottom of drained sand and 

lined with paloverde, mesquite, and a brushy understory; (3) riparian habitat—flowing water, occasionally 

flooded bottomlands lined with bunchgrasses such 

as sacaton (Sporobolus spp.) and broad-leaved trees 

such as willow (Salix spp.) and cottonwood (Populus

spp.), with mesquite further back; (4) farmland 

habitat—edges of brush along cleared fields, irrigation 

ditches lined with mesquite, elderberries (Sambucus

spp.), and gray thorn (Condalia spp.); and (5) deep-soil 

habitat—scattered mesquite trees interspersed with 

clumps of sacaton.

Phillips (1968) specifically notes that the desert 

washes are marginal habitat for the species, and 

Lowther and others (1999) point out that the riparian, 

farmland, and deep-soil sites are now so altered as to no 

longer support populations of rufous-winged sparrows. 

The desert washes and surrounding uplands presently 

occupied by rufous-winged sparrows may represent the 

best remaining habitat available. However, they are 

poor substitutes for the lush broad riparian floodplains 

Figure 4-21. Approximate current range boundaries of the 

rufous-winged sparrow (from National Geographic Society 

1999).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 119

originally favored by this species that have now been 

all but eliminated within their range.

The nest of the rufous-winged sparrow is most often 

a solid, deep cup composed of grasses and lined with 

horsehair, although its exact composition depends 

on the materials available in the habitat. Nests are 

usually placed in the fork or crotch of a low shrub or 

tree, placed so that the rim is supported on each side 

and concealed by overhanging branches (Lowther 

and others 1999). Nest heights also vary by habitat 

but may range from 0.15 to 3.04 m in height (average 

about 0.5 to 1.5 m; Lowther and others 1999). Phillips 

(1968) reported nests in swale habitats at 0.6 to 2.0 

m high in the edges of thick, tall desert hackberry; 

in desert washes, at 1.3 to 2.5 m in open paloverdes 

or occasionally in clumps of mistletoe (Phoradendron

spp.) within those trees. In farmlands, nests were 1.0 

to 1.3 m high in gray thorn. Nests may also be found in 

cholla in areas where that cactus is common (Lowther 

and others 1999, Wolfe 1977).

Timing of Breeding and Migration—Pair formation generally occurs in March, and eggs have been 

documented in nests from as early as 29 March in 

Arizona through 5 September, though June through 

August is considered the peak of egg-laying activity. 

In Mexico, eggs have been found as late as 16 October 

(Lowther and others 1999 and references therein). The 

nesting season “corresponds in all cases with a season 

of rainfall and warm to hot temperatures” (Phillips 

1968:909). Ohmart (1969) reported that the birds in 

his study often began nest construction within 1 day of 

precipitation exceeding 14 mm, and in all cases within 

4 days of such a rainfall event. However, Wolf (1977) 

found that birds in Mexico would eventually begin 

nesting activities even before the summer rains began 

if the rainy season was later than normal. Rufouswinged sparrows will most often raise second broods, 

even if the first brood was successful; egg dates after 

mid- to late July are considered to represent second 

nest attempts (Lowther and others 1999).

The rufous-winged sparrow is a nonmigratory resident, and territories are defended by males year-round 

(Lowther and others 1999). Individuals are reported 

to occasionally move a short distance in search of food 

during the winter (Moore 1946).

Area Requirements—In 1882, Bendire found no 

fewer than 43 nests of the rufous-winged sparrow in an 

area 100 m long by 400 m wide, in what Phillips (1968) 

calls their “original riparian habitat.” In favored swale 

habitats, Phillips (1968) reports most territories to be 

less than 0.5 ha. However, as is often the case, there 

tends to be an inverse relationship between territory 

size and habitat quality. In more marginal farmland 

habitats, territories are closer to 1 ha (Phillips 1968). 

More recent estimates report average territory sizes 

ranging from 0.7 to 1.2 ha at a single site in Arizona 

over a period of 3 years, although the type of habitat 

occupied there is not described (Lowther and others 

1999). Active nests may be placed as close as 14 m to 

one another, although again the type of habitat in which 

this occurred is not reported (Austin, personal communication as cited in Lowther and others 1999).

Management Issues and Recommendations—It 

is probably no coincidence that the disappearance of 

the rufous-winged sparrow in Arizona in the late 1800s 

occurred at the same time that cattle numbers there 

reached record numbers (Wildeman and Brock 2000). 

The overstocking of the range that characterized this 

period resulted in extensive degradation of the broad 

lowland riverine floodplains that once supported the 

greatest numbers of rufous-winged sparrows. Such 

intensive overgrazing results in the elimination of the 

tall bunchgrasses most favored by this species, such 

as tobosa (Phillips 1968). Because rufous-winged sparrows are currently found in Arizona only in relatively 

small, remnant areas of grassland, any activity such 

as development or improper grazing that decreases or 

eliminates these habitats will negatively impact this 

species. Protection of remaining areas of appropriate 

grassland habitats for the rufous-winged sparrow 

should be the primary conservation goal, followed 

closely by restoration of such habitats.

Arizona Partners in Flight offers the following 

specific management recommendations for the rufouswinged sparrow (Latta and others 1999):

• Grazing should be at light to moderate levels in 

prime habitat and should be closely monitored 

to maintain appropriate use.

• Maintain blocks of habitat between developments 

or green belts within developments.

• Maintain current management in areas that 

are considered to provide core habitat for the 

rufous-winged sparrow, including the Santa Rita 

Experimental Range, Buenos Aires National 

Wildlife Refuge, Saguaro National Park, and 

Tucson Mountain Park. Additional potential core 

areas identified are the Tohono O’odham lands 

around San Xavier Mission, along the western 

slopes of the Baboquivari and Coyote Mountains, 

and on the eastern and southern slopes of the 

Silver Bell Mountains.

Associated Species—Cactus wren, curve-billed 

thrasher, pyrrhuloxia, varied bunting, canyon towhee, 

Scott’s oriole (Latta and others 1999).

Cassin’s sparrow Aimophila cassinii

Distribution and Population Trends—Cassin’s 

sparrow is a true denizen of the Southwestern grasslands, being most common as a breeding bird in the 

arid shrubby grasslands of southeastern Colorado, 

eastern New Mexico, and western Texas. Its breeding 
120 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

distribution in the United States can be generally 

described as extending from southeastern Arizona 

across most of New Mexico, up into eastern Colorado 

and southwestern Nebraska, western Kansas, western 

Oklahoma, southward through the western two-thirds 

of Texas (fig. 4-22). This distribution extends into 

Mexico in the States of Chihuahua, Coahuila, in the 

interior south to Zacatecas and San Luis Potosí, and 

on the Atlantic slope from Nuevo Leon to Tamaulipas 

(Ruth 2000). Cassin’s sparrows are migratory in the 

northern portions of their breeding range, retreating 

southward into southeastern Arizona and western 

and south-central Texas, being found only rarely 

in southern New Mexico, and extending southward 

into Mexico on the Pacific slope from Sonora through 

Sinaloa to Nayarit and south in the interior including 

Chihuahua, Coahuila, Zacatecas, San Luis Potosí, and 

Guanajuato (Ruth 2000; see fig. 4-23).

The occurrence and abundance of this species 

tends to fluctuate dramatically between years, making it difficult to define its actual distribution or to 

define population trends (Dunning and others 1999, 

Ruth 2000). Particularly outside of its core breeding 

areas (southeastern Colorado, eastern New Mexico, 

and western Texas), the Cassin’s sparrow is highly 

erratic in its distribution; areas in which no Cassin’s 

sparrows are reported for several years may suddenly 

have large numbers of this species breeding in other 

years (Dunning and others 1999). Cassin’s sparrows 

show little breeding site fidelity between years, leading to the suggestion that they are almost nomadic 

(Dunning and others 1999). Defining the season of 

use and range of the Cassin’s sparrow is challenging due to the highly cryptic nature of this species. 

Inconspicuous and secretive by nature, the presence 

of Cassin’s sparrow is most reliably detected during 

the breeding season by the singing and skylarking 

display of the males. However, in the Southwest this 

behavior is closely associated with the onset of summer 

rains, which generally commence in July in Arizona 

and New Mexico. Thus, the problem remains that 

prior to these rains, Cassin’s sparrows may be present in an area but are not reliably detected (Dunning 

and others 1999, Ruth 2000). The actual limits of the 

winter range of this species are unknown because of 

the difficulty in identifying nonbreeding individuals 

(Dunning and others 1999).

The naturally variable nature of the abundance of 

Cassin’s sparrow throughout much of its breeding range 

also makes it difficult to interpret the standard data 

gathered from Breeding Bird Surveys, as well as to 

detect any historical changes in distribution (Dunning 

and others 1999). In general, Cassin’s sparrow populations appear to be stable across most of its range in 

the United States; although BBS data showed a significant decline surveywide in the years 1966 to 1996 

(-2.5 percent, p ≤ 0.01, n = 203), this downward trend 

appears to have been driven primarily by decreases in 

the Edwards Plateau and southern brushland regions 

of Texas (Ruth 2000). Current BBS data (1966 through 

2000) demonstrate significant downward trends in this 

species only in Texas (-2.6 percent, p = 0.00, n = 114) 

and Colorado (-5.0 percent, p = 0.02, n = 39; Sauer 

and others 2001). In New Mexico, the Cassin’s sparrow remains as the most abundant breeding bird in 

grasslands with a shrub component, and populations 

are apparently stable within the core of its range (H. 

Schwarz, S.O. Williams, personal communication as 

Figure 4-22. Distribution and densities of breeding 

Cassin’s sparrows in the United States and Canada, 

as mean numbers of individuals detected per route per 

year. Data averaged from Breeding Bird Surveys over 

the years 1982 to 1996 (Sauer and others 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 121

cited in Ruth 2000). Cassin’s sparrow is a USFWS 

Migratory Nongame Bird of Management Concern, a 

Highest Priority species for New Mexico Partners in 

Flight, and is a priority species in the Arizona Partners 

in Flight Bird Conservation Plan (table 4-1).

Preferred Habitat and Nest Placement 

Characteristics—Cassin’s sparrows are found in 

arid grasslands with a significant shrub component; 

alternatively, Cassin’s sparrow habitat may be defined 

as shrublands with a significant understory coverage 

of grasses (Ruth 2000). The shrub component may 

be actual shrubs or vegetative forms, such as yuccas, 

ocotillo, cacti, or bunchgrass that approximate shrub 

structure. The relative proportions of grasses and 

shrubs used by Cassin’s sparrows cover a wide range, 

but the structure provided by shrubs or similar plants 

is required by this species for both nest placement and 

for perches for males from which to sing or initiate 

their courtship flights (Ruth 2000). There is, however, 

an upper threshold of shrub density above which the 

Cassin’s sparrow will not utilize the habitat; some 

significant area of grass cover must be available as 

well (Ruth 2000).

In the Southwest, most studies of Cassin’s sparrow 

habitat have been conducted in southeastern Arizona. 

In this region, Cassin’s sparrow is found in extensive 

areas of mesquite grasslands and grassy habitats in 

the Sonoran Zones (Monson and Phillips 1981). Bock 

and Webb (1984) found that plots occupied by Cassin’s 

sparrows were characterized by a mean of 23 percent 

bare ground cover, 68.8 percent grass cover, 2.9 percent 

herb cover, and a shrub density of 23.6 shrubs per 100 

m2. Shrub canopy cover was 10.3 percent on average; 

mean grass height was 29.1 cm. Cassin’s sparrows rarely 

occupied plots with greater than 35 percent bare ground, 

and avoided areas with less than 6 percent shrub 

canopy cover. In south-central Texas, the highest 

breeding densities of Cassin’s sparrows were found in 

scrubby mesquite grasslands characterized by a shrub 

density of 717 shrubs per hectare; however, as mesquite 

increased in stature and density, Cassin’s sparrows 

decreased in abundance and eventually abandoned 

formerly used sites (Maxwell 1979, as cited in Ruth 

2000).

In southeastern Arizona, Cassin’s sparrows strongly 

avoid heavily grazed grasslands and demonstrate 

a preference for ungrazed or lightly grazed upland 

grasslands (Bock and Bock 1988, 1999, Bock and Webb 

1984). However, Ruth (2000) notes that some of the 

most substantial numbers of Cassin’s sparrows are 

supported in the grasslands of eastern New Mexico, 

most of which are grazed to some degree. As noted by 

Saab and others (1995), birds respond differently to 

grazing in different habitats and climatic conditions. 

In Arizona, Cassin’s sparrow was also found to show 

a preference for native grasslands over grasslands 

dominated by introduced lovegrasses (Eragrostis 

spp; Bock and others 1986). Cassin’s sparrows are 

only rarely found in agricultural fields (Dunning and 

others 1999).

As noted above, the use of breeding habitat by 

Cassin’s sparrows appears to be highly dependent upon 

patterns of precipitation. In the more arid grasslands 

of the Southwest, increased precipitation is positively 

correlated with abundance of Cassin’s sparrows as 

well as reproductive success (Dunning and others 

1999). In the more mesic, eastern portions of its range, 

increased precipitation may actually drive a decrease 

in Cassin’s sparrow numbers and breeding success 

(Andrews and Righter 1992, Lasley and Sexton 1993, 

Webster 1979). Although there has been speculation 

that Cassin’s sparrows are responding to the changes 

Average Count

< 1

1 - 3

3 - 10

10 - 30

30 - 100

> 100

Figure 4-23. Winter distribution and densities 

of Cassin’s sparrows in the United States 

and Canada, based on Christmas Bird Count 

data. Counts are average number of birds 

detected per survey over the years 1982 to 

1996 (from Sauer and others 2001).
122 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

in vegetative composition and structure that are 

associated with increased rainfall (for example, 

Dunning and others 1999), the birds generally react 

to the onset of summer rains within a matter of days, 

well before such changes take place (Ruth 2000). An 

alternative explanation is that these sparrows are 

responding to an increase in insect populations, their 

primary prey in the summer, which is associated with 

the increase in rainfall. For example, grasshoppers—an 

important component of the Cassin’s sparrow’s diet in 

the breeding season—respond positively to increased 

levels of precipitation in the Southwest (Dunning 

and others 1999, Ruth 2000 and references therein). 

Although the ultimate reason remains unknown, the 

erratic pattern of summer rains in the Southwest is 

apparently responsible for the similarly erratic or 

sporadic occupation of breeding habitat by Cassin’s 

sparrows in this region (Dunning and others 1999).

Winter habitat requirements are apparently similar 

to those demonstrated in the breeding season, except 

that Cassin’s sparrows retreat to the southern third of 

their range (Rising 1996). In southern Texas, wintering 

Cassin’s sparrows were found in habitats ranging from 

grasslands with scattered patches of shrubs and mesquite to relatively dense shrublands with an overstory 

of taller mesquite and acacia above the primary shrub 

layer (Emlen 1972). Wintering birds were not found in 

open grasslands without shrubs (100 percent grasses 

and forbs) or in scrub grasslands (60 percent grasses 

and forbs, 40 percent low-growing woody plants, no 

shrubs). In Arizona, Gordon (2000a) studied wintering 

Cassin’s sparrows in habitats described as semidesert 

or plains grasslands dominated by a variety of perennial 

bunchgrasses (Bouteloua, Eragrostis) with many forbs, 

small woody perennials, and a few scattered mesquites. 

As in the breeding season, wintering birds also appear 

to avoid using grazed grasslands (Bock and others 

1984, Bock and Bock 1999, Russell and Monson 1998). 

Cassin’s sparrows appear to require taller, denser 

grasses in the winter, probably for hiding and escape 

cover from predators, seed availability, and thermal 

cover (Zwartjes and others 2005). In Arizona, Cassin’s 

sparrows were more abundant in ungrazed grasslands 

that had received a high level of rainfall in the previous 

summer, presumably due to the resultant increases in 

grass cover and seed production (seeds are the dietary 

mainstay in the winter; Bock and Bock 1999).

The nests of Cassin’s Sparrows are placed in approximately equal proportions either on the ground or 

low in the base of shrubs or cacti (Ruth 2000, Williams 

and LeSassier 1968). Ground nests are generally 

concealed in clumps of grass or at the base of shrubs 

or cacti and are not sunk below the ground surface in 

a scrape (Baicich and Harrison 1997, Dunning and 

others 1999, Rising 1996, Williams and LeSassier 

1968). When placed in shrubs or cacti, nests are placed 

low (mean 4.0 to 15.0 cm from the ground) in plants of 

relatively short stature (mean 0.4 to 0.7 m in height; 

Ruth 2000 and references therein). Nests may be 

clustered, suggesting that Cassin’s sparrows may at 

times be semicolonial (Johnsgard 1979, Williams and 

LeSassier 1968).

Area Requirements—Although Cassin’s sparrows have been described as using extensive areas of 

grasslands, the impacts of habitat fragmentation on 

Cassin’s sparrows are not known (Dunning and others 

1999). In Texas, the mean territory size was 2.6 ha ± 

0.5 SD (range 1.7 to 3.3, n = 21; Schnase 1984, as cited 

in Dunning and others 1999). In Arizona, territories 

were estimated to be 0.26 to 0.35 ha, although the 

researcher postulates that this is likely an underestimate and that the true territory size is probably closer 

to that reported in the Texas study (R.K. Bowers, Jr., 

as cited in Dunning and others 1999). The Arizona 

Partners in Flight Bird Conservation Plan suggests 

that breeding Cassin’s sparrows require a minimum 

of 2.0 ha of dense Grama and bunchgrasses within 

a greater (minimum 16 ha) habitat matrix of mixed 

grasses and shrubs (Latta and others 1999).

Timing of Breeding and Migration—Breeding 

in this species varies widely according to local environmental conditions. In southeastern Arizona, 

breeding activities begin in response to the onset of the 

monsoon rains, which usually begin in July (Dunning 

and others 1999). In Texas, New Mexico, and Arizona, 

active nests have been found from March through 

September (Dunning and others 1999, Monson and 

Phillips 1981). In the more arid western portion of its 

range (Arizona and New Mexico), Dunning and others 

(1999) propose that Cassin’s sparrow may breed only 

in years of unusually high precipitation.

Whether Cassin’s sparrows raise more than one 

brood per season is unknown, although the long breeding season of this species has led some to speculate 

Cassin's sparrow. (Photo © Bill Schmoker, www.schmoker.

org/BirdPics, used with permission)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 123

that double-brooding may be possible (Baicich and 

Harrison 1997, Rising 1996, Wolf 1977). The incubation period is estimated to be 11 days, and the young 

leave the nest approximately 9 days after hatching 

(Dunning and others 1999).

Exactly when Cassin’s sparrows arrive and depart 

from their breeding and wintering grounds is hard 

to determine, based on the difficulty of detecting 

these birds when they are not actively engaged in 

courtship activities. Dunning and others (1999) note 

that the latest records in some regions do not reliably 

indicate when the last birds have departed, but only 

when the males have stopped singing. Birds probably arrive on their breeding grounds in Texas in 

March and return to New Mexico possibly as early as 

March and in large numbers by early April (Hubbard 

1977). Most of the birds have left New Mexico by late 

September (Hubbard 1977). In Texas, Cassin’s sparrows present in mid-October are considered wintering 

birds (Oberholser 1974). Cassin’s sparrows are 

highly territorial and sedentary on their winter ranges 

(Dunning and others 1999, Gordon 2000a). Studies of 

color-banded birds in Arizona indicate that Cassin’s 

sparrows here were not permanent residents but rather 

that wintering birds left in late April and early May, 

only to be replaced by a different population that then 

bred in the same area from July through September 

(Dunning and others 1999).

Management Issues and Recommendations—

As shrub-grassland specialists with a relatively 

limited distribution concentrated in the Southwest, 

Cassin’s sparrows are primarily threatened by the 

loss of habitat through conversion to agriculture, 

development, or desert scrublands (Ruth 2000 and 

references therein). Depending on local climatic conditions, poor range management and/or improper grazing 

practices may also contribute to loss of appropriate 

habitat (Bock and Bock 1988, Bock and others 1984). 

Because Cassin’s sparrows require a combination of 

shrubs and grass components in their habitat, they 

are potentially responsive to a number of management activities that affect these proportions, including 

grazing, prescribed burning, shrub control, and the 

introduction of exotic grasses (Bock and Bock 1988, 

Ligon 1961, Ruth 2000).

In mesquite grasslands of southeastern Arizona, 

both breeding and wintering Cassin’s sparrows avoided 

grazed plots (Bock and Bock 1988, Bock and Webb 

1984, Bock and others 1984). The preferred ungrazed 

plots had significantly greater grass cover and herb 

cover than did grazed plots, and also supported 

greater shrub cover; grazed plots had significantly 

greater coverage of bare ground (Bock and Bock 1988, 

Bock and others 1984). Grazing may reduce habitat 

suitability for Cassin’s sparrows by reducing vegetative height and/or density, and by altering species 

composition (Ruth 2000). Although such effects may 

actually improve habitat for Cassin’s sparrows in 

more mesic portions of its range, where vegetation 

may potentially be too dense or tall, grazing should 

probably be avoided or carefully controlled in arid 

grasslands of the Southwest where the vegetation 

is already short and sparse (Gillihan and Hutchings 

2000, Ruth 2000). Over the long term, grazing in 

shrubland habitats may result in an increase in shrub 

density to the extent that insufficient grass cover 

remains to provide appropriate habitat for Cassin’s 

sparrows. Alternatively, in pure grassland habitats, 

grazing may result in an increase in shrub cover over 

the long term, which would improve the habitat for 

Cassin’s sparrows (Ruth 2000).

Because Cassin’s sparrows require at least some 

shrub component in conjunction with grasslands (5 

percent cover has been suggested as an absolute minimum), any measures that would completely eliminate 

brush or shrubs (for example, chaining) are potentially 

detrimental (Gillihan and Hutchings 2000, Ruth 2000). 

However, controlled measures that would merely reduce shrub coverage in extremely dense shrub habitat 

(for example, selective mechanical or chemical removal) 

may benefit the species (Ruth 2000). The distribution 

of shrubs within the grasslands is apparently not 

important, as Cassin’s sparrows have been found in 

habitats with both widely scattered shrubs and with 

shrubs occurring in dense patches interspersed with 

grasslands (Dunning and others 1999, Ruth 2000 and 

references therein).

Prescribed fire may temporarily reduce local populations of Cassin’s sparrows; this species is generally 

absent from burned sites for the first year or two 

following a burn (Bock and Bock 1992, Gordon 2000b) 

probably due to the temporary alteration of grass cover 

and vegetative structure (Ruth 2000). In the long term, 

fire may be beneficial if it reduces overly dense shrub 

cover and stimulates growth of native grasses and seed 

production; short-term impacts may be mitigated by 

conducting burns in small patches to create a mosaic 

of burned and unburned sites (Ruth 2000).

The preservation of native grasslands with a shrub 

component is important for breeding Cassin’s sparrows, as this species was found in significantly lower 

numbers in habitats dominated by exotic lovegrasses 

(Bock and Bock 1988, Bock and others 1986). The 

grasslands dominated by exotics differed from native 

grasslands in terms of reduced plant species diversity, 

shrub density, and shrub canopy cover; significantly, 

numbers of grasshoppers were also lower in the exoticdominated grasslands, and grasshoppers are a diet 

staple for Cassin’s sparrows in the breeding season 

(Bock and others 1986). Seeding of exotic grasses as 

part of a range forage improvement program should be 

avoided, and native grasses should be maintained or 
124 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

restored whenever possible to provide quality habitat 

for Cassin’s sparrow and other grassland birds.

Associated Species—Loggerhead shrike, Botteri’s 

sparrow, grasshopper sparrow (Latta and others 

1999).

Baird’s Sparrow Ammodramus bairdii 

(Winter Only)

Distribution and Population Trends—The 

Baird’s sparrow is a true grassland specialist, one of the 

few species considered truly endemic to the prairies of 

the Great Plains (Knopf 1996a). This species occupies 

highly disjunct breeding and wintering grounds. Its 

breeding range is in the northern Great Plains, extending from eastern Montana, northern South Dakota, 

and western Minnesota northward into the prairies 

of southern Alberta and Manitoba. The majority of 

individuals winter in northern Mexico in the States 

of Chihuahua, Coahuila, Durango, and Sonora, but 

the winter range extends into the United States in 

southeastern Arizona, southwestern and south-central 

New Mexico, across into the Big Bend region of western 

Texas (American Ornithologists’ Union 1998, Howell 

and Webb 1995, see fig. 4-24).

In Arizona, wintering Baird’s sparrows are found 

primarily in the grasslands of the Sonoita Plains, 

the San Rafael Valley, and the upper Altar Valley; 

individuals are also found occasionally in the Sulphur 

Springs Valley (Jones and Green 1998, Monson and 

Phillips 1981). Historically this species was far more 

abundant in southeastern Arizona than it is today; it 

was described as abundant before 1878 but had become 

uncommon by the 1920s (Phillips and others 1964). 

Baird’s sparrows may be found in Arizona from midAugust through early May (Jones and Green 1998). 

The species was formerly much more common in New 

Mexico as well. Described as occurring in immense 

numbers in the 1870s (Bailey 1928), Baird’s sparrow 

today is considered a rare but regular migrant in the 

State (Jones and Green 1998). The most consistent 

records of Baird’s sparrow in New Mexico are from 

the Animas Valley in Hidalgo County and Otero Mesa, 

Fort Bliss Military Reservation, in Otero County (Jones 

and Green 1998, Ligon 1961). Other recent records of 

Baird’s sparrows are from Luna, Union, and Socorro 

Counties (New Mexico Department of Game and Fish 

1988). Baird’s sparrows may be found in the State 

starting in early August through late April, possibly 

as late as mid-May (Jones and Green 1998).

Baird’s sparrow populations have experienced significant declines in the United States in recent years. 

Although Breeding Bird Survey data did not reveal 

any significant population trends over the years 1966 

through 1979, Baird’s sparrow numbers dropped by 

an average of 4.3 percent annually over the period 

1980 through 2000 (p = 0.02, n = 46; Sauer and others 

2001). In 1991 and again in 1999, petitions to list the 

Baird’s sparrow as threatened under the Endangered 

Species Act were denied based on lack of supporting 

data (USFWS 1991, 1999b). The Baird’s sparrow 

remains on the USDI Fish and Wildlife Service’s list 

of Nongame Migratory Bird Species of Management 

Concern and is considered Threatened in both New 

Mexico and Arizona. It is a Highest Priority species for 

New Mexico Partners in Flight and is a priority species 

in the Arizona Partners in Flight Bird Conservation 

Plan (table 4-1).

Based on anecdotal evidence, it is likely that major 

declines in the abundance of Baird’s sparrows have 

occurred over a longer time frame, and much of this 

change took place between the late 1800s and early 

1900s before quantitative data were collected. In the 

late 1800s, for example, Coues (1878, as cited in Jones 

and Green 1998) referred to the Baird’s sparrow as one 

of the most abundant species in the Dakota Territory. 

Prior to 1880, Baird’s sparrows were reportedly so 

common in portions of North Dakota that they 

Figure 4-24. Breeding and wintering ranges of 

the Baird’s sparrow in Canada, the United States, 

and Mexico (from Jones and Green 1998).

Breeding

Wintering

North
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 125

outnumbered all other species of birds combined 

(Lane 1968). The presumably sharp drop in numbers 

of Baird’s sparrows on their breeding grounds in the 

northern prairies has been attributed primarily to the 

conversion of its native prairie habitat to agricultural 

uses and degradation of grassland habitat by overgrazing (Goossen and others 1993, Jones and Green 1998, 

Kantrud 1981, Owens and Myres 1973, Stewart 1975, 

Sutter and Brigham 1998). In Canada, for example, it 

is estimated that more than 90 percent of the native 

prairie has been converted or cultivated, and numbers 

of Baird’s sparrows have declined accordingly (DeSmet 

and Miller 1989). In Arizona, the primary threats to 

Baird’s sparrows are believed to be overgrazing and loss 

of habitat to residential development. In New Mexico, 

overgrazing and conversion of native grasslands to 

agricultural uses are believed to be the principal 

threats (Jones and Green 1998).

Preferred Habitat—Unfortunately, there has been 

little study of this species in its winter range. The 

following information on wintering Baird’s sparrows 

will be supplemented with information from studies 

on breeding birds where deemed appropriate and will 

be so noted.

Wintering Baird’s sparrows are most often found in 

expansive areas of relatively dense, tall grasses (Jones 

and Green 1998, NatureServe 2001). In Arizona, habitats used may be described as open semidesert or plains 

grasslands, dominated by perennial bunchgrasses in 

the genera Bouteloua and Eragrostris (Gordon 2000a). 

These grasslands also supported a variety of forbs and 

some small woody perennials, but woody plants less 

than 1 m in height were restricted to a few scattered 

mesquite trees (Prosopis velutina; Gordon 2000a). In 

New Mexico, Baird’s sparrows tend to be found primarily in mid-elevation grasslands at about 1,500 m, but on 

occasion have been found in mountain meadows up to 

about 3,600 m. In grasslands subject to grazing, Baird’s 

sparrows are found mostly in swales with taller grass 

or along ungrazed roadside edges (Jones and Green 

1998). On both the breeding and wintering grounds, 

Baird’s sparrows tend to prefer either ungrazed or 

only lightly grazed dense grasslands with relatively 

few shrubs (Jones and Green 1998). In the winter, 

Pulliam and Mills (1977) found that Baird’s sparrows 

were most abundant in open grasslands more than 

64 m from shrub cover. Preliminary results from a 

study of wintering birds in southeastern Arizona show 

a negative relationship between the abundance of 

Baird’s sparrows and shrub density (J. Ruth, personal 

communication 2002).

Some key characteristics of Baird’s sparrow breeding 

habitat are presented here based on the assumption 

that winter habitat preferences may be similar to 

some degree (Cartwright and others 1937, as cited 

in Jones and Green 1998). On the breeding range, 

Baird’s sparrows show a preference for native mixedgrass or tallgrass prairie and are found primarily in 

undisturbed or lightly grazed habitats characterized 

by dense grasses at least 10 to 20 cm in height and 

with a maximum of 20 to 25 percent shrub cover (Jones 

and Green 1998 and references therein, Sousa and 

McDonal 1983). In North Dakota, the highest density of 

Baird’s sparrows were associated with grasslands that 

had a relatively high but patchy coverage of forbs (20 

percent) and less than 10 percent woody cover (Winter 

1994, as cited in Jones and Green 1998). Traditionally 

Baird’s sparrows are believed to exhibit an aversion 

to exotic grasses, although this relationship seems to 

hold primarily for smooth brome Bromus inermis and 

shows more mixed results for species such as crested 

wheatgrass Agropyron cristatum (Jones and Green 

1998 and references therein). In Arizona, Gordon 

(2000a) found this species wintering in grasslands 

with a high proportion of the exotic Lehmann lovegrass 

Eragrostris lehmanniana. Gordon (personal communication as cited in Jones and Green 1998) speculates 

that the presence of exotic grasses may actually act to 

improve the suitability of some overgrazed habitats 

by providing the dense grass cover preferred by this 

species. Structural similarity of the vegetation may be 

more important in determining habitat quality for the 

Baird’s sparrow than the actual species composition 

(Jones and Green 1998).

Baird’s sparrows are sensitive to the structural 

changes in vegetation brought about by grazing. 

Habitats that are grazed for extended periods or 

subjected to excessive grazing pressure are rendered 

unsuitable for this species (Jones and Green 1998). 

Studies on the breeding grounds have shown that 

continuous grazing in relatively dry areas of mixedgrass prairie results in the virtual elimination of the 

Baird’s sparrow (Dale 1984, Karasiuk and others 1977 

as cited in Jones and Green 1998, Owens and Myres 

1973, Smith and Smith 1966). In the Southwest, overgrazing has probably degraded wintering habitat for 

Baird’s sparrows by promoting shrub encroachment 

on grasslands. Heavy grazing can also reduce habitat 

quality by decreasing grass density and height below 

the threshold preferred by this species. As wintering Baird’s sparrows depend primarily upon seeds 

for sustenance, levels of grazing that reduce seed 

production may also have severe negative impacts 

on the population (Latta and others 1999). Grazing 

need not be entirely incompatible with management 

for this sparrow on the wintering grounds, however. 

In Arizona, Gordon (2000b) found no significant difference between numbers of Baird’s sparrows on grazed 

and ungrazed grasslands; in one study year, Baird’s 

sparrows were actually more numerous on the grazed 

pasture. However, the grazed pasture in this case 

was stocked at well below the standard stocking rate 
126 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

in 2 years out of the 3 years of the study, and it was 

rated to be in excellent condition by the Bureau of Land 

Management (Gordon 2000b and references therein). 

The 1,501 ha pasture, with a stated desired stocking 

rate of 1,091 animal unit months (AUM), received 1,387, 

868, and 645 AUMs of grazing pressure respectively in 

the 3 years of the study, and grazing was concentrated 

primarily during the summer months. Gordon (2000b) 

is careful to point out that although this level of grazing 

did not appear to negatively impact Baird’s sparrows, 

this species was virtually eliminated from an adjacent 

pasture that received more intensive grazing pressure 

during the same period (actual stocking rates were not 

reported for this pasture).

Area Requirements—Although the actual extent 

of the area required is unknown, Baird’s sparrows 

are apparently quite sedentary in the winter, and 

individuals tend to remain within fixed home ranges 

once established on the wintering grounds (Gordon 

2000a). Arizona Partners in Flight recommends the 

conservation of areas of dense Grama and bunchgrasses 

a minimum of 2.5 ha within a greater matrix of mixed 

grass and shrub habitat (Latta and others 1999).

Timing of Migration— In Arizona and New 

Mexico, wintering Baird’s sparrows may arrive in 

early to mid-August. Most leave for their breeding 

grounds by late February, but some individuals may 

be present as late as mid-May (Jones and Green 1998, 

Terres 1991).

Management Issues and Recommendations—

Baird’s sparrows are vulnerable to habitat loss and 

fragmentation and show a preference for grasslands 

that are either of native species composition or at least 

composed of species that provide habitat structure and 

heterogeneity similar to that of native grasslands. 

Thus, the preservation or restoration, or both, of large 

expanses of native grasslands on the wintering range 

of the Baird’s sparrow is desirable for the conservation of this species. As Baird’s sparrow is relatively 

sedentary on the wintering grounds, it is especially 

vulnerable to the effects of fragmentation as it is 

unlikely to disperse far in search of higher quality 

habitat. Expansive core areas of contiguous grassland 

are required to sustain this species; an equal area of 

suitable grasslands that occur in the landscape as 

isolated fragments does not provide habitat of equivalent quality (Gordon 2000a). Conversion of grasslands 

to croplands probably represents a complete loss of 

usable habitat for this species and should be avoided 

(Goossen and others 1993).

Prescribed burning can be a potentially effective 

management tool for Baird’s sparrow on its wintering 

grounds because periodic fires will help to prevent shrub 

encroachment and thus prevent the high densities of 

woody plants that Baird’s sparrow avoids. Caution must 

be exercised, however, in tailoring the burning regime 

to the specific physical and vegetative characteristics of 

the site and in attempting to duplicate the natural fire 

frequency of the area (Jones and Green 1998, Madden 

1996, as cited in Dechant and others 2001g, Winter 

1999). If fire is used in extensive grassland areas, it 

may be helpful to burn small areas on a rotational 

basis to create a mosaic of successional stages (Johnson 

1997, Madden 1996, as cited in Dechant and others 

2001g, Renken and Dinsmore 1987). Alternatively, 

woody plants may be controlled in grasslands used 

by wintering Baird’s sparrows through mechanical 

or chemical means.

Baird’s sparrow appears to be tolerant of light to 

moderate levels of grazing. However, as this species 

prefers relatively dense grasses of moderate height, 

continuous grazing regimes or grazing at high levels 

of intensity are likely to result in the elimination of 

appropriate habitat for Baird’s sparrow. One study in 

southern Arizona found the continued presence of wintering Baird’s sparrows on pastures grazed primarily in 

the summer at a level of 0.4 to 0-.9 AUMs per hectare 

(Gordon 2000b). Long duration and/or heavy grazing 

levels should be avoided in areas of prime habitat for 

the Baird’s sparrow (Latta and others 1999).

In addition to controlling shrub encroachment, 

perhaps the most important overall aspect of managing for Baird’s sparrows in the Southwest is the 

implementation of range management practices that 

provide adequate cover for the birds and allow for 

greater levels of seed production to sustain wintering 

populations of this species (NMDGF 1988).

Associated Species—Northern aplomado falcon, 

horned lark, Sprague’s pipit, savannah sparrow, grasshopper sparrow, chestnut-collared longspur, eastern 

meadowlark (Latta and others 1999, NMPIF 2001).

Acknowledgments_________________

Sincere thanks to Carl Bock, Jane Bock, and Janet 

Ruth for their excellent and insightful reviews; their 

comments helped to greatly improve the quality of this 

chapter. Bill Howe, Christopher Rustay, and Sartor O. 

Williams III were all generous in sharing their knowledge with me. Additional thanks to Troy Corman, Jeff 

Kelly, Fritz Knopf, Janet Ruth, and Pat Zwartjes for 

providing information and resources to assist in my 

research. The species accounts for the rufous-winged 

sparrow and scaled quail presented here were adapted 

from those originally written for the Rocky Mountain 

Research Station’s Ecological Interactions of Ungulate 

Grazing and Native Species of the Southwest: Terrestrial 

Wildlife (Zwartjes and others 2005). Authors of the 

original accounts were Walter Haussamen (scaled 

quail) and Michele Merola-Zwartjes (rufous-winged 

sparrow). Any errors or omissions remaining in this 

manuscript are the sole responsibility of the author.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 127

References________________________

Agnew, W.D., W. Uresk, and R.M. Hansen. 1986. Flora and fauna 

associated with prairie dog colonies and adjacent ungrazed 

mixed-grass prairie in western South Dakota. Journal of Range 

Management. 39:135-139.

Aldrich, J.W. and C.S. Robbins. 1970. Changing abundance of 

migratory birds in North America. Smithsonian Contributions 

to Zoology. 26:17-26.

Allen, J.N. 1980. The ecology and behavior of the long-billed curlew 

in southeastern Washington. Wildlife Monographs. 73.

Ambuel, B. and S.A. Temple. 1982. Songbird populations in southern 

Wisconsin forests: 1954 and 1979. Journal of Field Ornithology. 

53:149-158.

American Bird Conservancy. 2001. Pesticides and Birds Campaign. 

Available at http://www.abcbirds.org/pesticides/pesticideindex.

htm. Accessed October 2001. 

American Ornithologists’ Union. 1931. Check-list of North American 

birds, 4th edition. american Ornithologists’ Union, Lancaster, 

Pennsylvania.

American Ornithologists’ Union. 1998. Check-list of North American 

Birds.Allen Press, Lawrence, Kansas.

Anable, M.E., M.P. McClaran, and G.B. Ruyle. 1992. Spread of 

introduced Lehmann lovegrass (Eragrostis lehmanniana Nees.) in 

southern Arizona, USA. Biological Conservation. 61:181-188.

Anderson, A.H. 1965. Notes on the behavior of the rufous-winged 

sparrow. Condor. 67:188-190.

Anderson, D.W. and J.J. Hickey. 1972. Eggshell changes in certain 

North American birds. Pages 514-540 in Proceedings of the 25th 

International Ornithological Congress.

Anderson, E., S.C. Forrest, T.W. Clark, and L. Richardson. 1986. 

Paleobiology, biogeography, and systematics of the black-footed 

ferret, Mustela nigripes (Audubon and Bachman), 1851.Great 

Basin Naturalist Memoirs. 8:11-62.

Anderson, W.L. and R.E. Duzan. 1978. DDE residues and eggshell 

thinning in Loggerhead Shrikes. Wilson Bulletin. 90:215-220.

Andrews, R. and R. Righter. 1992. Colorado birds. Denver Museum 

of Natural History, Denver, Colorado.

Anstey, D.A., S.K. Davis, D.C. Duncan, and M. Skeel. 1995. 

Distribution and habitat requirements of eight grassland songbird 

species in southern Saskatchewan. Unpublished report submitted 

to the Saskatchewan Wetland Conservation Corporation, Regina, 

Saskatchewan, Canada.

Applegate, R.D. and T.Z. Riley. 1998. Lesser Prairie-chicken 

Management. Rangelands. 20:13-15.

Archer, S. 1989. Have southern Texas savannas been converted to 

woodlands in recent history? American Naturalist. 134:545-561.

Arizona Game and Fish Department. 1988. Threatened native wildlife in Arizona. Arizona Game and Fish Department Publication, 

Phoenix, Arizona.

Askins, R.A. 1993. Population trends in grassland, shrubland, and 

forest birds in eastern North America. Current Ornithology. 

11:1-34.

Askins, R.A., J.F. Lynch, and R. Greenberg. 1990. Population 

declines in migratory birds in eastern North America. Current 

Ornithology 7, 1-57. 1990. 

Aspelin, A.L. and A.H. Grube. 1999. Pesticides industry sales and 

usage: 1996 and 1997 market estimates. U.S. Environmental 

Protection Agency Publication No. 733-R-99-001, Washington, 

D.C.

Bahre, C.J. 1987. Wild hay harvesting in southern Arizona: a casualty 

of the march of progress. Journal of Arizona History. 28:69-78.

Bahre, C.J. 1991. A legacy of change: historic human impact on 

vegetation in the Arizona borderlands. University of Arizona 

Press, Tucson, Arizona.

Bahre, C.J. 1995. Human impacts on the grasslands of southeastern Arizona. Pages 230-264. In: M.P. McClaran and T.R. Van 

Devender, eds. The Desert Grassland. The University of Arizona 

Press, Tucson, Arizona.

Baicich, P.J. and C.J.O. Harrison. 1997. A guide to the nests, eggs, 

and nestlings of North American birds. Academic Press, New 

York, New York.

Bailey, F.M. 1928. Birds of New Mexico. New Mexico Department 

of Game and Fish, Santa Fe, New Mexico.

Bailey, J.A. 1999. Status and trend of the lesser prairie-chicken in 

New Mexico and recommendations to list the species as threatened 

under the New Mexico Wildlife Conservation Act. Unpublished 

report submitted to the New Mexico Department of Game and 

Fish, Santa Fe, New Mexico.

Bailey, J.A., J. Klingel, and C.A. Davis. 2000.Status of nesting 

habitat for lesser prairie-chicken in New Mexico.The Prairie 

Naturalist. 32:149-156.

Bailey, J.A. and S.O. Williams III. 2000. Status of the lesser 

prairie-chicken in New Mexico, 1999. The Prairie Naturalist. 

32:157-168.

Bailey, V. 1932. Mammals of New Mexico. North American Fauna. 

53:119-131.

Balcomb, R. 1986. Songbird carcasses disappear rapidly from 

agricultural fields. Auk. 103:817-820.

Baril, A. 1993. Pesticides and wildlife in the prairies: current 

regulatory issues. Pages 44-48. In: G.L. Holroyd, H.L. Dickson, 

M. Regnier, and H.C. Smith, eds. Proceedings of the third endangered species and prairie conservation workshop. Provincial 

Museum of Alberta, Natural History Occasional Paper No. 19, 

Edmonton, Alberta.

Bartelt, G.A. 1997. Improving habitat quality of rotationally grazed 

pastures for grassland birds. Final report. Wisconsin Department 

of Natural Resources, Monona, Wisconsin.

Basili, G.D. and S.A. Temple. 1999. Winter ecology, behavior, and 

conservation needs of Dickcissels in Venezuela. Studies in Avian 

Biology. 19:289-299.

Beason, R.C. 1995. Horned Lark (Eremophila alpestris). In: A. 

Poole and F. Gill, eds. The Birds of North America No. 195. The 

Academy of Natural Sciences, Philadelphia, and The American 

Ornithologists’ Union, Washington, D.C 

Beason, R.C. and E.C. Franks. 1973. Development of young Horned 

Larks. Auk. 90:359-363.

Bechard, M.J. and J.K. Schmutz. 1995. Ferruginous Hawk Buteo 

regalis. In: A. Poole and F. Gill, eds. The Birds of North America, 

No. 172. The Academy of Natural Sciences, Philadelphia, and The 

American Ornithologists’ Union, Washington, D.C.

Beintema, A.J. and G.J.D.M. Muskens. 1987. Nesting success of 

birds breeding in Dutch agricultural grasslands. Journal of 

Applied Ecology. 24:743-758.

Bell, W.R. 1921. Death to the rodents. U.S. Department of Agriculture 

Yearbook. 1921:421-438.

Bent, A.C. 1932. Life histories of North American gallinaceous birds. 

United States National Museum Bulletin No. 162.

Bent, A.C. 1937. Life histories of North American birds of prey, Part 

I. United States National Museum Bulletin No. 167.

Bent, A.C. 1938. Life histories of North American birds of prey, Part 

II. United States National Museum Bulletin No. 170.

Bent, A.C. 1950. Sprague’s pipit. United States National Museum 

Bulletin. 197:52-62.

Bent, A.C. 1962. Life histories of North American shorebirds. Dover, 

New York, New York.

Berkey, G., R. Crawford, S. Galipeau, D. Johnson, D. Lambeth, 

and R. Kreil. 1993. A review of wildlife management practices in 

North Dakota: effects on nongame bird populations and habitats. 

Report submitted to U.S. Fish and Wildlife Service, Region 6, 

Denver Colorado.

Berry, M.E., C.E. Bock, and S.L. Haire. 1998. Abundance of diurnal 

raptors on open space grasslands in an urbanized landscape. 

Condor. 100:601-608.

Best,L.B. 1978. Field sparrow reproductive success and nesting 

ecology. Auk. 95:9-22.

Best, L.B., H. Campa III, K.E. Kemp, R.J. Robel, M.R. Ryan, J.A. 

Savidge, H.P. Weeks, Jr., and S.R. Winterstein. 1997. Bird abundance and nesting in CRP fields and cropland in the Midwest: a 

regional approach. Wildlife Society Bulletin. 25:864-877.

Best, T.L. and R.A. Smartt. 1985. Foods of scaled quail in southeastern 

New Mexico. Texas Academy of Sciences. 37:155-162.

Bicak, T.K. 1983. Vegetative interference: a factor affecting longbilled curlew (Numenius americanus) foraging success. Wader 

Study Group Bulletin. 39:57.

Biack, T.K., R.L. Redmond, and D.A. Jenni. 1982. Effects of grazing 

on long-billed curlew (Numenius americanus) breeding behavior 

and ecology in southwestern Idaho. Pages 74-85. In: J. eek and 

P.D. Dalke, eds. Wildlife-livestock relationships symposium: 
128 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Proceedings 10. University of Idaho, Forest, Wildlife and Range 

Experiment Station, Moscow, Idaho.

Bock, C.E. and J.H. Bock. 1987.Avian habitat occupancy following 

fire in a Montana shrubsteppe. Prairie Naturalist. 19:153-158.

Bock, C.E. and J.H. Bock. 1988. Grassland birds in Southeastern 

Arizona: impacts of fire, grazing, and alien vegetation. Pages 

43-58. In: P.D. Goriup, ed. Ecology and conservation of grassland 

birds. International Council for Bird Preservation, Publication 

No. 7, Cambridge, England.

Bock, C.E. and J.H. Bock. 1992. Response of birds to wildfire in 

native versus exotic Arizona grassland. Southwestern Naturalist. 

37:73-81.

Bock, C.E. and J.H. Bock. 1999. Response of winter birds to drought 

and short-duration grazing in southeastern Arizona. Conservation 

Biology. 13:1117-1123.

Bock, C.E., J.H. Bock, and B.C. Bennett. 1999. Songbird abundance 

in grasslands at a suburban interface on the Colorado High Plains. 

Studies in Avian Biology. 19:131-136.

Bock, C.E., J.H. Bock, K.L. Jepson, and J.L. Ortega. 1986. Ecological 

effects of planting African lovegrasses in Arizona. National 

Geographic Research. 2:456-463.

Bock, C.E., J.H. Bock, W.R. Kenney, and V.M. Hawthorne. 1984. 

Responses of birds, rodents, and vegetation to livestock exclosure 

in a semidesert grassland site.Journal of Range Management. 

37:239-242.

Bock, C.E., V.A. Saab, T.D. Rich, and D.S. Dobkin. 1993. Effects of 

livestock grazing on Neotropical migratory landbirds in western 

North America. Pages 296-309. In: D.M. Finch and P.W. Stangel, 

eds. Status and management of Neotropical migratory birds. 

USDA Forest Service, General Technical Report RM-229.

Bock, C.E. and B. Webb. 1984. Birds as grazing indicator species in southeastern Arizona.Journal of Wildlife Management. 

48:1045-1049.

Bock, J.H., C.E. Bock, and J.R. McKnight. 1976. A study of the effects 

of grassland fires at the Research Ranch in southeastern Arizona. 

Journal of the Arizona Academy of Sciences. 11:49-57.

Bollinger, E.K., P.B. Bollinger, and T.A. Gavin. 1990. Effects of 

hay-cropping on eastern populations of the Bobolink. Wildlife 

Society Bulletin. 18:142-150.

Botelho, E.S. and P.C. Arrowood. 1996.Nesting success of western 

burrowing owls in natural and human-altered environments 

Pages 61-68. In: D.M. Bird, D.E. Varland, and J.J. Negro, eds. 

Raptors in human landscapes: adaptations to built and cultivated 

environments. Academic Press, San Diego, California.

Boyd, C.S. and T.S. Bidwell. 2001. Influence of prescribed fire on 

lesser prairie-chicken habitat in shinnery oak communities in 

western Oklahoma. Wildlife Society Bulletin. 29:938-947.

Briggs, S.A. and J.H. Criswell. 1979. Gradual silencing of spring in 

Washington. Atlantic Naturalist. 32:19-26.

Brittingham, M.C. and S.A. Temple. 1983. Have cowbirds caused 

forest songbirds to decline? BioScience. 33:31-35.

Brooks, B.L. and S.A. Temple. 1990. Habitat availability and suitability for Loggerhead Shrikes in the upper Midwest. American 

Midland Naturalist. 123:75-83.

Brown, D.E. 1982. Grasslands. Desert Plants. 4:107-141.

Brown, D.E. 1989. Arizona game birds. University of Arizona Press, 

Tucson, Arizona.

Bryan, G.G. and L.B. Best. 1994. Avian nest density and success 

in grassed waterways in Iowa rowcrop fields. Wildlife Society 

Bulletin. 22:583-592.

Buffington, L.D. and C.H. Herbel. 1965. Vegetation changes on 

semidesert grassland range from 1858 to 1963. Ecological 

Monographs. 35:139-164.

Burger, L.D., L.W. Burger, Jr., and J. Faaborg. 1994. Effects of 

prairie fragmentation on predation on artificial nests. Journal 

of Wildlife Management. 58:249-254.

Busby, D.G., L.M. White, and P.A. Pearce. 1990. Effects of aerial 

spraying of fenitrothion on breeding White-throated Sparrows.

Journal of Applied Ecology. 27:743-755.

Butts, K.O. 1973. Life history and habitat requirements of Burrowing 

Owls in western Oklahoma. MS thesis. Stillwater, Oklahoma, 

Oklahoma State University. 

Butts, K.O. and J.C. Lewis. 1982. The importance of prairie dog towns 

to Burrowing Owls in Oklahoma. Proceedings of the Oklahoma 

Academy of Science. 62:46-52.

Cade, T.J. and C.P. Woods. 1994. Changes in distribution and abundance of loggerhead shrikes. Journal of Ornithology. 135:288.

Campbell, H. 1952. Habitat improvement for upland game birds 

in New Mexico. Proceedings of the Western Association of Game 

and Fish Commissions. 32:115-118.

Campbell, H. 1968. Seasonal precipitation and scaled quail in eastern 

New Mexico. Journal of Wildlife Management. 32:641-644.

Campbell, H. 1972. A population study of Lesser Prairie-chickens in 

New Mexico. Journal of Wildlife Management. 36:689-699. 

Campbell, H., D.K. Martin, P.E. Ferkovich, and B.K. Harris. 1973. 

Effects of hunting and some other environmental factors on scaled 

quail in New Mexico. Wildlife Monographs 34.

Campbell-Kissock, L., L.H. Blankenship, and L.D. White. 1984. 

Grazing management impacts on quail during drought in the 

northern Rio Grande Plain, Texas. Journal of Range Management. 

37:442-446.

Cannings, R.J. and W. Threlfall. 1981. Horned Lark Eremophila 

alpestris breeding biology at Cape St. Marys, Newfoundland, 

Canada. Wilson Bulletin. 93:519-530.

Cannon, R.W. and F.L. Knopf. 1981. Lek numbers as a trend index 

to prairie grouse populations. Journal of Wildlife Management. 

45:776-778.

Carson, R. 1962. Silent spring. Houghton Mifflin, Boston, 

Massachusetts.

Carter, M.F. and K. Barker. 1993. An interactive database for setting conservation priorities for western neotropical migrants. 

Pages 120-144. In: D.M. Finch and P.W. Stangel, eds. Status 

and management of neotropical migratory birds. USDA Forest 

Service, Rocky Mountain Forest and Range Experiment Station, 

GTR-RM-229, Fort Collins, Colorado. 

Clayton, K.M. and J.K. Schmutz. 1999. Is the decline of Burrowing 

Owls Speotyto cunicularia in prairie Canada linked to changes 

in Great Plains ecosystems? Bird Conservation International. 

9:163-185.

Cochran, J.F. and S.H. Anderson. 1987. Comparison of habitat 

attributes at sites of stable and declining long-billed curlew 

populations. Great Basin Naturalist. 47:459-466.

Collar, N.J., L.P. Gonzaga, N. Krabbe, A. Madrono Nieto, L.G. 

Naranjo, T.A. Parker III, and D.C. Wege. 1992. Threatened birds of 

the Americas. Smithsonian Institution Press, Washington, D.C.

Collins, C.T. and R.E. Landry. 1977. Artificial nest burrows for 

Burrowing Owls. North American Bird Bander. 2:151-154.

Cooke, W.W. 1915. Our shorebirds and their future. Condor. 

17:237-238.

Copelin, F.F. 1963. The lesser prairie-chicken in Oklahoma. Oklahoma 

Wildlife Conservation Department Technical Bulletin No. 6.

Coppock, D.L., J.K. Detling, J.E. Ellis, and M.I. Dyer. 1983. Plantherbivore interactions in a North American mixed-grass prairie 

II. Responses of bison to modification of vegetation by prairie 

dogs. Oecologia. 56:10-15.

Cox, J.R., H.L. Morton, T.N. Johnsen, Jr., G.L. Jordan, S.C. Martin, 

and L.C. Fierro. 1984. Vegetation restoration in the Chihuahuan 

and Sonoran Deserts of North America. Rangelands. 6:112-115.

Cox, J.R., G.B. Ruyle, and B.A. Roundy. 1990. Lehmann lovegrass 

in southeastern Arizona: biomass production and disappearance. 

Journal of Range Management. 43:367-372.

Craig, R.B. 1978. An analysis of the predatory behavior of the 

Loggerhead Shrike.Auk. 95:221-234.

Crawford, J.A. 1980. Status, problems, and research needs of the 

Lesser Prairie-chicken. Pages 1-7 in P.A. Vohs and F.L. Knopf, 

eds. Proceedings of the Prairie Grouse Symposium. Oklahoma 

State University, Stillwater, Oklahoma.

Cully, J.F. 1989. Plague in prairie dog ecosystems: importance for 

black-footed ferret management. Pages 47-55. In: T.W. Clar, D. 

Hinckley, and T. Rich, eds. The prairie dog ecosystem: managing 

for biological diversity. Montana Bureau of Land Management 

Wildlife Technical Bulletin No. 2.

Cully, J.F. 1991. Response of raptors to reduction of a Gunnison’s 

prairie dog population by plague. American Midland Naturalist. 

125:140-149.

Dale, B.C. 1984. Birds of grazed and ungrazed grasslands in 

Saskatchewan. Blue Jay. 42:102-105.

Dale, B.C., P.A. Martin, and P.S. Taylor. 1997. Effects of hay management on grassland songbirds in Saskatchewan. Wildlife Society 

Bulletin. 25:616-626.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 129

Daubenmire, R. 1970. Steppe vegetation of Washington. Washington 

Agricultural Experiment Station Technical Bulletin. 62.

Davis, C.A., R.C. Barkley, and W.C. Haussamen. 1975. Scaled 

quail foods in southeastern New Mexico. Journal of Wildlife 

Management. 39:496-502.

Davis, C.A., T.Z. Riley, R.A. Smith, H.R. Suminski, and M.J. Wisdom. 

1979. Habitat evaluation of Lesser Prairie-chickens in eastern 

Chaves County, New Mexico. New Mexico Agricultural Experiment 

Station, Las Cruces, New Mexico.

Davis, S.K. and D.C. Duncan. 1999. Grassland songbird abundance in 

native and crested wheatgrass pastures of southern Saskatchewan. 

Studies in Avian Biology. 19:211-218.

Day, K.S. 1994. Observations on Mountain Plovers (Charadrius montanus) breeding in Utah. Southwestern Naturalist. 39:298-300.

Dechant, J.A., M.L. Sondreal, D.H. Johnson, L.D. Igl, C.M. Goldade, 

P.A. Rabie, and B.R. Euliss. 2001a. Effects of management practices on grassland birds: Ferruginous Hawk. Northern Prairie 

Wildlife Research Center, http://www.npwrc.usgs.gov/resource/

literatr/grasbird/ferhawk/ferhawk.htm (Version 17FEB2000), 

Jamestown, North Dakota.

Dechant, J.A., M.L. Sondreal, D.H. Johnson, L.D. Igl, C.M. Goldade, 

M.P. Nenneman, and B.R. Euliss. 2001b. Effects of management 

practices on grassland birds: Mountain Plover. Northern Prairie 

Wildlife Research Center, available at http://www.npwrc.usgs.

gov/resource/literatr/grasbird/mtnplove/mtnplove.htm (Version 

17FEB2000), Jamestown, North Dakota.

Dechant, J.A., M.L. Sondreal, D.H. Johnson, L.D. Igl, C.M. Goldade, 

P.A. Rabie, and B.R. Euliss. 2001c. Effects of management 

practices on grassland birds: Long-billed Curlew. Northern 

Prairie Wildlife Research Center, available at http://www.npwrc.

usgs.gov/resource/literatr/grasbird/fplbcu/fplbcu.htm (Version 

29FEB2000), Jamestown, North Dakota.

Dechant, J.A., M.L. Sondreal, D.H. Johnson, L.D. Igl, C.M. Goldade, 

P.A. Rabie, and B.R. Euliss. 2001d. Effects of management 

practices on grassland birds: Burrowing Owl. Northern Prairie 

Wildlife Research Center http://www.npwrc.usgs.gov/resource/literatr/grasbird/buow/buow.htm (Version 17FEB2000), Jamestown, 

North Dakota.

Dechant, J.A., M.L. Sondreal, D.H. Johnson, L.D. Igl, C.M. Goldade, 

M.P. Nenneman, A.L. Zimmerman, and B.R. Euliss. 2001e. Effects 

of management practices on grassland birds: Loggerhead Shrike. 

Northern Prairie Wildlife Research Center, http://www.npwrc.

usgs.gov/resource/literatr/grasbird/logger/logger.htm (Version 

17FEB2000), Jamestown, North Dakota.

Dechant, J.A., M.L. Sondreal, D.H. Johnson, L.D. Igl, C.M. Goldade, 

M.P. Nenneman, and B.R. Euliss. 2001f. Effects of management 

practices on grassland birds: Sprague’s Pipit. Northern Prairie 

Wildlife Research Center, available at http://www.npwrc.usgs.gov/

resource/literatr/grasbird/pipit/pipit.htm (Version 17FEB2000), 

Jamestown, North Dakota.

Dechant, J.A., M.L. Sondreal, D.H. Johnson, L.D. Igl, C.M. Goldade, 

M.P. Nenneman, and B.R. Euliss. 2001g. Effects of management 

practices on grassland birds: Baird’s Sparrow. Northern Prairie 

Wildlife Research Center, http://www.npwrc.usgs.gov/resource/

literatr/grasbird/bairds/bairds.htm (Version 17FEB2000), 

Jamestown, North Dakota.

DeSante, D.F. and T.L. George. 1994. Population trends in the 

landbirds of Western North America. Studies in Avian Biology. 

15:173-190.

DeSmet, K.D. and M.P. Conrad. 1991. Management and research 

needs for Baird’s Sparrows and other grassland species in 

Manitoba. Pages 83-86.In: G.L .Holroyd, G. Burns, and H.C. 

Smith, eds. Proceedings of the second endangerd species and 

prairie conservation workshop. Natural History Occasional Paper 

15, Provincial Museum of Alberta, Edmonton, Alberta.

DeSmet, K.D. and W.S. Miller. 1989. Status report on the Baird’s 

Sparrow, Ammodramus bairdii. Report to the Committee on the 

Status of Endangered Wildlife in Canada, Ottawa, Ontario.

Desmond, M. J. 1991. Ecological aspects of Burrowing Owl nesting 

strategies in the Nebraska panhandle. M.S. thesis. University of 

Nebraska, Lincoln, Nebraska. 114 p.

Desmond, M.J. and J.A. Savidge. 1996. Factors influencing Burrowing 

Owl (Speotyto cunicularia) nest densities and numbers in western 

Nebraska. American Midland Naturalist. 136:143-148.

Desmond, M.J. and J.A. Savidge. 1998. Burrowing Owl conservation 

in the Great Plains. Page 9 In: Abstracts of the Second International 

Burrowing Owl Symposium. Ogden, Utah.

Desmond, M.J. and J.A. Savidge. 1999. Satellite burrow use by 

Burrowing Owl chicks and its influence on nest site. Studies in 

Avian Biology. 19:128-130.

Desmond, M.J., J.A. Savidge, and K.M. Eskridge. 2000. Correlations 

between Burrowing Owl and Black-tailed Prairie Dog declines: a 7-

year analysis. Journal of Wildlife Management. 64:1067-1075.

Desmond, J.J., J.A. Savidge, and T.F. Seibert. 1995. Spatial patterns 

of Burrowing Owl (Speotyto cunicularia) nests within black-tailed 

prairie dog (Cynomys ludovicianus) towns. Canadian Journal of 

Zoology. 73:1375-1379.

DeWeese, L.R., C.J. Henny, R.L. Floyd, K.A. Bobal, and A.W. Schultz. 

1979. Response of breeding birds to aerial sprays of trichlorfon 

(Fylox) and carbaryl (Sevin-4-oil) in Montana forests. U.S. Fish and 

Wildlife Service, Special Science Report on Wildlife. 224:1-29.

DeWeese, L.R., L.C. McEwen, G.L. Hensler, and B.E. Petersen. 

1986. Organochlorine contamination in passeriforms and other 

avian prey of the Peregrine falcon in the western United States. 

Environmental Toxicology and Chemistry. 5:675-693.

DeWeese, L.R., L.C. McEwen, L.A. Settimi, and R.D. DeBlinger. 

1983. Effects on birds of Fenthion aerial application for mosquito 

control. Journal of Economic Entomology. 76:906-911.

Dick-Peddie, W.A. 1993. New Mexico vegetation: past, present, future. 

University of New Mexico Press, Albuqueruqe, New Mexico.

Dinkins, M.F., A.L. Zimmerman, J.A. Dechant, B.D. Parkin, D.H. 

Johnson, L.D. Igl, C.M. Goldade, and B.R. Euliss. 2000. Effects 

of management practices on grasslands birds: Horned Lark. 

Northern Prairie Wildlife Research Center, Jamestown, North 

Dakota. Northern Prairie Wildlife Research Center Home Page, 

available at http://www.npwrc.usgs.gov/resource/literatr/grasbird/hola/hola.htm (Version 16JUN2000).

DiSilvestro, R. 1996. Poison in the Pampas: what’s killing the 

Swainson’s hawk? International Wildlife. 26:38-43.

Doerr, T.B. and F.S. Guthery. 1980. Effects of shinnery oak control 

on Lesser Prairie-chicken habitat.Pages 59-63.In: P.A.Vohs and 

F.L.Knopf, eds.Proceedings of the Prairie Grouse Symposium.

Stillwater, Oklahoma.

Duebbert, H.F. and J.T. Lokemoen. 1977. Upland nesting of 

American bitterns, marsh hawks, and short-eared owls. Prairie 

Naturalist. 9:33-40.

Dundas, H. and J. Jensen. 1995. Burrowing owl status and conservation. Bird Trends (Canadian Wildlife Service). 4:21-22.

Dunning, J.B. Jr., R.K. Bower Jr., S.J. Suter, and C.E. Bock. 1999. 

Cassin’s Sparrow (Aimophila cassinii). In: A. Poole and F. Gill, 

eds. The Birds of North America, No. 471. The National Academy 

of Sciences, Philadelphia, Pennsylvania, and The American 

Ornithologists’ Union, Washington, D.C.

Durrant, S.D.D. 1970. Faunal remains as indicators of Neothermal 

climates at Hogup Cave. Appendix II. Pages 241-245. n: C.M. 

Aikens, ed. University of Utah Anthropology Papers, No. 93.

Dyer, M.I. 1979. Consumers. Pages 73-86. In: R.T. Coupland, ed. 

Grassland ecosystems of the world: analysis of grasslands and 

their uses. Cambridge University Pres, Cambridge, England.

Ehrlich, P.R., D.S. Dobkin, and D. Wheye. 1988. The Birder’s 

Handbook. Simon and Schuster, Inc., New York, New York.

Ellison-Manning, A.E. and C.M. White. 2001. Breeding biology of 

Mountain Plovers (Charadrius montanus) in the Uinta Basin. 

Western North American Naturalist 61:223-228.

Emlen, J.T. 1972. Size and structure of a wintering avian community 

in southern Texas. Ecology. 53:317-329.

Environmental Protection Agency (EPA). 1989. Carbofuran: a special 

review technical support document. PA Office of Pesticides and 

Toxic Substances, Washington, D.C.

Faanes, C.A. 1983. Breeding birds of wooded draws in western 

North Dakota. Prairie Naturalist. 15:173-187.

Faanes, C.A. and G.R. Lingle. 1995. Breeding birds of the Platte River 

Valley of Nebraska. Northern Prairie Wildlife Research Center 

http://www.npwrc.usgs.gov/resource/distr/birds/platte/platte.htm 

(Version 16JUL97), Jamestown, North Dakota.

Finch, D.M., S.H. Anderson, and W.A. Hubert. 1987. Habitat suitability index models: lark bunting. U.S. Fish and Wildlife Service, 

Biological Report 82(10.137), Fort Collins, Colorado.
130 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Findley, J.S. 1987. The natural history of New Mexican mammals. 

University of New Mexico Press, Albuquerque, New Mexico.

Finzel, J.E. 1964. Avian populations of four herbaceous communities 

in southeastern Wyoming. Condor. 66:496-510.

Fleischner, T.L. 1994. Ecological costs of livestock grazing in western 

North America. Conservation Biology. 8:629-644.

Forbush, E.H. 1912. A history of game birds, wild-fowl and shore 

birds of Massachusetts and adjacent states. Massachusetts State 

Board of Agriculture, Boston, Massachusetts.

Fox, G.A., P. Mineau, B. Collins, and P.C. James. 1989. The impact 

of insecticide carbofuran (Furadan 480F) on the Burrowing Owl 

in Canada. Canadian Wildlife Service, Technical Report Series 

No. 72, Ottawa, Canada.

Frawley, B.J. and L.B. Best. 1991. Effects of mowing on breeding 

bird abundance and species composition in alfalfa fields. Wildlife 

Society Bulletin. 19:135-142.

Friedmann, H. 1929. The cowbirds: a study in the biology of social 

parasitism. C. Thomas, Springfield, Illinois.

Friedmann, H. 1963. Host relations of the parasitic cowbirds. U.S. 

National Museum Bulletin 233.

Friedmann, H. and L.F. Kiff. 1985. The parasitic cowbirds and 

their hosts. Proceedings of the Western Foundation of Zoology

2:226-304.

Gard, N.W. and M.J. Hooper. 1995. An assessment of potential 

hazards of pesticides and environmental contaminants. Pages 294-

310. In: T.E. Martin and D.M. Finch, eds. Ecology and management 

of neotropical migratory birds: a synthesis and review of critical 

issues. Oxford University Press, New York, New York.

Gates, J.E. and L.W. Gysel. 1978. Avian nest dispersion and fledgling 

success in field-forest ecotones. Ecology. 59:871-883.

Giesen, K.M. 1994. Movements and nesting habitat of Lesser Prairiechicken hens in Colorado. Southwestern Naturalist. 39:96-98.

Giesen, K.M. 1998. Lesser Prairie-Chicken (Tympanuchus pallidicinctus). In: A. Poole and F. Gill, eds. The Birds of North 

America, No. 364. The Academy of Natural Sciences, Philadelphia, 

Pennsylvania, and The American Ornithologists’ Union, 

Washington, D.C.

Giesen, K.M. 2000. Population status and management of lesser prairie-chicken in Colorado. The Prairie Naturalist. 32:137-148.

Gillihan, S.W., D.J. Hanni, S.W. Hutchings, T. Toombs, and T. 

VerCauteren. 2001. Sharing your land with shortgrass prairie 

birds. Rocky Mountain Bird Observatory, Brighton, Colorado.

Gillihan, S.W. and S.W. Hutchings. 2000. Best management practices 

for shortgrass prairie birds: a landowner’s guide. Colorado Bird 

Observatory, Brighton, Colorado.

Gleason, R.S. and D.R. Johnson. 1985. Factors influencing nesting 

success of Burrowing Owls in southeastern Idaho. Great Basin 

Naturalist. 45:81-84.

Glendening, G.E. 1952. Some quantitative data on the increase of 

mesquite and cactus on a desert grassland range in southern 

Arizona.Ecology. 33:57-74.

Glinski, R.L. 1998. Ferruginous Hawk. Pages 105-108. In: R.L. 

Glinski, ed. The raptors of Arizona. University of Arizona Press, 

Tucson, Arizona.

Goodwin, J.G. Jr. and C.R. Hungerford. 1977. Habitat used by native Gambel’s and Scaled quail and released masked bobwhite 

quail in southern Arizona. USDA Forest Service Research Paper 

RM-197, Rocky Mountain Forest and Range Experiment Station, 

Fort Collins, Colorado.

Goossen, J.P., S. Brechtel, K.D. DeSmet, D. Hjertass, and C. 

Werschler. 1993. Canadian Baird’s Sparrow recovery plan. 

Recovery of Nationally Endangered Wildlife, Report No. 3. 

Canadian Wildlife Federation, Ottawa, Ontario.

Gordon, C.E. 2000a. Movement patterns of wintering grassland 

sparrows in Arizona. Auk. 117:748-759.

Gordon, C.E. 2000b. Fire and cattle grazing on wintering sparrows in Arizona grasslands. Journal of Range Management. 

53:384-389.

Graber, R.R., J.W. Graber, and E.L. Kirk. 1973. Illinois birds: 

Laniidae. Illinois Natural History Survey Biological Notes. 

83:1-18.

Grant, R.A. 1965. The Burrowing Owl in Minnesota. Loon. 37:

2-17.

Graul, W.D. 1975. Breeding biology of the Mountain Plover. Wilson 

Bulletin. 87:6-31.

Graul, W.D. and L.E. Webster. 1976. Breeding status of the Mountain 

Plover. Condor. 78:265-267.

Green, G.A. and R.G. Anthony. 1989. Nesting success and habitat 

relationships of Burrowing Owls in the Columbia Basin, Oregon. 

Condor. 91:347-354.

Green, G. A., and R. G. Anthony. 1997. Ecological considerations for 

management of breeding burrowing owls in the Columbia Basin. 

Pages 117-121 in J. L. Lincer and K. Steenhof, eds. The Burrowing 

Owl, its biology and management: including the Proceedings of 

the First International Symposium. Raptor Research Report 

Number 9.

Grue, C.E., W.J. Fleming, D.G. Busby, and E.F. Hill. 1983. Assessing 

hazards of organophosphate pesticides to wildlife. Transactions 

of the North American Wildlife Natural Resources Conference. 

48:200-220.

Grue, C.E., P.L. Gilbert, and M.E. Seeley. 1997. Neurophysiological 

and behavioral changes in non-target wildlife exposed to organophosphate and carbamate pesticides: Thermoregulation, 

food consumption, and reproduction. American Zoologist. 

37:369-388. 

Grue, C.E. and B.K. Shipley. 1984. Sensitivity of nestling and 

adult starlings to dicrotophos, an organophosphate insecticide. 

Environmental Research. 35:454-465.

Gryzbowski, J.A. 1991. Ecology and management of isolated populations of the black-capped vireo (Vireo atricapillus) in Oklahoma. 

Oklahoma Department of Conservation Performance Report 

E-1-6. 

Hands, H.M., R.D. Drobney, and M.R. Ryan. 1989. Status of the 

Loggerhead Shrike in the northcentral United States. U.S. Fish 

and Wildlife Service, Missouri Cooperative Fish and Wildlife 

Research Unit, University of Missouri, Columbia, Missouri.

Hart, A.D.M. 1993. Relationship between behavior and the inhibition 

of acetyl-cholinesterase in birds exposed to organophosphorus pesticides. Environmental Toxicology and Chemistry. 12:321-336.

Haug, E.A. 1985. Observations on the breeding ecology of Burrowing 

Owls in Saskatchewan. M.S. thesis. Saskatoon, Saskatchewan, 

Canada, University of Saskatchewan.

Haug, E.A. and A.B. Didiuk. 1991. Updated status report on the 

burrowing owl Athene cunicularia hypugaea in Canada. Committee 

on the Status of Endangered Wildlife in Canada.

Haug, E.A., B.A. Millsap, and M.S.Martell. 1993. Burrowing Owl 

(Speotyto cunicularia). In: A. Poole and F. Gill, eds. The Birds 

of North America, No. 61. The Academy of Natural Sciences, 

Philadelphia, Pennsylvania, and The American Ornithologists’ 

Union, Washington, D.C.

Haug, E.A. and L.W. Oliphant. 1987. Breeding biology of Burrowing 

Owls in Saskatchewan. Pages 269-271. In: G.L. Holroyd, W.B. 

McGillivray, P.H.R. Stepney, D.M. Ealey, G.C. Trottier, and K.E. 

Eberhart, eds. Endangered species in the Prairie Provinces.

Provincial Museum of Alberta Occasional Paper No. 9, Edmonton, 

Alberta.

Haug, E.A. and L.W. Oliphant. 1990. Movements, activity patterns, 

and habitat use of Burrowing Owls in Saskatchewan. Journal of 

Wildlife Management. 54:27-35.

Haukos, D.A. and L.M. Smith. 1989. Lesser Prairie-Chicken nest 

site selection and vegetation characteristics in tebuthiurontreated and untreated sand shinnery oak in Texas. Great Basin 

Naturalist. 49:624-626.

Hawks Aloft. 1998. Nest site selection, reproductive success, and 

territory reoccupation of Ferruginous Hawks in three regions of 

New Mexico. Report submitted to the Bureau of Land Management, 

New Mexico Department of Game and Fish, the National Fish 

and Wildlife Foundation, the U.S.Fish and Wildlife Service, and 

the Turner Foundation. Albuquerque, New Mexico.

Henny, C.J., F.P. Ward, K.E. Riddle, and R.M. Prouty. 1982. Migratory 

Peregrine Falcons, Falco peregrinus, accumulate pesticides 

in Latin America during winter. Canadian Field Naturalist. 

96:333-338.

Herkert, J.R. 1991. Study suggests increases in restored prairie 

fragments to conserve breeding bird communities. Restoration 

and Management Notes. 9:107.

Herkert, J.R. 1994. The effects of habitat fragmentation on midwestern grassland bird communities. Ecological Applications. 

4:461-471.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 131

Herkert, J.R., D.W. Sample, and R.E. Arner. 1996. Management of 

midwestern grassland landscapes for the conservation of migratory birds. Pages 89-116. In: F.R. Thompson III, ed. Managing 

midwestern landscapes for the conservation of neo-tropical 

migratory birds. USDA Forest Service North Central Forest 

Experimental Station, General Technical Report GTR-NC-187, 

St. Paul, Minnesota.

Hill, E.F. and W.J. Fleming. 1982. Anticholinesterase poisoning 

of birds: field monitoring and diagnosis of acute poisoning. 

Environmental Toxicology and Chemistry. 1:27-38.

Hobbs, R.J. and L.F. Huenneke. 1992. Disturbance, diversity, and 

invasion: implications for conservation. Conservation Biology. 

6:324-337.

Hoffman, D.M. 1963. The lesser prairie-chicken in Colorado. Journal 

of Wildlife Management. 27:726-732.

Holland, E.A. and J.K. Detling. 1990. Plant response to herbivory 

and below ground nitrogen cycling. Ecology . 71:1040-1049.

Holochek, J.L., R.D. Piper, and C.H. Herbel. 1989. Range management principles and practices. Prentice Hall, Englewood Cliffs, 

New Jersey.

Hooper, M.J., L.W. Brewer, G.P. Cobb, and R.J. Kendall. 1990. An 

integrated laboratory and field approach for assessing hazards of 

pesticides exposure to wildlife. Pages 271-283 In: L. Somerville 

and C.H. Walker, eds. Pesticide effects on terrestrial wildlife. 

Taylor and Francis, London, England. 

Horton, R.E. 2000. Distribution and abundance of lesser prairiechicken in Oklahoma. The Prairie Naturalist. 32:189-195.

Howard, R.P. and M.L. Wolfe. 1976. Range improvement practices and 

Ferruginous Hawks.Journal of Range Management. 29:33-37.

Howell, S.N.G. and S. Webb. 1995. A guide to the birds of Mexico 

and northern Central America. Oxford University Press, Oxford, 

England.

Hubbard, J.P. 1977. The status of Cassin’s sparrow in New Mexico 

and adjacent states. American Birds. 31:933-941.

Hubbard, J.P. 1978. Revised check-list of the birds of New Mexico. 

New Mexico Ornithological Society Publication No. 6.

Hubbard, J.P. and C.G. Schmitt. 1984. The black-footed ferret in New 

Mexico. New Mexico Game and Fish Department, unpublished 

report to the New Mexico Bureau of Land Management, Santa 

Fe, New Mexico.

Humphrey, R.R. 1958. The desert grassland. Botanical Review. 

24:193-253.

Humphrey, R.R. 1987. 90 years and 535 miles: vegetation changes 

along the Mexican border. University of New Mexico Press, 

Albuquerque, New Mexico.

Hungerford, C.R. 1964. Vitamin A and productivity in Gambel’s 

quail. Journal of Wildlife Management. 28:141-147.

Hunter Jr., M.L. and J.W. Witham. 1985. Effects of a carbaryl induced 

depression of arthropod abundance on the behavior of Parulinae 

warblers. Canadian Journal of Zoology. 63:2612-2616.

Hurley, R.J. and E.C. Franks. 1976. Changes in the breeding ranges 

of two grassland birds. Auk. 93:108-115.

Jackson, A.S. and R. DeArment. 1963. The lesser prairie-chicken 

in the Texas panhandle. Journal of Wildlife Management. 

27:733-737.

James, P.C. and R.H.M. Espie. 1997. Current status of the burrowing owl in North America: an agency survey. Pages 3-5. In: J.L. 

Lincer and K. Steenhof, eds. The burrowing owl: its biology and 

management. Raptor Research Report No. 9. Raptor Research 

Foundation.

James, P.C. and T.J. Ethier. 1989. Trends in the winter distribution 

and abundance of burrowing owls in North America. American 

Birds. 43:1224-1225.

James, P.C. and G.A. Fox. 1987. Effects of some insecticides on 

productivity of Burrowing Owls. Blue jay. 45:65-71.

James, P.C., G.A. Fox, and T.J. Ethier. 1990. Is the operational use of 

strychnine to control ground squirrels detrimental to Burrowing 

Owls? Journal of Raptor Research. 24:120-123.

James, T.R. and R.W. Seabloom. 1968. Notes on the burrow ecology 

and food habits of the Burrowing Owl in southwestern North 

Dakota. Blue Jay. 26:83-84.

Janes, S.W. 1985. Habitat selection in raptorial birds. Pages 159-

188. In: M.L. Cody, ed. Habitat selection in birds. Academic Press, 

New York, New York.

Jensen, J.K. 1925. Late nesting of scaled quail. Auk. 42:129-130.

Jewett, S.G. 1936. Bird notes from Harney County, Oregon, during 

May 1934. Murrelet. 17:41-47.

Johnsgard, P.A. 1979. Birds of the Great Plains. University of 

Nebraska Press, Lincoln, Nebraska.

Johnsgard, P.A. 1980. A preliminary list of the birds of Nebraska 

and adjacent Plains states. University of Nebraska, Lincoln, 

Nebraska.

Johnson, D.H. 1997. Effects of fire on bird populations in mixedgrass prairie. Pages 181-206. In: F.L. Knopf and F.B. Samson, eds. 

Ecology and conservation of Great Plains vertebrates. SpringerVerlag, New York, New York.

Johnson, D.H. and L.D. Igl. 1995. Contributions of the Conservation 

Reserve Program to populations of breeding birds in North Dakota. 

Wilson Bulletin . 107:709-718.

Johnson, D.H. and L.D. Igl. 2001. Effects of management practices 

on grassland birds. Northern Prairie Wildlife Research Center 

Home Page, available at http://www.npwrc.usgs.gov/resource/

literatr/grasbird/grasbird.htm (Version 11APR2001). Northern 

Prairie Wildlife Research Center, Jamestown, North Dakota. 

Johnson, D.H. and M.D. Schwartz. 1993. The Conservation Reserve 

Program: habitat for grassland birds.Great Plains Research. 

3:273-295.

Johnson, R.G. and S.A. Temple. 1986. Assessing habitat quality for 

birds nesting in fragmented tallgrass prairies. Pages 245-249. 

In: J.A. Verner, M.L. Morrison, and C.J. Ralph, eds. Wildlife 

2000: Modeling habitat relationships of terrestrial vertebrates. 

University of Wisconsin Press, Madison, Wisconsin.

Johnson, R.G. and S.A. Temple. 1990. Nest predation and brood parasitism of tallgrass prairie birds.Journal of Wildlife Management. 

54:106-111.

Johnston, D.W. 1975. Organochlorine pesticide residues in small 

migratory birds, 1974-1973. Pesticide Monitoring Journal. 

9:79-88.

Jones, A. 2000. Effects of cattle grazing on North American arid 

ecosystems: a quantitative review. Western North American 

Naturalist. 60:155-164.

Jones, R.E. 1963. Identification and analysis of Lesser and Greater 

Prairie-chicken habitat. Journal of Wildlife Management. 

27:757-778.

Jones, S.L. and M.T. Green. 1998. Baird’s Sparrow status assessment and conservation plan. May, 1998. Administrative Report. 

U.S. Department of Interior, Fish and Wildlife Service, Denver, 

Colorado.

Kantrud, H.A. 1981. Grazing intensity effects on the breeding 

avifauna of North Dakota native grasslands. Canadian Field 

Naturalist. 95:404-417.

Kantrud, H.A. and K.F. Higgins. 1992. Nest and nest site characteristics of some ground-nesting, non-passerine birds of northern 

grasslands. Prairie Naturalist. 24:67-84.

Kantrud, H.A. and R.L. Kologiski. 1982. Effects of soils and grazing on breeding birds of uncultivated upland grasslands of the 

northern Great Plains. USDI Fish and Wildlife Service Research 

Publication no. 15, Washington, D.C.

King, J.W. and J.A. Savidge. 1995. Effects of the Conservation 

Reserve Program on wildlife in southeast Nebraska. Wildlife 

Society Bulletin. 23:377-385.

Knight, R.L. and Gutzwiller, K.G. 1995. Wildlife and recreationists: coexistence through research and management. Covelo, 

California, Island Press. 

Knopf, F.L. 1994. Avian assemblages on altered grasslands. Studies 

in Avian Biology. 15:247-257.

Knopf, F.L. 1996a. Prairie Legacies—Birds. Pages 135-148. In: 

F.B. Samson and F.L. Knopf, eds. Prairie conservation: preserving North America’s most endangered ecosystem. Island Press, 

Washington, D.C.

Knopf, F.L. 1996b. Grazing nongame bird habitats. Pages 51-58 in

P.R. Krausmann, ed. Rangeland wildlife. The Wildlife Society, 

Denver, Colorado.

Knopf, F.L. 1996c. Mountain Plover (Charadrius montanus). in A. 

Poole and F. Gill, eds. The birds of North America, No. 211. The 

Academy of Natural Sciences and The American Orntihologists’ 

Union, Philadelphia, Pennsylvania and Washington, D.C.

Knopf, F.L. and B.J. Miller. 1994. Charadrius montanus—montane, 

grassland, or bare-ground plover? Auk. 11:504-506.
132 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Knopf, F.L. and J.R. Rupert. 1995. Habits and habitats of Mountain 

Plovers in California. Condor. 97:743-751.

Knopf, F.L. and J.R. Rupert. 1996. Reproduction and movements 

of mountain plovers breeding in Colorado. Wilson Bulletin. 

108:28-35.

Knopf, F.L. and J.R. Rupert. 1999. Use of cultivated fields by 

breeding mountain plovers in Colorado. Studies in Avian Biology. 

19:81-86.

Knowles, C. 1986. Some relationships of black-tailed prairie dogs to 

livestock grazing. Great Basin Naturalist. 46:198-203.

Knowles, C.J. and P.R. Knowles. 1984. Additional records of Mountain 

Plovers using prairie dog towns in Montana. Prairie Naturalist. 

16:183-186.

Knowles, C.J., C.J. Stoner, and S.P. Gieb. 1982. Selective use of 

black-tailed prairie dog towns by Mountain Plovers. Condor. 

84:71-74.

Koerth, B.H., W.M. Webb, F.C. Bryant, and F.S. Guthery. 1983. 

Cattle trampling of simulated ground nests under short duration and continuous grazing. Journal of Range Management. 

36:385-386.

Koford, C.B. 1958. Prairie dogs, whitefaces, and blue grama. Wildlife 

Monographs. 3:1-78.

Konrad, P.M. and D.S. Gilmer. 1984. Observations on the nesting 

ecology of Burrowing Owls in central North Dakota. Prairie 

Naturalist. 16:129-130.

Korschgen, L.J. 1970. Soil food-chain pesticide relationships in aldrintreated fields. Journal of Wildlife Management. 34:186-199.

Kotliar, N.B., B.W. Baker, A.D. Whicker, and G. Plumb. 1999. A 

critical review of assumptions about the prairie dog as a keystone 

species. Environmental Management. 24:177-192.

Krueger, K. 1986. Feeding relationships among bison, pronghorn, and 

prairie dogs: an experimental analysis. Ecology. 67:760-770.

Lane, J. 1968. Baird’s Sparrow. United States National Museum 

Bulletin. 237:745-765.

Lanyon, W.E. 1995. Eastern Meadowlark (Sturnella magna). in

A. Poole and F. Gill, eds. The birds of North America, No. 160. 

Academy of Natural Sciences, Philadelphia, PA, and American 

Ornithologists’ Union, Washington, D.C.

Lasley, G.W. and C. Sexton. 1993. The nesting season. American 

Birds. 47:1128.

Latta, M.J., C.J. Beardmore, and T.E. Corman. 1999. Arizona 

Partners in Flight Bird Conservation Plan version 1.0. Nongame 

and endangered wildlife program technical report. Arizona Game 

and Fish Department, Phoenix, Arizona.

Laursen, K. 1981. Birds on roadside verges and the effect of mowing on 

frequency and distribution. Biological Conservation. 20:59-68.

Leary, A.W., R. Mazaika, and M.J. Bechard. 1998. Factors affecting 

the size of Ferruginous Hawk home ranges. Wilson Bulletin. 

110:198-205.

Lehmann, V.W. 1953. Bobwhite population fluctuations and vitamin 

A. Transactions of the North American Wildlife Conference. 

18:199-246.

Ligon, J.S. 1927. Wildlife of New Mexico, its conservation and 

management. New Mexico State Game Commission, Santa Fe, 

New Mexico.

Ligon, J.S. 1937. Tragedy of upland game birds throughout the 

west and southwest. Transactions of the North American Wildlife 

Conference. 2:476-480.

Ligon, J.S. 1961. New Mexico birds and where to find them. University 

of New Mexico Press, Albuquerque, New Mexico.

Loftin, S.R., C.E. Bock, J.H. Bock, and S.L. Brantley. 2000. Desert 

grasslands. Pages 53-96 In: R. Jemison and C. Raish, eds. Livestock 

management in the American Southwest: ecology, society, and 

economics. Elsevier, Amsterdam, The Netherlands.

Lokemoen, J.T. and H.F. Duebbert. 1976. Ferruginous Hawk nesting ecology and raptor populations in northern South Dakota. 

Condor. 78:464-470.

Lovejoy, T.C., R.O. Bieregaard Jr., A.B. Rylands, J.R. Malcolm, 

C.E. Quintela, L.H. Harper, K.S. Brown Jr., A.H. Powell, G.V.N. 

Powell, H.O.R. Schubart, and M.B. Hays. 1986. Edge and other 

effects of isolation on Amazon forest fragments. Pages 257-285. 

In: M.E. Soulé, ed. Conservation biology: The science of scarcity 

and diversity. Sinauer Associates, Sunderland, Massachusetts.

Lowther, P.E., K.D. Groschupf, and S.M. Russell. 1999. Rufouswinged Sparrow (Aimophila carpalis). In: A. Poole and F. Gill, eds. 

The birds of North America, No. 422. The Academy of Natural 

Sciences, Philadelphia, and The American Ornithologists’ Union, 

Washington, D.C.

MacCracken, J.G., D.W. Uresk, and R.M. Hansen. 1984. Burrowing 

Owl nesting habitat use in Conata Basin, South Dakota.Abstracts 

of the Society for Range Mangaement 37th annual Meeting, Rapid 

City, South Dakota.

MacCracken, J.G., D.W. Uresk, and R.M. Hansen. 1985. Vegetation 

and soils of Burrowing Owl nest sites in Conata Basin, South 

Dakota. Condor. 87:152-154.

Mack, R.N. 1981. Invasion of Bromus tectorum L. into western North 

America: an ecological chronicle. Agro-Ecosystems. 7:145-165.

Mack, R.N. and J.N. Thompson. 1982. Evolution in steppe with few 

large, hooved mammals.American Naturalist. 119:757-773.

Madden, E.M., A.J. Hansen, and R.K. Murphy. 1999. Influence of 

prescribed fire history on habitat and abundance of passerine 

birds in northern mixed-grass prairie. Canadian Field Naturalist. 

113:627-640.

Maher, W.J. 1973. Matador Project: Birds I. Population dynamics. 

Canadian Committee for the International Biological Programme, 

Matador Project, Technical Report 34. University of Saskatchewan, 

Saskatoon, Saskatchewan.

Maher, W.J. 1979. Nesting diets of prairie passerine birds at Matador, 

Saskatchewan, Canada. Ibis. 121:437-452.

Manzano-Fischer, P., R. List, and G. Ceballos. 1999. Grassland 

birds in prairie-dog towns in northwestern Chihuahua, Mexico. 

Studies in Avian Biology. 19:263-271.

Marlette, G.M. and J.E. Anderson. 1986. Seed banks and propagule dispersal in crested wheatgrass stands. Journal of Applied 

Ecology. 23:161-175.

Martin, D.J. 1973. Selected aspects of Burrowing Owl ecology and 

behaviour in central New Mexico. Condor. 75:446-456.

Martin, S.J. 1983. Burrowing Owl occurrence on white-tailed prairie 

dog colonies. Journal of Field Ornithology. 54:422-423.

Mastrota, F.N. 1999. Wildlife mortality incidents caused by pesticides: 

an analysis of the EIIS database. U.S. Environmental Protection 

Agency, Washington, D.C.

Mayfield, H.F. 1965. The brown-headed cowbird with old and new 

hosts. Living Bird. 4:13-28.

McCallum, D.A., W.D. Graul, and R. Zaccagnini. 1977. The breeding status of the long-billed curlew in Colorado. The Auk. 94:

599-601.

McCarthey, T.D. and T.E. Corman. 1996. Arizona breeding 

bird atlas project 1996 project report. Nongame Endangered 

Wildlife Program Technical Report 108, Arizona Game and Fish 

Department, Phoenix, Arizona.

McEwen, L.C., C.E. Knittle, and M.L. Richmond. 1972. Wildlife 

effects from grasshopper insecticides sprayed on short-grass 

range. Journal of Range Management. 25:188-194.

McLeod, J.M. 1967. The effect of phosphamidon on bird populations in jack pine stands in Quebec. Canadian Field Naturalist. 

81:102-106.

McPherson, G.R. 1995. The role of fire in the desert grasslands. Pages 

130-151 in M.P. McClaran and T.R. Van Devender, eds. The desert 

grassland. University of Arizona Press, Tucson, Arizona.

Mearns, E.A. 1907. Mammals of the Mexican boundary of the 

United States, Part I. United States National Museum Bulletin 

No. 56.

Medina, A.L. 1988. Diets of scaled quail in southern Arizona. Journal 

of Wildlife Management. 32:753-757.

Mellink, E. and H. Madrigal. 1993. Ecology of Mexican prairie dogs. 

Journal of Mammalogy. 74:631-635.

Merriam, C.H. 1902. The prairie dog of the Great Plains. Pages 

257-270 in Yearbook of the U.S. Department of Agriculture 1901. 

U.S. Government Printing Office, Washington, D.C.

Milchunas, D.G., O.E. Sala, and W.K. Lauenroth. 1988. A generalized 

model of the effects of grazing by large herbivores on grassland 

community structure. American Naturalist. 132:87-106.

Miller, A.H. 1951. A comparison of the avifaunas of Santa Cruz and 

Santa Rosa islands, California. Condor. 53:117-123.

Miller, B., G. Ceballos, and R. Reading. 1994. The prairie dog and 

biotic diversity. Conservation Biology. 8:677-681.

Millikin, R.L. and J.N.M. Smith. 1990. Sublethal effects of 

fenitrothion on forest passerines. Journal of Applied Ecology. 

27:983-1000.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 133

Møller, A.P. 1989. Nest site selection across field-woodland ecotones: 

The effect of nest predation. Oikos. 56:240-246.

Monson, G. and A.R. Phillips. 1981. Annotated checklist of the birds 

of Arizona. University of Arizona Press, Tucson, Arizona.

Moore, R.T. 1946. The Rufous-winged Sparrow, its legends and 

taxonomic status. Condor. 48:117-123.

Mora, M.A., M.C. Lee, J.P. Jenny, T.W. Schultz, J.L. Sericano, and 

N.J. Clum. 1997. Potential effects of environmental contaminants 

on recovery of the aplomado falcon in south Texas. Journal of 

Wildlife Management. 61:1288-1296.

Mote, K.D., R.D. Applegate, J.A. Bailey, K.M. Giesen, R. Horton, and 

J.L. Sheppard. 1999. Assessment and conservation strategy for 

the lesser prairie-chicken (Tympanuchus pallidicinctus). Kansas 

Department of Wildlife and Parks, Emporia, Kansas.

Moulding, J.D. 1976. Effects of a low-persistence insecticide on 

forest bird populations. Auk. 93:692-708.

Mulhern, D.W. and C.J. Knowles. 1996. Black-tailed prairie dog status 

and future conservation planning. Pages 19-29. In: D.W. Uresk, 

G.L. Schenbeck, and J.T. O’Rourke, eds. Conserving biodiversity 

on native rangelands: symposium proceedings; August 17, 1995; 

Fort Robinson State Park, Nebraska. USDA Forest Service, Rocky 

Mountain Forest and Range Experiment Station, RM-GTR-298, 

Fort Collins, Colorado.

National Geographic Society. 1999. Field Guide to the Birds of 

North America (Third Edition). National Geographic Society, 

Washington, D.C.

NatureServe: An online encyclopedia of life [web application]. 2001. 

Version 1.5. Available: http://www.natureserve.org/. (Accessed 

November 23, 2001). Association for Biodiversity Information, 

Arlington, Virginia.

Neilson, R.P. 1986. High-resolution climatic analysis and Southwest 

biogeography. Science. 232:27-34.

New Mexico Department of Game and Fish. 1988. Handbook of species endangered in New Mexico, F-150. New Mexico Department 

of Game and Fish, Santa Fe, New Mexico.

New Mexico Department of Game and Fish. 2002. Threatened and 

endangered species of New Mexico: biennial review and recommendations. New Mexico Department of Game and Fish, Santa 

Fe, New Mexico.

New Mexico Partners in Flight. 2001. Draft Land Bird Conservation 

Plan for the State of New Mexico, Version 1.0. Available at http://

www.hawksaloft.org/pif.html. Compiled by Christopher Rustay, 

Albuquerque, New Mexico.

Norman, R.F. and R.J. Robertson. 1975. Nest-searching behavior 

in the brown-headed cowbird. Auk. 92:610-611.

Noss, R.F. and A.Y. Cooperrider. 1994. Saving nature’s legacy: 

Protecting and restoring biodiversity. Island Press, Covelo, 

California.

Oberholser, H.C. 1974. The bird life of Texas. University of Texas 

Press, Austin, Texas.

Olawsky, C.D. and L.M. Smith. 1991. Lesser Prairie-chicken densities on tebuthiuron-treated and untreated san shinnery oak 

rangelands. Journal of Range Management. 44:364-368.

O’Leary, C.H. and D.W. Nyberg. 2000. Treelines between fields 

reduce the density of grassland birds. Natural Areas Journal. 

20:243-249.

Olendorff, R.R. 1993. Status, biology and management of Ferruginous 

Hawks: a review. Raptor Research and Technical Assistance 

Center, Special Report. USDI Bureau of Land Management, 

Boise, Idaho.

Olson, S.L. and D. Edge. 1985. Nest site selection by Mountain 

Plovers in northcentral Montana. Journal of Range Management. 

38:280-282.

Olson-Edge, S.L. and D. Edge. 1987. Density and distribution of 

the Mountain Plover on the Charles M. Russell National Wildlife 

Refuge. Prairie Naturalist. 19:233-238.

O’Meilia, M.E., F.L. Knopf, and J.C. Lewis. 1982. Some consequences 

of competition between prairie dogs and beef cattle. Journal of 

Range Management. 35:580-585.

Owens, R.A. and M.T. Myres. 1973. Effects of agriculture upon populations of native passerine birds of an Alberta fescue grassland. 

Canadian Journal of Zoology. 51:697-713.

Paine, L.K., D.J. Undersander, D.W. Sample, G.A. Bartelt, and T.A. 

Schatteman. 1996. Cattle trampling of simulated ground nests 

in rotationally grazed pastures. Journal of Range Management. 

49:294-300.

Palmer, R.S. 1988. Ferruginous Hawk. Pages 135-151. In: Handbook 

of North American birds. Volume 5, Part 2. Yale University Press, 

New Haven, Connecticut.

Pampush, G.J. and R.G. Anthony. 1993. Nest success, habitat 

utilization and nest-site selection of long-billed curlews in the 

Columbia Basin, Oregon. The Condor. 95:957-967.

Parmenter, R.R. and T.R. Van Devender. 1995. Diversity, spatial variability, and functional roles of vertebrates in the desert grassland. 

Pages 196-229 in M.P. McClaran and T.R. Van Devender, eds. 

The desert grassland. University of Arizona Press, Tucson, 

Arizona.

Parrish, T.L., S.H. Anderson, and W.F. Oelklaus. 1993. Mountain 

Plover habitat selection in the Powder River Basin, Wyoming. 

Prairie Naturalist. 25:219-226.

Patnode, K.A. and D.H. White. 1991. Effects of pesticides on songbird 

productivity in conjunction with pecan cultivation in southern 

Georgia: a multiple-exposure experimental design. Environmental 

Toxicology and Chemistry. 10:1479-1486.

Paton, P.W.C. and J. Dalton. 1994. Breeding ecology of long-billed 

curlews at Great Salt Lake, Utah. Great Basin Naturalist. 

54:79-85.

Patterson, M.P. and L.B. Best. 1996. Bird abundance and nesting 

success in Iowa CRP fields: the importance of vegetation structure 

and composition. American Midland Naturalist. 135:153-167.

Payne, N.F. and F.C. Bryant. 1994. Techniques for wildlife habitat 

management of uplands. McGraw-Hill, New York, New York.

Peterjohn, B.G. and J.R. Sauer. 1999. Population status of North 

American grassland birds from the North American Breeding Bird 

Survey 1966-1996. Studies in Avian Biology. 19:27-44.

Phillips, A.R. 1968. Aimophila carpalis carpalis (Coues) Rufouswinged Sparrow. U.S. National Museum Bulletin. 237:902-919.

Phillips, A.R., J. Marshall, and G. Monson. 1964. The birds of Arizona. 

University of Arizona Press, Tucson, Arizona.

Pimentel, D. 2001. Environmental effects of pesticides on public 

health, birds, and other organisms. Talk given at conference 

proceedings: Rachel Carson and the Conservation Movement: Past, 

Present, Future. Shepherdstown, West Virginia, August 2001. 

Pimentel, D., H. Acquay, M. Biltonen, P. Rice, M. Silva, J. Nelson, 

V. Lipner, S. Giordano, A. Horowitz, and M. D’Amore. 1992 

.Environmental and economic costs of pesticide use. BioScience. 

42:750-760.

Pimentel, D., L. McLaughlin, A. Zepp, B. Lakitan, T. Kraus, P. 

Kleinman, F. Vancini, W.J. Roach, E. Graap, W.S. Keeton, and 

G. Selig. 1991.Environmental and economic effects of reducing 

pesticide use. BioScience. 41:402-409.

Plumpton, D.L. and D.E. Andersen. 1998. Anthropogenic effects 

on winter behavior of ferruginous hawks. Journal of Wildlife 

Management. 62:340-346.

Plumpton, D.L. and R.S. Lutz. 1993. Nesting habitat use by 

Burrowing Owls in Colorado. Journal of Raptor Research. 

27:175-179.

Porter, D.K. and R.A. Ryder. 1974. Avian density and productivity 

studies and analyses on the Pawnee Site in 1972. U.S. International 

Biological Program, Grassland Biome Technical Report 252. 

Colorado State University, Fort Collins, Colorado.

Porter, D.K., M.A. Strong, J.B.Giezentanner, and R.A. Ryder. 1975. 

Nest ecology, productivity, and growth of the Loggerhead Shrike on 

the shortgrass prairie. Southwestern Naturalist. 19:429-436. 

Power, M.E., D. Tilman, J.A. Estes, B.A. Menge, W.J. Bond, L.S. 

Mills, G. Daily, J.C. Castilla, J. Lubchenco, and R.T. Paine. 1996. 

Challenges in the quest for keystones. BioScience. 466:9-20.

Prescott, D.R.C. and A.J. Murphy. 1996. Habitat associations 

of grassland birds on native and tame pastures of the aspen 

parkland in Alberta. NAWMP-021. Alberta NAWMP Centre, 

Edmonton, Alberta.

Prescott, D.R.C. and A.J. Murphy. 1999. Bird populations in seeded 

nesting cover on North American Waterfowl Management Plan 

properties in the aspen parkland of Alberta. Studies in Avian 

Biology. 19:203-210.

Pulliam, H.R. and G.S. Mills. 1977. The use of space by wintering 

sparrows. Ecology. 58:1393-1399.

Pylypec, B. 1991. Impacts of fire on bird populations in a fescue 

prairie. Canadian Field Naturalist. 105:346-349.
134 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Ratti, J.T. and K.P. Reese. 1988. Preliminary test of the ecological 

trap hypothesis. Journal of Wildlife Management. 52:484-491.

Ratti, J.T. and J.M. Scott. 1991. Agricultural impacts on wildlife: 

problem review and research needs. Environmental Professional. 

13:263-274.

Redmond, R.L. 1986. Egg size and laying date of long-billed curlews 

(Numenius americanus): implications for female reproductive 

tactics. Oikos. 46:330-338.

Redmond, R.L., T.K. Bicak, and D.A. Jenni. 1981. An evaluation 

of breeding season census techniques for long-billed curlews 

(Numenius americanus).Studies in Avian Biology. 6:197-201.

Redmond, R.L. and D.A. Jenni. 1982. Natal philopatry and breeding area fidelity of long-billed curlews (Numenius americanus): 

patterns and evolutionary consequences. Behavioral Ecology and 

Sociobiology. 10:277-279.

Redmond, R.L. and D.A. Jenni. 1986. Population ecology of the 

long-billed curlew (Numenius americanus) in western Idaho. 

The Auk. 103:755-767.

Renaud, W.E. 1980. The long-billed curlew in Saskatchewan: status 

and distribution. Blue Jay. 38:221-237.

Renken, R.B. and J.J. Dinsmore. 1987. Nongame bird communities on 

managed grasslands in North Dakota. Canadian Field-Naturalist. 

101:551-557.

Rich, T. 1986. Habitat and nest-site selection by Burrowing Owls in 

the sagebrush steppe of Idaho. Journal of Wildlife Management. 

50:548-555.

Riley, T.Z. and C.A. Davis. 1993. Vegetative characteristics of 

lesser prairie-chicken brood foraging sites. Prairie Naturalist. 

25:243-248.

Riley, T.Z., C.A. Davis, M. Ortiz, and M.J. Wisdom. 1992. Vegetative 

characteristics of sucessful and unsuccessful nests of Lesser 

Prairie-chickens. Journal of Wildlife Management. 56:383-387.

Riley, T.Z., C.A. Davis, and R.A. Smith. 1993a. Autumn and winter 

foods of the lesser prairie-chicken (Tympanuchus pallidicinctus) 

(Galliformes: Tetraonidae). Great Basin Naturalist. 53:186-

189.

Riley, T.Z., C.A. Davis, and R.A. Smith. 1993b. Autumn-winter habitat 

use of Lesser Prairie-Chickens (Tympanuchus pallidicinctus, 

Tetraonidae). Great Basin Naturalist. 53:409-411.

Rising, J.D. 1996. A guide to the identification and natural history of 

the sparrows of the United States and Canada. Academic Press, 

New York, New York.

Robbins, C.S., J.R. Sauer, R.S. Greenberg, and S. Droege. 1989. 

Population declines in North American birds that migrate to 

the neotropics. Proceedings of the National Academy of Sciences 

(USA). 86:7658-7662.

Robbins, M.B. and B.C. Dale. 1999. Sprague’s Pipit Anthus spragueii. 

In: A. Poole and F. Gill, eds. The Birds of North America, No. 439. 

The Academy of Natural Sciences, Philadelphia, and American 

Ornithologists’ Union, Washington, D.C.

Robinson, S.K., S.I. Rothstein, M.C. Brittingham, L.J.Petit, and J.A. 

Grzybowski. 1995a. Ecology and behavior of cowbirds and their 

impact on host populations. Pages 428-460. In: T.E. Martin and 

D.M. Finch, eds. Ecology and management of neotropical migratory 

birds. Oxford University Press, New York, New York.

Robinson, S.K., F.R. Thompson III, T.M. Donovan, D.R. Whitehead, 

and J. Faaborg. 1995b. Regional forest fragmentation and the 

nesting success of migratory birds. Science. 267:1987-1990.

Rodenhouse, N.L., L.B. Best, R.J. O’Connor, and E.K. Bollinger. 

1993. Effects of temperate agriculture on neotropic migrant 

landbirds. Pages 280-295. In: D.M. Finch and P.W. Stangel, eds. 

Status and management of neotropical migratory birds. USDA 

Forest Service Rocky Mountain Research Station GTR-RM-229, 

Fort Collins, Colorado.

Rodriguez Estrella, R., F. Chavez Ramirez, and G.L. Holroyd. 1998. 

Current knowledge of the burrowing owl in Mexico: what is needed 

for a conservation plan? Abstract and notes. Second International 

Burrowing Owl Symposium, September 29-30, 1998, Ogden, Utah. 

Canadian Wildlife Service, Environment Canada.

Rotenberry, J.T. and J.A. Wiens. 1980. Temporal variation in habitat 

structure and shrubsteppe bird dynamics. Oecologia. 47:1-9.

Rothstein, S.I. 1975. An experimental and teleonomic investigation 

of avian brood parasitism. Condor. 77:250-271.

Rothstein, S.I. 1977. Cowbird parasitism and egg recognition of the 

Northern Oriole. Wilson Bulletin. 89:21-32.

Rothstein, S.I. 1982. Successes and failures in avian egg and nestling 

recognition with comments on the utility of optimality reasoning. 

American Zoologist. 22:547-560.

Russell, S.M. and G. Monson. 1998. The birds of Sonora. University 

of Arizona Press, Tucson, Arizona.

Ruth, J.M. 2000. Cassin’s Sparrow (Aimophila cassinii) status assessment and conservation plan. Biological Technical Publication 

BTP-R6002-1999. U.S. Fish and Wildlife Service, Denver, 

Colorado.

Ryan, M.R. 1986. Nongame management in grassland and agricultural ecosystems. Pages 117-136. In: J.B. Hale, L.B. Best, 

and R.L. Clawson, eds. Management of nongame wildlife in the 

Midwest. Northcentral Section of the Wildlife Society, Grand 

Rapids, Michigan.

Ryder, R.A. 1980. Effects of grazing on bird habitats. In: R.M. 

DeGraaf and N.G. Tilghman, eds. Management of western 

forests and grasslands for nongame birds. USDA Forest Service, 

Intermountain Forest and Range Experiment Station General 

Technical Report INT-86, Ogden, Utah.

Saab, V.A., C.E. Bock, T.D. Rich, and D.S. Dobkin. 1995. Livestock 

grazing effects in western North America. Pages 311-353. In: 

T.E. Martin and D.M. Finch, eds. Ecology and management of 

neotropical migratory birds: a synthesis and review of critical 

issues. Oxford University Press, New York, New York.

Saiwana, L., J.L. Holechek, A. Tembo, R. Valdez, and M. Cardenas. 

1998. Scaled quail use of different seral stages in the Chihuahuan 

desert. Journal of Wildlife Management. 62:550-556.

Salt, W.R. and A.L. Wilk. 1958.The birds of Alberta. Department of 

Economic Affairs, Edmonton, Alberta.

Samson, F.B. 1980. Island biogeography and the conservation of 

prairie birds. Pages 293-305. In: C.L. Kucera, ed. Proceedings 

of the Seventh North American Prairie Conference. Southwest 

Missouri State University, Springfield, Missouri.

Samson, F. and F. Knopf. 1994. Prairie conservation in North 

America. BioScience. 44:418-421.

Sandoz, M. 1954. The buffalo hunters. Hasting House, New York, 

New York.

Sands, J.L. 1968. Status of the lesser prairie-chicken. Audubon 

Field Notes. 22:454-456.

Sauer, J.R. and S. Droege. 1992. Geographic patterns in population 

trends of neotropical migrants in North America. Pages 26-42. In: 

Hagan III, J.M. and D.W. Johnston, eds. Ecology and conservation 

of neotropical migrant landbirds. Smithsonian Institution Press, 

Washington, D.C.

Sauer, J.R., J.E. Hines, and J. Fallon. 2001. The North American 

Breeding Bird Survey, Results and Analysis 1966 - 2000. Version 

2001.2. Available at http://www.mbr.nbs.gov/bbs/bbs.html, USGS 

Patuxent Wildlife Research Center, Laurel, Maryland.

Sauer, J.R., J.E. Hines, G. Grough, I. Thomas, and B.G. Peterjohn. 

1997. The North American Breeding Bird Survey Results and 

Analysis. Version 96.4. Patuxent Wildlife Research Center, 

Laurel, Maryland.

Sauer, J.R.; B.G. Peterjohn, S. Schwartz, and J.E. Hines. 1995. 

The grassland bird home page. Version 95.0 http://www.mbr.

nbs.gov/bbs/grass/grass.htm. Patuxent Wildlife Research Center, 

Laurel, Maryland.

Sauer, J.R., S. Schwartz, and B. Hoover. 1996. The Christmas Bird 

Count Home Page. Version 95.1. Available at http://www.mbrpwrc.usgs.gov/bbs/cbc.html. Patuxent Wildlife Research Center, 

Laurel, Maryland. 

Saunders, A.A. 1914. The birds of Teton and northern Lewis and 

Clark counties, Montana. Condor. 16:124-144.

Sawyer, P.E. 1973. Habitat-use by scaled quail and other birds 

in southeastern New Mexico. M.S. thesis. New Mexico State 

University, Las Cruces, New Mexico. 

Schemnitz, S.D. 1961. Ecology of the scaled quail in the Oklahoma 

panhandle.Wildlife Monographs. 8.

Schemnitz, S.D. 1994. Scaled quail Callipepla squamata. In: A. 

Poole and F. Gill, eds. The birds of North America, No. 106. The 

Academy of Natural Sciences, Philadelphia, Pennsylvania, and 

the American Ornithologists’ Union, Washington, D.C.

Schlesinger, W.H., J.F. Reynolds, G.L. Cunningham, L.F. Huenneke, 

W.M. Jarrell, R.A. Virginia, and W.G. Whitford. 1990. Biological 

feedbacks in global desertification. Science. 247:1043-1048.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 135

Schmutz, J.K. and R.W. Fyfe. 1987. Migration and mortality of 

Alberta Ferruginous Hawks. Condor. 89:169-174.

Schmutz, J.K., R.W. Fyfe, D.A. Moore, and A.R. Smith. 1984. Artificial 

nests for Ferruginous and Swainson’s hawks. Journal of Wildlife 

Management. 48:1009-1013.

Schmutz, J.K. and D.J. Hungle. 1989. Populations of Ferruginous 

and Swainson’s hawks increase in synchrony with ground squirrels. Canadian Journal of Zoology. 67:2596-2601.

Scott, D.M. 1977. Cowbird parasitism on the gray catbird at London, 

Ontario. Auk. 94:18.

Sealy, S.G. 1992. Removal of yellow warbler eggs in association with 

cowbird parasitism. Condor. 94:40-54.

Shackford, J.S. 1991. Breeding ecology of the Mountain Plover in 

Oklahoma. Bulletin of the Oklahoma Ornithological Society. 

24:9-13.

Shackford, J.S. 1994. Nesting of long-billed curlews on cultivated 

fields. Bulletin of the Oklahoma Ornithological Society. 27:17-

20.

Shackford, J.S., D.M. Leslie Jr., and W.D. Harden. 1999. Range-wide 

use of cultivated fields by Mountain Plovers during the breeding 

season. Journal of Field Ornithology. 70:114-120.

Sharps, J.C. and D.W. Uresk. 1990. Ecological review of black-tailed 

prairie dogs and associated species in western South Dakota. 

Great Basin Naturalist. 50:339-345.

Sheffield, S.R. 1997. Current status, distribution, and conservation of the burrowing owl (Speotyto cunicularia) in Midwestern 

North America. Pages 399-407. In: J.R. Dundan, D.H. Johnson, 

and T.H. Nicholls, eds. Biology and conservation of owls of the 

Northern Hemisphere. USDA Forest Service, North Central 

Forest Experiment Station, General Technical Report NC-190, 

St. Paul, Minnesota.

Silloway, P.M. 1900. Notes on the long-billed curlew. Condor. 

2:79-82.

Sissons, R., K. Scalise, and T.I. Wellicome. 1998. Nocturnal foraging 

habitat use of the Burrowing Owl in a heavily cultivated region 

of southern Saskatchewan. Pages 22-23. Abstracts of the Second 

International Burrowing Owl Symposium. Ogden, Utah.

Skinner, R.M. 1975. Grassland use patterns and prairie bird populations in Missouri. Pages 171-180. In: M.K. Wali, ed. Prairie: a 

multiple view. University of North Dakota Press, Grand Forks, 

North Dakota.

Skinner, R.M., T.S. Baskett, and M.D. Blendon. 1984. Bird habitat 

on Missouri prairies. Terrestrial Series 14. Missouri Department 

of Conservation, Jefferson City, Missouri.

Smith, G., J.L. Holechek, and M. Cardenas. 1996. Wildlife numbers 

on excellent and good condition Chihuahuan Desert rangelands: 

an observation. Journal of Range Management. 49:489-493.

Smith, H. and J. Smith. 1966. A breeding bird survey on uncultivated 

grasslands at Regina. Blue Jay. 24:129-131.

Sousa, P.J. and W.N. McDonal. 1983. Habitat suitability index 

models: Baird’s Sparrow. U.S. Fish and Wildlife Service, FWS/

OBS-82/10.44, Washington, D.C.

Stewart, R.E. 1975. Breeding birds of North Dakota. Tri-college 

Center for Environmental Studies, Fargo, North Dakota.

Stone, W.B. 1979. Poisoning of wild birds by organophosphate and 

carbamate pesticides. New York Fish and Game. 26:37-47.

Stormer, F.A. 1984. Night-roosting habitat of scaled quail.Journal 

of Wildlife Management. 48:191-197.

Stotz, D.F., J.W. Fitzpatrick, T.A. Parker III, and D.K. Moskovits. 

1996. Neotropical birds: ecology and conservation. University of 

Chicago Press, Chicago, Illinois.

Sugden, J.W. 1933. Range restriction of the long-billed curlew. 

Condor. 35:3-9.

Sutter, G.C. 1997. Nest-site selection and nest-entrance orientation 

in Sprague’s pipit. Wilson Bulletin. 109:462-269.

Sutter, G.C. and R.M. Brigham. 1998. Avifaunal and habitat changes 

resulting from conversion of native prairie to crested wheatgrass: 

patterns at songbird community and species levels. Canadian 

Journal of Zoology. 76:869-875.

Sutton, G.M. 1927. Flocking, mating and nest-building habits of the 

Prairie Horned Lark. Wilson Bulletin. 34:131-141.

Swarth, H.S. 1929. The faunal areas of southern Arizona: a study 

in animal distribution. Proceedings of the California Academy 

of Sciences. 18:267-383.

Szmedra, P.I. 1991. Pesticides use in agriculture. Pages 649-677. In: 

D. Pimentel, ed. CRC handbook of pest management in agriculture. 

CRC Press, Boca Raton, Florida.

Taylor, M.A. 1980. Lesser prairie-chicken use of man-made leks. 

Southwestern Naturalist. 24:706-707.

Taylor, M.A. and F.S. Guthery. 1980. Status, ecology and management of the lesser prairie-chicken. USDA Forest Service, Rocky 

Mountain Forest and Range Experiment Station, General 

Technical Report RM-77, Fort Collins, Colorado.

Tellman, B., D.M. Finch, C. Edminster, and R. Hamre. 1998. The 

future of arid grasslands: identifying issues, seeking solutions. 

1996 Oct. 9-13; Tucson, AZ. Fort Collins, Colorado, USDA 

Forest Service, Rocky Mountain Research Station, Proceedings 

RMRS-P-3. 

Temple, S.A., B.M. Fevold, L.K. Paine, D.J. Undersander, and D.W. 

Sample. 1999. Nesting birds and grazing cattle: accommodating 

both on midwestern pastures. Studies in Avian Biology. 19:196-

202.

Terborgh, J. 1989. Where have all the birds gone? Princeton 

University Press, Princeton, New Jersey.

Terres, J.K. 1991. The Audubon Society Encyclopedia of North 

American Birds. Alfred A. Knopf, Inc., New York, New York.

Thompson, E.E. 1893. Additions to the list of Manitoban birds. 

Auk. 10:49-50.

Thomsen, L. 1971. Behavior and ecology of Burrowing Owls on the 

Oakland municipal airport. Condor. 73:177-192.

Timken, R.L. 1969. Notes on the long-billed curlew. Auk. 86:750-

751.

Tisdale, E.W. 1961. Ecological changes in the Palouse. Northwest 

Sciences. 35:134-138.

Trulio, L.A. 1997. Strategies for protecting western burrowing owls 

(Speotyto cunicularia hypugaea). Pages 461-465. In: J.R. Dundan, 

D.H. Johnson, and T.H. Nicholls, eds. Biology and conservation of 

owls in the Northern Hemisphere. USDA Forest Service, North 

Central Forest Experiment Station, General Technical Report 

NC-190, St. Paul, Minnesota.

Ure, J., P. Briggs, and S.W. Hoffman. 1991. Petition to list as 

endangered the Ferruginous Hawk (Buteo regalis), as provided 

by the Endangered Species Act of 1973, as amended in 1982. 

Ferruginous Hawk Project, Salt Lake City, Utah.

U.S. Fish and Wildlife Service. 1991. Endangered and threatened 

wildlife and plants: animal candidate review for listing as endangered or threatened species. Proposed Rule.Federal Register. 

56:58804-58836.

U.S. Fish and Wildlife Service. 1992. Endangered and threatened 

wildlife and plants: notice of finding on petition to list the 

Ferruginous Hawk. Federal Register. 57:37507-37513.

U.S. Fish and Wildlife Service. 1995. Migratory nongame birds of 

management concern in the United States: the 1995 list. U.S. 

Fish and Wildlife Service, Washington, D.C.

U.S. Fish and Wildlife Service. 1998. Endangered and threatened 

wildlife and plants; 12-month finding for a petition to list the 

Lesser Prairie-Chicken as threatened and designate critical 

habitat. Federal Register. 63:31400-31406.

U.S. Fish and Wildlife Service. 1999a. Endangered and threatened 

wildlife and plants; proposed threatened status for the Mountain 

Plover. Federal Register. 64:7587-7601.

U.S. Fish and Wildlife Service. 1999b. Endangered and threatened 

wildlife and plants; 90-Day finding for a petition to list the Baird’s 

Sparrow as threatened with critical habitat. Federal Register. 

64:27747-27749.

U.S. Fish and Wildlife Service. 2000. Endangered and threatened 

wildlife and plants; 12-month finding for a petition to list the 

black-tailed prairie dog as threatened. Federal Register. 65:5476-

5488.

U.S. Fish and Wildlife Service. 2001. Endangered and threatened 

wildlife and plants; review of plant and animal species that are 

candidates or proposed for listing as endangered or threatened, 

annual notice of findings on recycled petitions, and annual 

description of progress on listing actions; proposed rule. Federal 

Register. 66:54807-54832.

U.S. Fish and Wildlife Service. 2002. Species information: threatened 

and endangered animals and plants. Available at http://ecos.fws.

gov/wildlife.html. Accessed August 2002.
136 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Van Pelt, W.E. 1999. The black-tailed prairie dog conservation assessment and strategy - fifth draft. Nongame and Endangered 

Wildlife Program, Arizona Game and Fish Department, Phoenix, 

Arizona.

Verbeek, N.A.M. 1967. Breeding biology and ecology of the Horned 

Lark in alpine tundra. Wilson Bulletin. 79:208-218.

Vickery, P.D., J.R. Herkert, F.L. Knopf, J. Ruth, and C.E. Keller. 

1999a. Grasslands birds: an overview of threats and recommended 

management strategies. In: R.E. Bonney Jr., D.N. Pashley, R. 

Cooper, and L. Niles, eds. Strategies for bird conservation: the 

Partners in Flight planning process. Cornell Laboratory of 

Ornithology, Ithaca, New York. Available at http://birds.cornell.

edu/pifcapemay.

Vickery, P.D., M.L. Hunter Jr., and S.M. Melvin. 1994. Effects of 

habitat area on the distribution of grassland birds in Maine.

Conservation Biology. 8:1087-1097.

Vickery, P.D., P.L. Tubaro, J.M.C. Da Silva, B.G. Peterjohn, J.G. 

Herkert, and R.B. Cavalcanti. 1999b. Conservation of grassland 

birds of the Western Hemisphere. Studies in Avian Biology. 

19:2-26.

Vyas, N.B. 1999. Factors influencing estimation of pesticide-related 

mortality. Toxicology and Industrial Health. 15:186-191.

Wakely, J.S. 1978. Activity budgets, energy expenditures, and energy 

intakes of nesting Ferruginous Hawks. Auk. 95:667-676.

Wallis, C.A. and C.R. Wershler. 1981. Status and breeding of 

Mountain Plovers (Charadrius montanus) in Canada. Canadian 

Field Naturalist. 95:133-136.

Wallmo, O.C. 1957. A study of blues. Texas Game and Fish. 15:4-

7.

Wallmo, O.C. and P.B. Uzell. 1958. Ecological and social problems 

in quail management in west Texas. Transactions of the North 

American Wildlife Conference. 23:320-327.

Warner, R.E. 1992. Nest ecology of grassland passerines on road 

rights-of-way in central Illinois. Biological Conservation. 59:1-

7.

Warner, R.E. 1994. Agriculture, land use, and grassland habitat 

in Illinois: future shock for midwestern birds? Conservation 

Biology. 8:147-156.

Warner, R.E. and S.L. Etter. 1989. Hay cutting and the survival of 

pheasants. Journal of Wildlife Management. 53:455-461.

Warnock, R. 1997. Is habitat fragmentation a factor in the decline 

of the Burrowing Owl in Saskatchewan? Blue Jay. 55:222-228.

Warnock, R.G. and P.C. James. 1997. Habitat fragmentation and 

Burrowing Owls (Speotyto cunicularia) in Saskatchewan. Pages 

477-486. In: J.R. Duncan, D.H. Johnson, and T.H. Nicholls, eds. 

Biology and conservation of owls of the Northern Hemisphere. 

USDA Forest Service, North Central Forest Experiment Station, 

General Technical Report NC-190, St. Paul, Minnesota.

Weatherhead, P.J. 1989. Sex ratios, host-specific reproductive success, and impact of brown-headed cowbirds. Auk. 106:358-366.

Weaver, T., E.M. Payson, and D.L. Gustafson. 1996. Prairie ecology: 

the shortgrass prairie. Pages 67-76. In: F.B. Samson and F.L. 

Knopf, eds. Prairie conservation: preserving North America’s most 

endangered ecosystem. Island Press, Washington, D.C.

Webb, E.A. and C.E. Bock. 1990. Relationship of the Botteri’s 

sparrow to sacaton grassland in Southeastern Arizona. Pages 

199-209. In: P.R. Krausman and N.S. Smith, eds. Managing 

wildlife in the Southwest. Arizona Chapter of the Wildlife Society, 

Phoenix, Arizona.

Webster, F.S. 1979. The nesting season.American Bird. 33:879.

Wedgwood, J.A. 1976. Burrowing Owl in south-central Saskatchewan. 

Blue Jay. 34:26-44.

Wedgwood, J.A. 1978. The status of the Burrowing Owl in Canada. 

A report prepared for the Committee on the Status of Endangered 

Wildlife in Canada, Ottawa, Canada.

Wellicome, T.I. 1994. Is reproduction in Burrowing Owls limited by 

food supply? Picoides. 7:9-10.

Wellicome, T.I. 1997. Status of the burrowing owl (Speotyto cunicularia hypugaea) in Alberta. Wildlife Status Report No. 11. 

Alberta Environmental Protection, Wildlife Management Division, 

Edmonton, Alberta.

Wellicome, T.I. and E.A. Haug. 1995. Second update of status report 

on the Burrowing Owl. Committee on the Status of Endangered 

Wildlife in Canada, Ottawa, Ontario.

Wellicome, T.I., G.L. Holroyd, K. Scalise, and E.R. Wiltse. 1997. 

The effects of predator exclusion and food supplementation on 

Burrowing Owl (Speotyto cunicularia) population change in 

Saskatchewan. Pages 487-497. In: J.R. Duncan, D.H. Johnson, 

and T.H. Nicholls, eds. Biology and conservation of owls of the 

Northern Hemisphere. USDA Forest Service, North Central 

Forest Experiment Station, General Technical Report NC-190, 

St. Paul, Minnesota.

Weltzin, J.F., S. Archer, and R.K. Heitschmidt. 1997. Small-mammal regulation of vegetation structure in a temperate savanna. 

Ecology. 78:751-763.

Wershler, C., W.W. Smith, and C. Wallis. 1991. Status of the 

Baird’s Sparrow in Alberta: 1987/1988 update with notes on 

other grasslands sparrows and Sprague’s Pipit. Pages 87-89. In: 

G.L. Holroyd, G. Burns, and H.C. Smith, eds. Proceedings of the 

second endangered species and prairie conservation workshop. 

Natural History Occasional Paper 15, Provincial Museum of 

Alberta, Edmonton, Alberta.

Westemeier, R.L. 1985. Management plan for native prairie-chicken 

in Illinois. Illinois Natural History Survey, Springfield, Illinois.

Whicker, A.D. and J.K. Detling. 1988. Ecological consequences of 

prairie dog disturbances. BioScience. 38:778-785.

White, D.H., K.A. King, C.A. Mitchell, and A.J. Krynitsky. 1981. 

Body lipids and pesticide burdens of migrant Blue-winged Teal. 

Journal of Field Ornithology. 52:23-28.

White, D.H. and A.J. Krynitsky. 1986. Wildlife in some areas of 

New Mexico and Texas accumulate elevated DDE residues, 

1983. Archives of Environmental Contaminants and Toxicology. 

15:149-157.

Wickersham, C.W. 1902. Sickle-billed curlew. Auk. 19:353-356.

Wiens, J.A. 1970. Avian populations and patterns of habitat occupancy at the Pawnee site, 1968-1969. U.S. International Biological 

Program, Grassland Biome Technical Report 63. Colorado State 

University, Fort Collins, Colorado.

Wiens, J.A. 1971. Avian ecology and distribution in the comprehensive network, 1970. U.S. International Biological Program, 

Grassland Biome Technical Report 77. Colorado State University, 

Fort Collins, Colorado.

Wiens, J.A. 1973. Pattern and process in grassland bird communities. Ecological Monographs. 43:237-270.

Wilcove, D.S. 1985. Nest predation in forest tracts and the decline 

of migratory songbirds. Ecology. 66:1211-1214.

Wildeman, G. and J.H. Brock. 2000. Grazing in the southwest: 

history of land use and grazing since 1540. Pages 1-26. In: R. 

Jemison and C. Raish, eds. Livestock management in the American 

Southwest: ecology, society, and economics. Elsevier, Amsterdam, 

The Netherlands.

Williams, F.C. and A.L. LeSassier. 1968. Cassins Sparrow. Pages 

981-990. In: O.L. Austin, Jr., ed. Life histories of North American 

cardinals, grosbeaks, buntings, towhees, finches, sparrows, and 

allies. Dover, New York, New York.

Winter, M. 1999. Relationship of fire history to territory size, breeding 

density and habitat of Baird’s Sparrow in North Dakota. Studies 

in Avian Biology. 19:171-177.

Winter, M., D.H. Johnson, and J. Faaborg. 2000. Evidence for edge 

effects on multiple levels in tallgrass prairie. Condor. 102:256-

266.

With, K.A. and D.R. Webb. 1993. Microclimate of groundnests: the 

relative importance of radiative cover and wind breaks for three 

grassland species. Condor. 95:401-413.

Wobeser, G. and A.G. Wobeser. 1992. Carcass disappearance and 

estimation of mortality in a simulated die-off of small birds. 

Journal of Wildlife Diseases. 28:548-554.

Woffinden, N.D. and J.R. Murphy. 1989. Decline of a Ferruginous 

Hawk population: a 20-year summary. Journal of Wildlife 

Management. 53:1127-1132.

Wolf, L.L. 1977. Species relationships in the avian genus Aimophila. 

Ornithological Monographs. 23.

Wray II, T., Strait, K.A., and R.C. Whitmore. 1982. Reproductive 

success of grassland sparrows on a reclaimed surface mine in 

West Virginia. Auk. 99:157-164.

Wydeven, A.P. and R.B. Dahlgren. 1985. Ungulate habitat relationships in Wind Cave National Park .Journal of Wildlife 

Management. 49:805-813.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 137

Yahner, R.H. 1983. Seasonal dynamics, habitat relationships, and 

management of avifauna in farmstead shelterbelts. Journal of 

Wildlife Management. 47:85-104.

York, J.C. and W.A. Dick-Peddie. 1969. Vegetation changes in 

southern New Mexico during the past hundred years. In: W.G. 

McGinnies and B.J. Goldman, eds. Arid lands in perspective. The 

American Association for the Advancement of Science, Washington, 

D.C., and The University of Arizona Press, Tucson, Arizona.

Yosef, R. 1996. Loggerhead Shrike (Lanius ludovicianus). In: A. 

Poole and F. Gill, eds. The Birds of North America, No. 231. The 

Academy of Natural Sciences, Philadelphia, Pennsylvania, and 

The American Ornithologists’ Union, Washington, D.C.

Yosef, R. and T.C. Grubb. 1994. Resource dependence and territory 

size in Loggerhead Shrikes. Auk. 111:465-469.

Zimmerman, J.L. 1983. Cowbird parasitism of Dickcissels in different habitats and at different nest densities. Wilson Bulletin. 

95:7-22.

Zwartjes, P.W., J-L.E. Cartron, P.L.L. Stoleson, W.C. Haussamen, 

and T.E. Crane. 2005. Ecological interactions of ungulate grazing 

and native species of the Southwest: terrestrial wildife. Gen. 

Tech. Rep. RMRS-GTR-142. Fort Collins, CO: U.S. Department 

of Agriculture Forest Service Rocky Mountain Research Station. 

71 p. plus CD.
138 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Appendix 4-A. Common and scientific names of birds in text, listed in 

 alphabetical order of the scientific names

Common name Scientific name

Northern harrier Circus cyaneus

Swainson’s hawk Buteo swainsoni

Ferruginous hawk Buteo regalis

Golden eagle Aquila chrysaetos

American kestrel Falco sparverius

Northern Aplomado falcon Falco femoralis septentrionalis

Prairie falcon Falco mexicanus

Ring-necked pheasant Phasianus colchicus

Sage grouse Centrocercus urophasianus

White-tailed ptarmigan Lagopus leucurus

Sharp-tailed grouse Tympanuchus phasianellus

Greater prairie-chicken Tympanuchus cupido

Lesser prairie-chicken Tympanuchus pallidicinctus

Scaled quail Callipepla squamata

Northern bobwhite Colinus virginianus

Masked bobwhite Colinus virginianus ridgwayi

Montezuma quail Cyrtonyx montezumae

Killdeer Charadrius vociferus

Mountain plover Charadrius montanus

Upland sandpiper Bartramia longicauda

Long-billed curlew Numenius americanus

Common snipe Gallinago gallinago

Wilson’s phalarope Phalaropus tricolor

Mourning dove Zenaida macroura

Burrowing owl Athene cunicularia

Short-eared owl Asio flammeus

Lesser nighthawk Chordeiles acutipennis

Common nighthawk Chordeiles minor

Common poorwill Phalaenoptilus nuttallii

Broad-tailed hummingbird Selasphorus platycercus

Ladder-backed woodpecker Picoides scalaris

Southwestern willow flycatcher Empidonax traillii extimus

Say’s phoebe Sayornis saya

Ash-throated flycatcher Myiarchus cinerascens

Cassin’s kingbird Tyrannus vociferans

Western kingbird Tyrannus verticalis

Scissor-tailed flycatcher Tyrannus forficatus

Loggerhead shrike Lanius ludovicianus

Bell’svireo Vireo bellii

Gray vireo Vireo vicinior

Black-capped vireo Vireo atricapillus

Chihuahuan raven Corvus cryptoleucus

Common raven Corvus corax
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 139

Horned lark Eremophila alpestris

Cactus wren Campylorhynchus brunneicapillus

Sedge wren Cistothorus platensis

Mountain bluebird Sialia currucoides

Northern mockingbird Mimus polyglottos

Bendire’s thrasher Toxostoma bendirei

Curve-billed thrasher Toxostoma curvirostre

American pipit Anthus rubescens

Sprague’s pipit Anthus spragueii

Yellow warbler Dendroica petechia

Common yellowthroat Geothlypis trichas

Yellow-breasted chat Icteria virens

Canyon towhee Pipilo fuscus

Rufous-winged sparrow Aimophila carpalis

Cassin’s sparrow Aimophila cassinii

Botteri’s sparrow Aimophila botterii

Brewer’s sparrow Spizella breweri

Black-chinned sparrow Spizella atrogularis

Vesper sparrow Pooecetes gramineus

Lark sparrow Chondestes grammacus

Black-throated sparrow Amphispiza bilineata

Sage sparrow Amphispiza belli

Lark bunting Calamospiza melanocorys

Savannah sparrow Passerculus sandwichensis

Grasshopper sparrow Ammodramus savannarum

Arizona Grasshopper sparrow Ammodramus savannarum ammolegus

Baird’s sparrow Ammodramus bairdii

Henslow’s sparrow Ammodramus henslowii

Le Conte’s sparrow Ammodramu leconteii

Fox sparrow Passerella iliaca

Song sparrow Melospiza melodia

White-crowned sparrow Zonotrichia leucophrys

McCown’s longspur Calcarius mccownii

Chestnut-collared longspur Calcarius ornatus

Pyrrhuloxia Cardinalis sinuatus

Varied bunting Passerina versicolor

Dickcissel Spiza americana

Bobolink Dolichonyx oryzivorus

Red-winged blackbird Agelaius phoeniceus

Eastern meadowlark Sturnella magna

Eastern meadowlark (Southwestern subspecies) Sturnella magna lilianae

Western meadowlark Sturnella neglecta

Brewer’s blackbird Euphagus cyanocephalus

Brown-headed cowbird Molothrus ater

Scott’s oriole Icterus parisorum

Brown-capped rosy-finch Leucosticte australis
140 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 141

Introduction_______________________

The world possesses an amazing number and variety of freshwater fishes. Fish, the most diverse taxon 

on Earth, exhibit more species than all vertebrate 

taxa combined (Hocutt and Wiley 1986, May 1988, 

Nelson 1994, Matthews 1998) (table 5-1). Worldwide, 

an estimated 28,500 fish species are in 57 orders 

and 482 families (Nelson 1994). The North American 

freshwater fish fauna is also relatively rich and 

highly diverse ranging from primitive forms such as 

the Petromyzontidae (lampreys) to more modern and 

advanced forms such as the Percidae (perches) (Miller 

1959, Gilbert 1976, Briggs 1986, Cavender 1986, 

Nelson 1994). Currently, there are about 1,000 known 

freshwater fish species representing 50 families in 201 

genera in an area from Canada and Alaska south to the 

Isthmus of Tehuantepec (Gilbert 1976, Lee and others 

1980, Briggs 1986, Mayden 1992a,b). Some of these 

fish families are endemic to North America whereas 

others are also found on bordering continents (Briggs 

1986). Despite persistent research for more than 100 

years, numbers of described fishes in North America 

continue to increase over time. For example, Lee and 

others (1980) described 770 North American fishes, 

whereas just over a decade later Mayden (1992a,b) 

listed 971.

Regionally significant differences in North American 

freshwater fish species numbers and diversity are 

evident. The fish fauna east of the Continental Divide, 

consisting of eight speciose genera that comprise the 

majority of freshwater fish species on the continent, 

is species rich compared to the fish fauna of the 

Southwestern grasslands of New Mexico, Texas, and 

Oklahoma and the fish fauna west of the Continental 

Divide (Smith 1981b, Briggs 1986, Minckley and others 

1986, Moyle and Herbold 1987, Etnier and Starnes 

1993). Concentrations or “hot spots” of fish species 

diversity east of the Divide occur in the Appalachian 

and Ozark highlands and in Eastern streams. Here, 

some river systems contain as many as 100 to 200 

fish species (Robison and Beedles 1974, Hocutt and 

others 1986, Schmidt 1986, Starnes and Etnier 1986, 

Mayden 1987, Robison and Buchanan 1988, Etneier 

and Starnes 1993, Matthews 1998). Diversity of fishes 

is also high throughout the Mississippi-Missouri River 

drainage, the region’s major basin, where at least 260 

Chapter 5:

Fishes of Southwestern Grasslands:

Ecology, Conservation, and Management

Bob Calamusso

Table 5-1. Fish versus tetrapod diversity. Numbers given are 

for described species worldwide in each taxon. (Adapted 

from Nelson 1994 and May 1988)

Species Numbers

Fishes 28,618

Amphibians 2,800

Reptiles 6,000

Birds 4,500

Mammals 4,500
142 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

species of fresh water fishes are found (Cross 1967, 

Smith 1981b, Cross and others 1986, Robison 1986, 

Moyle and Herbold 1987). Moving west into the western 

Mississippi Basin (fig. 5-1) the number of native fish 

species decline to about 235 and continues to decline as 

one moves westward into Western short-grass prairie 

and Southwestern grasslands east of the Continental 

Divide (fig. 5-2).

In contrast, the fish fauna west of the Continental 

Divide is characterized by relatively depauperate fish 

assemblages rich in endemic species, many of which 

are confined to a single spring, stream, or drainage 

(Minckley 1973, Pister 1974, Smith 1981a,b, Briggs 

1986, Minckley and others 1986, Moyle and Herbold 

1987). The fauna west of the Divide has only about 

one-fourth as many species as the waters of Eastern 

North America (Miller 1959, Briggs 1986, Minckley 

and others 1986). For example, Burr and Mayden 

(1992) suggest that the Colorado River basin has 

approximately 32 native fish species, whereas 375 

native fishes could be found in the Mississippi River. 

Matthews and Gelwick (1990) indicated that a typical 

reach (200 m) in a medium or large stream in Eastern 

North America contained 20 species of fish, whereas 

a reach in semiarid or arid lands of the Midwestern 

to Western North America may only contain eight 

to 12 species. Moyle and Herbold (1987) suggested 

that a sample of fishes in Western North America 

could contain fewer than 10 species, whereas in the 

Mississippi River basin two to three times as many 

fishes could be found.

The variation in fish species diversity and numbers 

between the Mississippi Basin and Southwestern 

grasslands east of the Continental Divide and the fish 

fauna west of the Continental Divide can be linked to 

a myriad of obvious and less apparent factors such 

as historic and recent evolutionary geology, effects 

of inland seas, climate, past and present erosive 

processes, and extant local fish faunas (Endler 1977, 

Briggs 1986, Minckley and others 1986, Smith and 

Miller 1986). In general, the fish fauna east of the 

Divide experienced a relatively calm geologic history. 

During the Pleistocene, the fish fauna of the East 

was able to avoid a high number of extinctions due to 

the region’s large watershed size, stability, and lack 

of barriers to fish migration to refugia during glaciation and subsequent recolonization of habitats as ice 

sheets retreated northward (Smith 1981). In contrast, 

the fish fauna of the eastern Rocky Mountain slopes 

and west of the Continental Divide experienced more 

intense geological events during their evolution. West 

Figure 5-1. Western Mississippi Basin.

Figure 5-2. Semiarid grasslands, deserts, 

and major rivers of the Southwestern United 

States.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 143

of the Divide, the fish fauna evolved in small isolated 

basins that were influenced by periods of extreme 

tectonic uplift, regional erosion, mountain building, 

volcanism, and increasing aridity (Dickinson 1981, 

Minckley and others 1986). In contrast to events east 

of the Continental Divide, these processes created 

barriers to fishes moving in search of refugia and 

prevented dispersal and recolonziation of waters 

subsequent to these geologic events (Smith 1981b, 

Dott and Batten 1981). Isolation, fragmentation, and 

allopatric speciation prevailed in this region. Many 

fish populations were extirpated, others were isolated 

or restricted to small habitats where genetic variation of the population was limited and speciation was 

allowed to proceed (Miller 1946a,b, 1959, Simberloff 

1966, Moyle 1976, Smith 1978, 1981a,b, Minckley 

and others 1986). These factors, coupled with loss of 

genetic variation, contributed to rapid evolutionary 

change (Hubbs and Miller 1948, Hubbs and others 

1974). These events account for a fish fauna that is 

characterized by special adaptations, endemism, and 

relatively depauperate populations.

Population Trends__________________

As illustrated, the Southwestern fish fauna evolved 

in a region with highly variable climate and a dynamic 

geologic history. These environmental influences are 

reflected in the specializations and unique adaptations 

of the native Southwestern fish fauna. Unfortunately, 

the adaptations that have allowed these fishes to survive the harsh hydrologic conditions of the Southwest 

uplands and grasslands also made them vulnerable to 

large-scale human-induced ecological changes (Johnson 

and Rinne 1982, Minckley and Deacon 1991).

Native fishes of the Southwestern United States 

have experienced a severe reduction in both range and 

numbers since the region was first surveyed in the 

19th century (Williams and others 1985). Significant 

fish faunal changes were first observed in the early 

1900s (Miller 1946a, 1961, Minckley and Deacon 1968, 

Minckley and Deacon 1991, Koster 1957, Tyus and 

others 1982, Sublette and others 1990). By this time, 

extirpation and decline of native forms due to habitat 

alteration and competition with nonnative fish species 

had already begun (Miller 1961, Scurlock 1998). As early 

as 1904, nonnative common carp (Cyprinus carpio), 

brown bullhead (Ameiurus nebulosus), mosquito fish 

(Gambusia affinis), and green sunfish (Lepomis cyanellus) were collected in the Gila River of Arizona and New 

Mexico (Miller and others 1991). By 1926, 50 percent 

of the native fish fauna had been extirpated in the Salt 

River near Tempe, AZ, and by the late 1950s only two 

(14 percent) of the 14 original native fishes remained 

(Minckley and Deacon 1968). By the turn of the century, 

four fish species inhabiting grassland stream reaches 

including shovelnose sturgeon (Scaphirhynchus platorynchus), spotted gar (Lepisosteus osculates), bonytail 

chub (Gila elegans), and freshwater drum (Aplodinotus 

grunniens), had been extirpated in New Mexico (Koster 

1957, Sublette and others 1990). Noticing this downward trend prompted Miller (1946b) to call for studies 

of Western fishes to ensure their persistence.

It was not until the middle of the 20th century, 

however, when large-scale continental and regional 

changes in the native fish faunas occurred and were 

finally recognized by the scientific community (Miller 

1961, 1972, Minckley and Deacon 1991). Miller 

(1961) examined and provided data on changes in 

the Southwestern fish fauna. In a more extensive 

effort, Miller (1972) provided a preliminary effort to 

identify threatened fishes of the 50 States and again 

highlighted the decline of Southwestern native fishes, 

provided reasons for the decline, and listed species 

of threatened fishes for each of the 50 States in the 

United States (Miller 1972). He listed 305 fishes as 

threatened. Nine fish species were listed as threatened 

in Arizona and five in New Mexico. Miller and others 

(1989) documented the extinction of three genera, 27 

species, and 13 subspecies of fishes from North America 

during the past 100 years.

Today, most native Western fishes have become 

listed as threatened or endangered by the USDI Fish 

and Wildlife Service, and others are given protection 

afford by individual States (Johnson and Rinne 1982, 

Williams and others 1989, Minckley and Douglas 1991, 

USFWS 2003). Of the 150 fishes recognized by Lee and 

others (1980) west of the Continental Divide, 122 are 

considered to be in danger of extirpation (Minckley and 

Deacon 1991). A large percentage (34 percent) of the 

fishes currently listed by the USDI Fish and Wildlife 

Service are from the Western United States (USFWS 

TESS 2003). Viability of many of these populations 

is questionable. Many species not afforded protection 

under the Endangered Species Act or by State regulations are also believed to be in decline.

 In New Mexico, a minimum of 66 and possibly as 

many as 70 native fishes were extant when EuroAmerican settlers arrived (approximately 1525) 

(Sublette and others 1990, Propst 1999). Using the 

latter value, since that time, two (3 percent) native 

fish species have gone extinct, 12 (17 percent) have 

been extirpated from the State, 25 (36 percent) are 

considered to be in decline, and 31 (44 percent) are 

listed as stable (table 5-2). Of the remaining native 

fish fauna, 23 (41 percent) are listed as threatened or 

endangered by the State of New Mexico (Propst 1999). 

Of these, 10 (18 percent) are listed as threatened and 

endangered by the USDI Fish and Wildlife Service, and 

one additional species is proposed for listing (table 5-2). 

Many of these are species found in reaches that flowed 

through grasslands of the Southwest (Koster 1957, 

Sublette and others 1990, Propst 1999).
144 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Table 5-2. Status of native fishes in New Mexico.

Species Status

Scientific name Common name USFWS NM

Scaphirhynchus platorynchus Shovelnose sturgeon Ex

Lepsiosteus oculatus Spotted gar Ex

Lepisosteus osseus Longnose gar 

Anguila rostrata American eel 

Dorosoma cepedianum Gizzard shad 

Oncorhynchus clarki virginalis Rio Grande cutthroat trout

Oncorhynchus gilae Gila trout En Th

Oncorhynchus clarki Colorado River cutthroat Ex

Astyanax mexicanus Mexican tetra Th

Agosia chrysogaster Longfin dace 

Campostoma anomalum Central stoneroller 

Cyprinella Formosa Beautiful shiner Ext

Cyprinella lutrensis Red shiner

Dionda episcopa Roundnose minnow

Extrarius aestivalis Speckled chub 

Macrhybopsis aestivalis tetranemus Arkansas River speckled chub Th

Gila elegans Bonytail chub Ex

Gila nigrescens Chihuahua chub En

Gila Pandora Rio Grande chub 

Gila nigrescens Chihuahua chub Th En

Gila intermedia Gila chub Ex

Gila robusta Roundtail chub En

Hybognathus amarus Rio Grande silvery minnow En En

Hybognathus placitus Plains minnow

Meda fulgida Spikedace Th Th

Notropis amabilis Texas shiner Ex

Notropis girardi Arkansas shiner Th En

Notropis jemezanus Rio Grande shiner

Notropis orca Phantom shiner Ext Ext

Notropis simus simus Bluntnose shiner Ext Ext

Notropis simus pecosensis Pecos Bluntnose shiner Th

Notropis stramineus Sand shiner Th

Phenacobius mirabilis suckermouth minnow Th

Phoxinus erythrogaster Southern redbelly dace En

Pimephales promelas Fathead minnow 

Platygobio gracilis Flathead chub 

Ptychocheilus lucius Colorado cikeminnow En En

Rhinichthys cataractae Longnose dace

Rhinichthys osculus Speckled dace

Semotilus atromaculatus Creek chub

Tiaroga cobitis Loach minnow Th Th

Carpoides carpio River carpsucker

Catostomus commersoni White sucker

Catostomus plebeius Rio Grande sucker

Catostomus clarki Desert sucker 

Catostomus discobolus Zuni bluehead sucker En

Catostomus insignis Sonoran sucker

Catostomus latipinnis Flannelmouth sucker

Cycleptus elongatus Blue sucker En

Xyrauchen texanus Razorback sucker

Ictiobus bubalus Smallmouth buffalo Th

Moxostoma congestum Gray redhorse

Ictalurus furcatus Blue catfish

Ictalurus lupus Headwater catfish

Ictalurus melas Black bullhead

Ictalurus punctatus Channel catfish

Pylodictus olivaris Flathead catfish
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 145

Similar to New Mexico, the native fish fauna of 

Arizona has been greatly modified over the past 120 

years. Of the 36 fishes native to the waters of Arizona, 

one (3 percent) is now extinct, three (8 percent) have 

been extirpated from the State, 21 (58 percent) are in 

decline, and 11 (25 percent) are considered stable (table 

5-3). Of the remaining native fishes, 25 (69 percent) are 

listed as Wildlife of Special Concern in Arizona, 19 (53 

percent) are listed as threatened or endangered by the 

USDI Fish and Wildlife Service, and one is classified 

as a candidate for listing (table 5-3).

The decline of native fishes and native fish communities can be linked to resource extraction, alteration of 

stream channels, and water diversion by indigenous 

peoples and Euro-Americans settling in the region 

(approximately 1521) (Minckley and Douglas 1991, 

Rinne and Platania 1995, Calamusso and Rinne 1999). 

Introduction of nonnative aquatic species and associated competition and hybridization have resulted in 

the decline, alteration, and extirpation of many native 

fish species and entire assemblages. Remaining populations of native fish are often isolated, then decline 

to low levels leading to categorization as threatened, 

endangered, or at risk (Miller 1961, Deacon and others 

1979, Williams and Sada 1985, Williams and others 

1985, 1989, Miller and others 1989, Rinne and Minckley 

1991). Often resource management agencies do not 

have complete knowledge of populations that in fact 

warrant listing or some type of protection.

Biological Threats: Introduction of 

Nonnative Fish Species_____________

Other than dewatering of a stream reach or desiccation of some type of standing body of water the 

primary threat to the persistence of native fishes in 

Southwestern grasslands has been the introduction and 

range expansion of nonnative fishes. In fact, Minckley 

and Douglas (1991) suggested that the spread of nonnative fishes may be more detrimental than all other 

environmental perturbations combined. The effects of 

dewatering, habitat loss, and pollution are sometimes 

temporary, but once a nonnative species is established 

it is difficult or impossible to remove it from many 

systems (Rinne and Turner 1991).

Nonnative fishes have been introduced into the 

waters of North America since the late 1600s. Fuller 

and others (1999) listed 536 fish taxa (representing 75 

families) that have been transplanted into the United 

States outside their native range. Unfortunately, rates 

of introduction have continued to increase over the 

past 45 years due to more efficient transportation 

and demand for nonnative forms for sport, fish farming, bait, and aquarium use (Fuller and others 1999, 

Heidinger 1999, Li and Moyle 1999, Nico and Fuller 

1999). Of the myriad of intentional fish stockings, only 

a few have been positive, and no unplanned introductions have been considered beneficial (Steirer 1992, 

Courtnay and Williams 1992). Introductions have, 

on the whole, caused many problems for native fishes 

(Deacon 1979, Deacon and others 1979, Williams and 

Table 5-2. Continued.

Species Status

Scientific name Common name USFWS NM

Cyprinodon pecosensis Pecos pupfish Th Th

Cyprinodon tularosa White sands pupfish Th

Cyprinodon sp. Palomas pupfish Ext Ext

Fundulus zebrinus Plains killifish

Lucania parva Rainwater killifish

Poeciliopsis occidentalis Gila topminnow En Th

Gambusia affinis Mosquito fish

Gambusia noblis Pecos gambusia En En

Lepomis cyanellus Green sunfish

Lepomis macrochirus Blugill

Lepomis meglotis Longear sunfish

Micropterus salmoides Largemouth bass

Etheostoma lepidum Greenthroat darter Th

Percina macrolepida Bigscale logperch Th

Aplodinotus grunniens Freshwater drum Ex

Cottus bairdi Mottled sculpin

En=Endangered, Th =Threatened, Ext = Extinct, Ex = Extirpated (For more detailed description of Federal and State threatened 

and endangered status of native fishes see Propst 1999.)
146 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

others 1989). The nonnative fishes currently found in 

almost all major watersheds of the Continental United 

States have led to the decline and listing of native fish 

populations across North America (Taylor and others 

1984, Courtenay and others 1984, Courtenay and 

Stauffer 1990, Hendrickson and Brooks 1991, U.S. 

Congress, OTA 1993). Today, 70 percent of the fishes 

listed under the Endangered Species Act have been 

negatively impacted by nonnative fishes (Lassuy 1995, 

USFWS 2003, Tyus and Saunders 2000). Further, it is 

estimated that the extinction of 20 native fishes in the 

1900s is primarily due to interactions with nonnative 

forms (Miller and others 1989).

Impacts of nonnative fish on native fishes/communities depend on the physiological, behavioral, and 

ecological capabilities of the nonnative species and the 

biological and physical components of the ecosystem 

and its inhabitants (Taylor and others 1984, Haines 

and Tyus 1990, Rinne and others 1998). Nonnative 

fish often have adaptations that provide advantages in 

the aquatic environment in which they are introduced. 

These capabilities (reproductive, environmental tolerances, behavior) enable them to exploit environmental 

resources more effectively than native inhabitants. This 

efficient exploitation can lead to higher reproductive 

success and greater growth rates, which then allow 

nonnative species to biologically “swamp” resident 

fishes. The end result is replacement of native forms 

by nonindigenous species.

Many examples of these phenomena exist in the 

Southwest. Native Gila topminnow (Poeciliopsis occidentalis occidentalis) have been eliminated through 

Table 5-3. Status of native fishes in Arizona.

Status

Species USFWS AZ

Elops affinis Machete

Mugil cephalus Striped mullet

Oncorhynchus gilae apache Apache trout Th X

Oncorhynchus gilae gilae Gila trout En X

Campostoma ornatum Mexican stoneroller

Cyprinella formosa Beautiful shiner Th X

Gila cypha Humpback chub En X

Gila ditaenia Sonora chub Th X

Gila elegans Bonytail chub En X

Gila intermedia Gila chub C X

Gila nigra Headwater chub 

Gila purpurea Yaqui chub En X

Gila robusta Roundtail chub X

Gila seminuda Virgin chub En X

Meda fulgida Spikedace Th X

Lepidomeda mollispinis mollispinis Virgin spinedace X

Lepidomeda vittata Little Colorado spinedace Th X

Plagopterus argentissimus Woundfin En X

Ptychocheilus lucius Colorado sikeminnow En X

Rhinichthys osculus Speckled dace

Tiaroga cobitis Loach minnow Th X

Catostomus bernardini Yaqui sucker X

Catostomus clarki Desert sucker

Catostomus discobolus Bluehead sucker

Catostomus d. yarrowi Zuni Mountain sucker

Catostomus insignis Sonora sucker

Catostomus latipinnis Flannelmouth sucker X

Catostomus sp. Little Colorado sucker

Xyrauchen texanus Razorback sucker En X

Ictalurus pricei Yaqui catfish Th X

Cyprinodon eremus Quitobaquito pupfish En X

Cyprinodon macularius Desert pupfish En X

Cyprinodon sp. Monkey spring pupfish Ext Ext

Poeciliopsis occidentalis occidentalis Gila topminnow En X

Poeciliopsis occidentalis sonoriensis Yaqui topminnow En X

En = Endangered, Th = Threatened, Ext = Extinct, X = Species of Special Concern by State of Arizona, Ex = Extirpated
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 147

aggression and predation by mosquito fish (Gambusia 

affinis) (Minckley and others 1977, Meffe and others 

1983, Weedman and others 1996, 1997a,b). Brown trout 

(Salmo trutta), a native trout of Europe, have wider 

environmental tolerances, especially water temperature, and are more aggressive and piscivorous. They 

have replaced native trout across North America, and 

in the Southwest have been implicated in the demise 

of Rio Grande cutthroat trout (Oncorhynchus clarki 

virginalis), Gila trout (O. gilae), and Apace trout (O. 

apache) (Behnke and Zarn 1976). Red shiner has been 

implicated in the replacement of native fishes through 

competitive exclusion throughout low elevation grassland streams in the Southwest. Predation by species 

nonnative channel (Ictalurus punctatus) and flathead 

catfish (Pylodictus olivaris) directly prey on larval 

young and adult native fishes. Both of these species 

have been implicated in the demise of the razorback 

sucker (Xyrauchen texanus) (Marsh and Langhorst 

1988). Some species of introduced chiclid have yearround breeding seasons. These fish reproductively 

swamp native fishes by sheer number, which appears 

also to be the case in the replacement of Rio Grande 

sucker (Catostomus plebeius) by nonnative white sucker 

(Catostomus commersoni) in the Rio Grande drainage 

of Colorado and New Mexico.

Habitat alteration, it is hypothesized, often allows 

nonnative fishes to dominate indigenous fish faunas 

(Maddux and others 1993). Minckley and Carufel (1967) 

concluded that red shiner contributed to the demise 

of native fish populations throughout the Southwest. 

In many areas Red shiners have replaced woundfin 

(Plagopterus argentissimus), spikedace (Meda fulgida), 

and loach minnow (Rhinichthys (tiaroga) cobitis), all 

of which are Federally listed species in Arizona and 

New Mexico (Deacon 1988, Marsh and others 1989, 

Douglas and others 1994). While the mechanisms of 

replacement are not known, Rinne (1991) and Douglas 

and others (1994) indicated that red shiner adults 

occupied habitats used by juvenile spikedace and 

loach minnow in the Verde River. Nonnative trout 

have been shown to have detrimental affects on the 

Little Colorado River spinedace (Lepidomeda vittata 

vittata) in terms of predation and displacement (Blinn 

and others 1993). Centrarchids, especially, have been 

problematic for native fishes. This taxon has been 

introduced throughout Southwestern streams for 

sport, and by nature they are predacious, which has 

had severe impacts on all native forms. In the Verde 

River of Arizona, smallmouth bass (Micropterus dolmieui) and Green sunfish (Lepomis cyanellus) have 

eliminated a complete native fish fauna (Rinne and 

others 1999).

Similar to continental trends, establishment of 

nonnative fishes and replacement of native fishes 

has occurred on a regional scale throughout the 

Southwestern grasslands since the late 1800s (Rinne 

1990, Fuller and others 1999). Nonnative fishes were 

imported for sport, biological control, and bait along 

with accidental introductions of unwanted species 

(Deacon and others 1964). Currently, 84 nonindigenous 

fishes are now found in the waters of Arizona and 78 

in New Mexico (tables 5-4 and 5-5) (Minckley 1973, 

Hanson 1980, Sublette and others 1990, Fuller and 

others 1999). In New Mexico, introduced fishes account for more than half of the total fish fauna in the 

Rio Grande, Mimbres, Tularosa, San Juan, Zuni, San 

Francisco, and Gila Rivers (Propst and others 1987, 

Platania 1991, Sublette and others 1990, Rinne and 

Platania 1995, Propst 1999). In the Pecos and Canadian 

Rivers, nonnative fishes comprise 42 percent and 48 

percent of the current fish fauna, respectively (Sublette 

and others 1990). Similar effects are seen in Arizona. 

Nonnative fishes comprise 82 percent of the fauna 

in the Colorado River basin, 68 percent in the Little 

Colorado, 50 percent in the San Juan, 40 percent in 

the Gila, 38 percent in the Bill Williams, 25 percent 

in the Rio Yaqui basin, and 10 percent in the Virgin 

River. The Sonotya and Rio Magdalena basins are free 

of nonnative fishes.

Native fish of the montane and grassland reaches 

of the Eastern Rocky Mountains and Western North 

America may be at greater risk from the introduction of 

nonnative species than native fishes in other parts of the 

country because of their isolated evolutionary history 

(Moyle and others 1986, Minckley and Douglas 1991). 

Additionally, depauperate native Western species live 

in dynamic watershed and aquatic environments, both 

of which make them vulnerable to replacement by nonnative fishes (Minckley and Deacon 1968, Moyle and 

Nichols 1973, 1974, Moyle 1976). Many native fishes 

in the West have not developed protective mechanisms 

against predation, general competition, and genetic 

hybridization, and they have low tolerance to biological interactions with nonnative fishes (Miller 1961, 

Deacon and Minckley 1974, Deacon 1979, Minckley 

and others 1986, Smith and Miller 1986, Sublette and 

others 1990). To a large degree, many Southwestern 

fishes exist as relict populations in single habitats, 

which exacerbates their precarious state (Deacon and 

Minckley 1974).

Effects of nonnative fishes on native fish communities are not well understood and are typically observed 

from major alterations in the native fish fauna. Direct 

effects from the interactions with nonnative fishes come 

in the form of predation (Schoenherr 1981, Minckley 

1983, Blinn and others 1993), competition (Deacon and 

others 1964), hybridization (Rinne and Minckley 1985, 

Dowling and Childs 1992), elimination, reduced growth 

and survival, changes in community structure, and no 

effect (Moyle and others 1986). Impacts on various life 

stages also occur in time and space. Indirect effects, 

such as stress or other behavioral interactions are more 

difficult to document. The outcome of many of these 
148 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Table 5-4. Current distribution introduced nonnative grassland fishes by drainage basin in New Mexico. (Table 

adapted from Sublette and others 1990)

Species Rio Grande Canadian Pecos San Juan Mimbres Tularosa Gila

Threadfin shad X X X X

Cutthroat trout X X X X

Coho Salmon X X X X

Rainbow trout X X X X X X

Kokanee salmon X X X X X

Brown trout X X X X X X

Brook trout X X X X X

Lake trout X

Northern pike X X X X

Zebra danio X

Goldfish X

Grass carp

Carp X X X X X

Fathead minnow X X X X X X

Bullhead minnow X 

Yellow bullhead X X X

Guppy X

Sailfin molly X

Inland silverside X

White bass X X X

Striped bass X X

Rock bass X

Warmouth X X

Smallmouth bass X X X X X

Spotted bass X 

White crappie X X X X X

Black crappie X X X

Walleye X X X

Key: X = presence.

Table 5-5. Current distribution of nonnative grassland fishes by drainage basin in Arizona.

 Little Bill

Species Colorado Colorado Williams San Juan Sonoyta Yaqui Gila Virgin

White sturgeon

American eel X X

Threadfin shad X X

Coho salmon

Sockeye salmon

Rainbow trout X X X X

Cutthroat trout

Brown trout

Brook trout

Arctic grayling

Northern pike

Banded tetra X ?

Carp X X X X X

Goldfish 

Grass carp

Golden shiner X X X X
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 149

Table 5-5. Continued.

 Little Bill

Species Colorado Colorado Williams San Juan Sonoyta Yaqui Gila Virgin

Utah chub X ? 

Red shiner X X X X X

Redside shiner X ? X X ?

Sand shiner X ?

Fathead minnow X X X X X

Bigmouth buffalo X

Black buffalo X

Smallmouth buffalo X

Flathead catfish X X

Channel catfish X X X X

Blue catfish X ?

Black bullhead X X X X

Yellow bullhead X X X

Brown bullhead X

Plains Killifish X X

Mosquitofish X X X X X X

Pacu

Variable platyfish

Green swordtail

Sailfin molly X X

Mexican molly X X

Guppy X X

Striped bass X

White bass X

Yellow bass X

Smallmouth bass X X

Spotted bass X X

Largemouth bass X

Warmouth

Green sunfish X X X X X

Bluegill X X X X X

Redear sunfish X X

Pumpkinseed X

Rockbass

White crappie X

Black crappie X X X

Sacramento perch 

Walleye X

Yellow perch X X X

Sargo

Bairdiella

Corvina

Convict cichlid X

Mozambique mouthbrooder X X

Blue talapia

Red-breasted tilapia X

Striped mullet X

Spotted sleeper X

Longjaw mudsleeper X

Mottled sculpin X

Key: X = presence. X? = questionable.
150 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

negative interactions is that remaining populations 

of native fish are often isolated, decline to low levels 

leading to categorization as threatened, endangered, or 

at risk (Miller 1961, Deacon and others 1979, Williams 

and others 1985, Miller and others 1989, Rinne and 

Minckley 1991).

Habitat Threats____________________

General Landscape Change and Habitat 

Loss

Beginning in the 1880s, marked ecological changes 

over the Southwestern landscape occurred. By the 

1890s, regional degradation was evident (Hasting 

and Tuner 1965). Dense stands of Southwestern 

grasslands, lush riparian areas, free flowing rivers and 

unaltered watersheds that were in place at the time 

of the Spanish explorers and later by Anglo settlers, 

had long since been irrevocably altered by livestock 

grazing and the damming and diverting of river water 

for human consumption (Griffiths 1904, Mehrhoff 1955, 

Scurlock 1998, Ambos and others 2002). Cottonwood 

gallery forests that once lined Southwestern rivers 

died due to channel incision and dropping water tables 

and were being replaced by nonnative phreatophytic 

species such as salt cedar (Tamarisk sp.).

Examples were noted throughout the Southwest. The 

San Pedro River, once a slow moving river with cienegas, 

had been changed to a steep-walled entrenched channel 

through which water moved rapidly, thus transporting large amounts of sediment into the Gila River. By 

the end of the 19th century, the once perennial San 

Pedro River had changed to an intermittent system 

exhibiting dry reaches throughout the year. Events 

such as these were relatively unknown prior to 1890 

(Hastings and Turner 1965). In the grasslands of New 

Mexico, similar events occurred in the Rio Puerco (Rio 

Grande drainage) and Rio Chaco (San Juan drainage) 

watersheds (Bryan 1928, Scurlock 1998).

Water Use

The original aquatic habitats of the Southwestern 

grasslands once consisted of large rivers (5th order), 

streams, cienegas, and springs. Since settlement 

by Native Americans and Euro-Americans, these 

resources have undergone extreme modification 

(Leopold and others 1964, Hastings and Turner 1965, 

Sublette and others 1990, Minckley and Deacon 1991). 

Water development and its use for agriculture began 

as early as 900 A.D. when pueblo dwelling Native 

Americans collected surface runoff or diverted water 

from Southwestern rivers for agriculture and domestic 

uses. As the Spanish-Americans began settling the 

region (approximately 1521), increased demand was 

placed on limited Southwestern waters resources 

(Dozier 1983, Carlson 1990, Scurlock 1998). The use 

of metal tools, domesticated animals, and engineering 

skills used by Spanish-American settlers made even 

a greater impact on water resources in the region. In 

New Mexico, the acequia or irrigation ditch was so 

important that they were dug prior to the construction of human dwellings. By the end of the 1600s, all 

or portions of present-day New Mexico’s perennial 

streams were being diverted into irrigation networks 

(Scurlock 1998). As Anglo-Americans (nonnative, nonHispanic peoples) entered the Southwest in the early 

1820s, more demand was laced on water resources 

for expanding commercial agricultural and economic 

interest (Hastings 1959, Scurlock 1998). New irrigation 

systems were constructed, and older systems expanded 

by private companies whose economic well being was 

tied directly to the growing economy (Briggs and Van 

Ness 1987, Clark 1987, Scurlock 1998). Technological 

advances allowed wells to be dug that tapped underlying aquifers. Windmills also gave access to underground 

water resources. By 1880 water shortages had begun 

in the Southwest (Westphall 1965).

Large-scale change in natural stream flow and surface water dispersion occurred during the latter part 

of the 19th and beginning of the 20th centuries when 

dams were built on the major rivers of New Mexico 

and Arizona (Mueller and Marsh 2002). Then, with the 

passage of the Reclamation Act of 1903, the free-flowing waters and cienegas of the American Southwest 

were changed forever (Hastings and Tuner 1965). The 

perennial shortage of water for human use inherent in 

the region along with the post-World War II population boom precipitated the need for larger dams and 

diversion systems. Large-scale dams and diversion 

projects were built to store water, control sedimentation 

and flooding, generate power, and disperse water for 

irrigation. Today, dams and diversions are found on all 

major rivers and many of their tributaries throughout 

the Southwest (fig. 5-3 and 5-4). Large, artificial impoundments now make up the major surface waters in 

the Southwestern grasslands (Minckley 1973, Sublette 

and others 1990). Expanding agriculture, especially in 

the lower elevation grassland valleys, has necessitated 

the pumping of groundwater because of the shortages 

in surface water allocations; often times the result is 

the extraction of water that exceeds rates at which 

water recharges. Some aquifers are essentially being 

“mined” for their fossil water. Pumping of aquifers has 

placed many rivers and springs in the grasslands of 

the Western United States at risk by lowering regional 

and local water tables.

Interbasin transfer projects, such as the Central 

Arizona Project (fig. 5-5) now transport water across 

hundreds of kilometers into foreign basins. Natural 

drought cycles (one of which at publication time is in 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 151

Figure 5-3. Major dams and diversions on 

rivers in Arizona.

Figure 5-4. Major dams and diversions on 

rivers in New Mexico.
152 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

progress in the Southwest) add additional stress to 

a system that is largely overallocated and overused. 

Though completed with the best of intentions, all of 

these activities have disrupted the natural flows of 

grassland rivers and altered their physical properties. 

Today much of Arizona’s and New Mexico’s grassland 

river flows are found in canals, ditches, and drains.

In New Mexico, major dams or diversions exist on 

every large mainstem river in the State (table 5-6; fig. 

5-4). There are also 32 small watershed projects with 

71 flood control dams and 41 miles of diversions that 

were constructed since 1957. On the Rio Grande there 

are 13 major dams and diversions in New Mexico. 

Along the upper and middle reaches of the Rio Grande, 

six dams and three major diversions have been built. 

There are 321 km of canals (9.6 km concrete lined), 928 

km of laterals (6 km lined), and 648 km of open and 

concrete pipe drains. The objective of these projects 

is to provide irrigation water, control sediment, and 

prevent flooding in the Middle Rio Grande Valley. Over 

36,000 ha of agricultural lands are irrigated by water 

diverted from this reach of the Rio Grande. On the lower 

Rio Grande in New Mexico and West Texas, there are 

two major dams, six diversion dams, 222 km of canals, 

731 km miles of laterals, 744 km of drains, and one 

hydroelectric power generating facility. Most drains 

in this reach of river are unlined earthen structures. 

These structures provide irrigation water for 79,320 ha 

of farm land in addition to power generation for cities 

and industry in the region. Water is also provided to 

irrigate 10,117 ha in the Juarez Valley, Mexico. Today, 

the lower Rio Grande is essentially a conveyance ditch 

from the headwaters of Elephant Butte to the Texas 

border. In the grassland reaches of the Pecos River, four 

dams have been constructed in New Mexico. There are 

242 km of laterals, 59 km of canals, and about 38 km 

of drains in this system. Water stored in Pecos River 

reservoirs irrigates almost 10,000 ha in the region.

As with New Mexico, all major rivers and many 

tributaries in Arizona now are impounded, controlled, 

diverted, and in some cases overallocated (table 5-7; fig. 

5-3). On the Colorado River there six dams, on the Bill 

Williams River one, four on the Salt River, two on the 

Verde River, and two on the Gila River. Impoundments 

on the Colorado River and its tributaries now account 

for the major surface waters in Arizona. The Colorado 

River—the once turbid and swift-flowing water that 

meandered through braided channel system bordered 

by riparian vegetation and marshes—has now been 

dammed, diverted, channelized, and confined by levees 

(Mueler and Marsh 2002). The largest artificial bodies 

of water in Arizona—Lakes Havasu, Powell, Mead, and 

Mohave—are all located on the Colorado River. Remote 

reaches of the Colorado River retain characteristics of 

a free-flowing stream as does the Salt River Canyon 

and the upper reaches of the Verde River.

Effect on Fishes—Water, above all, is of course the 

key habitat element needed to sustain fishes (Heede 

Figure 5-5. Central Arizona Project.

Table 5-6. Dams and diversions on major rivers in the New 

Mexico.

Structure River Date built

Heron Rio Grande 1971

El Vado Rio Grande 1935

Abiquiu Resrvoir Rio Grande 1963

Cochiti Dam Rio Grande 1975

Galisteo Dam Rio Grande 1970

Nambe Falls Dam Rio Nambe 1974

Jemez Canyon Dam Jemez 1953

Angostura Diversion Rio Grande 1938

Isleta Diversion Rio Grande 1934

San Acacia Diversion Rio Grande 1934

Elephant Butte Dam Rio Grande 1916

Caballo Dam Rio Grande 1938

Leasburg Diversion Rio Grande 1919

McMillan Dam Pecos 1908

Brantley Dam Pecos 1991

Avalon Dam Pecos 1888

Sumner Dam Pecos 1937
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 153

and Rinne 1990, Rinne 1990, 2000). The impact of 

impounding, diverting, dewatering, and draining of 

Southwestern waters along with mining of ancient aquifers has eliminated or drastically altered the aquatic 

habitats needed for survival of native Southwestern 

grassland fishes. The affect of these activities has accelerated the early decimation of native fishes found 

in grassland reaches throughout Southwestern rivers 

(Minckley and Deacon 1968).

Damming—Dams present a suite of problems for 

native fishes. First order impacts begin soon after the 

dam is in place (Neel 1963, Orth and White 1993). 

The physical structure acts as a barrier to upstream 

migration, and the impoundment can prevent downstream migration. Natural cycles of flooding and 

drought flows are now determined by power and water 

demands of cities and agriculture (Johnson and Rinne 

1982). Other first order effects are altered sediment 

load, changes in water quality, and plankton (fig. 5-6). 

Second order impacts alter channel morphology and 

change substrate composition. Third order impacts 

relate to the cumulative effects of first and second order 

effects. It is the third order effects that impact fish 

and invertebrates. The physical structure interferes 

or prevents upstream migration, and often times the 

impoundment above precludes downstream migration 

due to vastly different water quality than preimpoundment conditions. For example, the Rio Grande once had 

large runs of female American eel (Anguilla rostrata). 

Other species affected by dams on the Rio Grande 

and Pecos Rivers were freshwater drum (Aplodintus 

grunniens), river carpsucker (Carpoides carpio), blue 

sucker (Cycleptus elongates), grey redhorse (Moxostoma 

congestum). Dams on the Lower Colorado system have 

been particularly destructive to the long-lived big 

river fishes such as flannelmouth sucker (Catostomous 

latipinnis), razorback sucker (Xyrauchen texanus), 

Colorado pikeminnow (Ptychoceilus lucius), humpback 

chub (Gila cypha), and bonytailed chub (G. elegans). 

Dams alter the riverine environment upstream and 

downstream of the structure. Stratification and point 

of release from the dam also determine water quality 

and temperature (fig. 5-7). For example, hypolimnetic 

releases have low relative water temperatures, are 

often anoxic, and can also have elevated nutrient, 

iron, manganese, and hydrogen sulfide concentrations (Reid and Wood 1976, Orth and White 1993). 

These conditions may be unsuitable for native fishes 

and native macroinvertebrates, but may be excellent 

for introduced cool-water game fishes. In addition to 

unfavorable habitat conditions, native fish now face 

the threat of predation and competition from these 

alien forms. Releases such as these occur at Stewart 

Mountain Dam on the Salt River and Caballo Dam on 

Figure 5-7. Impacts of dams on upstream and downstream 

water temperature.

Figure 5-6. Impacts of impoundments on riverine systems.

Table 5-7. Major dams and diversions of the Lower Colorado 

River Basin, Arizona.

Structure River Date built

Granite Reef Diversion Salt River 1908

Theodore Roosevelt Dam Salt River 1911

Horse Mesa Dam Salt River 1927

Mormon Flat Dam Salt River 1925

Stewart Mountain Dam Salt River 1930

Horseshoe Dam Verde River 1946

Bartlett Dam Verde River 1939

Laguna Dam Colorado River 1909

Boulder Dam Colorado River 1936

Imperial Dam Colorado River 1938

Parker Dam Colorado River 1938

Headgate Rock Diversion Colorado River 1944

Palo Verde Diversion Colorado River 1957

Coolidge Dam Gila River 1928

Painted Rock Dam Gila River 1960

New Waddell Dam Agua Fria River 1928

Alamo Dam Bill Williams River 1968 
154 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

the Rio Grande. Water chemistry and conditions that 

once favored native grassland fishes below Elephant 

Butte and Caballo Reservoirs now support species coolwater species such as trout and walleye (Stizostedion 

vitreum), along with various sunfishes and catfishes. 

In many reaches of Southwestern rivers (such as the 

Rio Grande, Salt, Colorado), dams have dewatered 

hundreds of kilometers of river channel while permanently inundating others. Flood plain areas that 

may have acted as nursery or rearing areas are now 

inaccessible along most major rivers.

Diversion—Diversion of stream waters during 

irrigation season desiccates many riverine habitats 

causing local extirpations and extinctions in reaches 

that have become permanently dewatered. Impacts 

are more severe during drought years. For example, 

Rio Grande silvery minnow living below the San Acia 

diversion on the Rio Grande often run the risk of coping 

with decreasing flows and drying of the channel during 

drought years. Fish are unable to ascend the diversion 

as larvae. As the stream dries many individuals are lost 

and recruitment is reduced. Similar situations occur 

on the Salt River below Stewart Mountain Dam and 

in the lower reaches of the Colorado River near the 

Mexican border. During irrigation season many fishes 

become stranded in irrigation canals as water flows 

are decreased and headgates are closed. The effects 

of dewatering appear too be more pronounced in the 

middle to lower elevations of Southwestern streams 

(fig. 5-3 and 5-4). These reaches flow through grassland 

and desert ecotypes that are farther from headwater 

sources and are more prone to natural dewatering. The 

placement of diversion systems above these reaches 

increases the probability of dewatering. Canals and 

laterals often times are earthen structures, but many 

miles of these are concrete lined. Lining canals with 

concrete, while increasing irrigation efficiency, prevents 

water from filtering into the water table. Seepage 

from unlined irrigation systems recharges aquifers, 

flows into drains, and recharges rivers from which 

it was diverted, thus supporting fish and riparian 

communities.

Earthen canals are regularly dredged to clean and 

maintain efficient flow of water. During this operation, 

riparian vegetation is either partially or completely 

removed so that equipment can complete the dredging 

operation. Fish communities in these systems (common along the Ro Grande) often harbor native fishes. 

Applications of herbicides to control aquatic weeds 

and pesticides to control mosquitoes are common. 

Such chemicals alter water quality and often make 

conditions uninhabitable for certain fishes (Minckley 

and Deacon 1968).

Irrigation systems present a suite of problems for 

native fishes that once were abundant in the unaltered 

river systems from which present systems are watered. 

The irrigation systems that are often lined with concrete 

and are highly regulated cause organic components that 

would typically be found in natural riverine systems 

to be almost nonexistent in the artificial environment. 

Mechanical and chemical vegetation removal and 

dredging of such systems are regularly conducted to 

clean, facilitate flow, and effect efficient transfer of the 

precious commodity—water. The threat of dewatering 

and drying is ever present. Even in primitive, earthen 

irrigation canals, aquatic habitats have been highly 

simplified, flow regimes altered, and the fish community 

impacted through the introduction of nonnative fishes 

and other alien aquatic species such as crayfish and 

other invertebrates. Though these problems for native 

fish are salient, irrigation systems do allow fishes 

to persist, even if that occurs at reduced levels. The 

drains that parallel the Rio Grande in New Mexico 

for all practical purposes harbor the fish community 

of its middle and lower reaches.

Pumping—Pumping of groundwater lowers water 

tables; this causes interruption in arid land streams. 

Springs and marshes can also become desiccated from 

lowering of water tables. The upshot is the demise of 

many aquatic habitats and the loss of native fishes 

populations One of the classic examples of this, while 

not in the grasslands reaches of Arizona or New Mexico, 

was the near extirpation of the Devil’s Hole pupfish 

(Cyprinodon diabolis). In 1969, corporate farm interests 

intended to pump the Ash Meadows aquifer, located in 

southeastern Nevada near Las Vegas,to increase farm 

production. The result would have been the lowering 

of the water table, which in turn would dewater the 

spring, ultimately leading to the extinction of this indigenous endemic pupfish species. Fortunately, for the 

Devil’s Hole pupfish, concerned scientists and citizens 

were able to bring a halt to this action.

Grazing, Riparian Corridors, and Fish 

Relationships______________________

The ecological conditions of Southwestern watersheds and riparian corridors have undergone dramatic 

changes since the settlement of the region by Native 

Americans and Euro-Americans (Carton and others 

2000). Many of these changes have been attributed to 

grazing of domestic livestock over the past 460 years. 

The first livestock grazed in the region were part 

of the Spanish expeditions that entered the region 

in 1539. Spanish settlers grazed 4,000 sheep, 1,000 

cattle, 1,000 goats, and 150 mares with colts in the 

Middle Rio Grande reach of present-day New Mexico 

around 1598 (Baxster 1987, Bayer and Montoya 1994, 

Scurlock 1998, Floyd and others 2003). Overgrazing 

was observed in the area as early as 1630 (Stewart 

1936, Scurlock 1998). From 1785 to 1900 some 6,000 

goats grazed the foothills of the Sandia Mountains. 

These lands later came under the jurisdiction of the 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 155

Cibola National Forest (Scurlock 1998). In addition to 

Spanish settlers, many contemporary Native American 

peoples grazed livestock. As many as 30,000 head of 

sheep grazed the lands of the Hopi Pueblos during 

the 1700s (Schickendanz 1980). By 1820 there was an 

estimated one million sheep grazing in New Mexico, and 

by the mid 1800s about three million sheep grazed the 

grasslands and riparian areas of that region (Baxter 

1987, Scurlock 1986, 1998). The ending of hostilities 

between the United States government and Native 

Americans along with the completion of transcontinental railways and their spur routes led to sharp 

increases in livestock from 1870 to 1890 (Scurlock 

1998). By 1880 five million sheep and 250,000 head of 

cattle grazed the grasslands and river valleys of New 

Mexico (Bailey and Bailey 1986).

Arizona experienced a long history of grazing by 

domestic livestock similar to New Mexico (Young 1998, 

Wildeman and Brock 2000). Spanish horses were first 

grazed in present-day Arizona in 1515, cattle in 1521. 

Goats and sheep, imported by Spanish settlers and 

expeditions, were grazed on the Arizona grasslands 

by 1700 (Barnes 1926, Peplow 1958). Peplow (1958) 

reported that the most important (and probably 

the most well stocked) ranches were at present-day 

San Bernardino, Babocomari, San Pedro, Arivaca, 

Calabasas, Sapori, Radenton, San Rafael de la Zanje, 

Sonoita, Tubac, and ranches in the San Simon Valley, 

Aqua Prieta, and Pueblo Viejo. After wars with the 

Apache and Navajo Tribes, most if not all of the haciendas were abandoned (Wildeman and Brock 2000). 

Subsequently, large herds of domestic livestock roamed 

freely across the grasslands of Arizona.

Anglo settlement and livestock grazing began in the 

early 1870s (Young 1998). Cattle driven from Texas were 

grazed the Salt River watershed and 100,000 head of 

cattle grazed the headwaters of the San Pedro River 

in southern Arizona (Wildeman and Brock 2000). By 

1891 about 1.5 million cattle grazed the grasslands 

and watersheds of southern Arizona (Cox and others 

1983, Ferguson and Ferguson 1983, Hendrickson and 

Minckley 1985).

By the early 1870s, even untrained observers noted 

that range conditions across the Southwest were rapidly 

deteriorating. Overstocking of upland and riparian 

grassland habitats and drought (late 1880s) was implicated to be the primary cause of decline (Leopold 1946, 

Hastings 1959, Gresswell and others 1989, Scurlock 

1998, Kruse and Jemison 2000). Arroyo cutting and 

erosion of topsoil began to occur as vegetation was 

stripped from watershed slopes. Overutilized riparian 

areas, which had formerly acted as filter strips in their 

pristine state, were unable to trap the large quantities of soil moving down slope into stream channels 

(Hastings 1959, Miller 1961, York and Dick-Peddie 

1969, Dick-Peddie 1993, Minckley 1969, Sublette and 

others 1990, Bahre 1991, 1995, Friedman and others 1997, Wildeman and Brock 2000). Southwestern 

grasslands east of the Rocky Mountains (Great Plains 

type grasslands) faired better than desert grasslands 

of western New Mexico and Arizona. Great Plains type 

grasslands evolved under the grazing pressure of large 

herds of native ungulates (Milchunas and others 1988, 

Loftin and others 2000).

In contrast, desert grasslands west of the Continental 

Divide evolved under light grazing pressure from native ungulates. Thus, making the latter susceptible to 

degradation by livestock grazing (Elmore 1992, Bock 

and Bock 1993, Loftin and others 2000). Grassland 

reaches of streams at intermediate and higher elevations in the American Southwest suffered far less 

impact than middle and lower elevation grassland 

stream reaches. Many of these stream reaches are 

found within National Forest System lands and Native 

American lands. Greater amounts of precipitation, 

relative to lower elevation grasslands, made these 

riparian zones more resistant to the effects of grazing 

and other land use practices. Steeper gradient reaches 

also are less accessible to grazing livestock and use 

by agriculture.

As one can see, the long history of grazing in the 

Southwest has had definite impacts on Southwestern 

watersheds. Today, many riparian areas in the 

Southwest are considered to be functionally impaired 

due to environmental stressors such as overgrazing 

(Leopold 1974, Galliziolli 1977, Davis 1977, Hibbert 

1979, Platts 1983, Clarkson and Wilson 1995, Patten 

1998, Baker and others 2000). Grazing of livestock 

affects watershed condition and function through 

the direct removal of vegetation and the mechanical 

action of trampling (Johnston 1962, Johnson and 

others 1978, Tromble and others 1974, Hibbert 1979, 

Ryder 1980, Reid 1993, Peiper 1994, Belsky and others 

1999). Removal of riparian vegetation and trampling 

through grazing can reduce bank stability and the 

channel-narrowing capacity of a stream (Platts 1979, 

1982, Platts and others 1983, Abdel-Magid and others 1987a,b, Medina 1995, Neary and Medina 1995, 

Clary and Kinney 2000). Hoof action disturbs stream 

substrate, loosens and/or compacts soils, which in 

turn reduces infiltration, increases surface runoff, 

and accelerates movement of soils into watersheds 

(Wilcox and Wood 1988, Neary and Medina 1996, 

Payne and Lapointe 1997, Loftin and others 2000). 

Typical changes in stream morphology and aquatic 

habitat are streambank sloughing, widening and 

shallowing of the streambed, channel entrenchment, 

velocity reduction (Behnke and Raleigh 1978, Platts 

1981a,b, Allen-Diaz and others 1998). The result is a 

less complex aquatic habitat. Fine sediments (less than 

2 mm) that would typically be suspended in the water 

column and transported downstream are deposited on 
156 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

the streambed as water velocity decreases. Erosion 

of bank materials, due to loss of ground cover and 

increased surface runoff, adds to the fine sediment 

portion of the stream bedload (Dune and Leopold 1978, 

Pogacnik and Marlow 1983, Williams and Sada 1985, 

Abdel-Magid and others 1987b, Schultz and Leininger 

1990, Platts 1991, Medina 1995, Neary and Medina 

1996). Fines in bedload fill interstitial spaces and 

sedimentation of channel substrate begins (Lisle 1989, 

Bevenger and King 1995). Increases in temperature 

can occur due to alteration of species composition and 

reduction in plant density through direct removal of 

vegetation by grazing livestock (Amaranthus and others 1989, Schultz and Leininger 1990). Overgrazing of 

livestock of watersheds in arid and semiarid regions 

such as the Southwestern grasslands may have more 

severe consequences to watersheds, riparian habitats, 

and their inhabitants than in more mesic habitats 

where plant vigor is assumed to be greater (Behnke 

and Raleigh 1978).

Physical and chemical changes caused by grazing 

in grassland watersheds, riparian areas, and instream 

habitats clearly may have a direct affect on the quality 

and quantity of fish habitat (Meehan and Platts 1978, 

Platts and Nelson 1985, Platts 1981a,b, 1991, Bestchta 

and Platts 1986). Although the linkage between fish 

and grazing is not well defined and many cases cannot 

be correlated (Rinne 1988a b, 1999a,b, 2000), evidence 

does suggest that as grazing negatively impacts fish 

habitat, it also affects fish populations (Stuber 1985). 

For example, Platts (1981a) reported that a stream 

reach receiving high intensity grazing from sheep 

had a channel width five times greater, stream depth 

one-fifth as deep, and stream shore depth one-third as 

deep in heavily grazed reaches compared to moderately 

or lightly grazed adjoining reaches. Livestock grazing 

in riparian areas can reduce overhanging banks that 

can serve as cover for fish, and cause stream banks to 

become reposed with a decrease in stream shore depth 

(Overton and others 1994). Platts (1991) concluded, 

as did Behnke (1977), that degradation of riparian 

habitats through grazing was one of the major vectors 

causing the decline of native fishes in the Western 

United States. Grazing causes temporal and spatial 

changes in the riparian canopy and composition, 

which in turn may elevate stream temperature, and 

thus may have negative effects on the persistence or 

productivity of some fish populations, particularly 

cold water species (Meehan and others 1977, Meehan 

1996). For example, Platts and Nelson (1989a,b) found 

that salmonid biomass was correlated with riparian 

canopy density in Great Basin streams. The authors 

suggested that solar insolation was a limiting factor 

in these waters. Sponholtz and Rinne (1997) reported 

that on reaches that were grazed by livestock in the 

Verde River, Arizona streambank vegetation was 60 

percent lower, channel width 50 percent greater, and 

stream velocity was 85 percent lower than in grazed 

reaches. Fish density in the grazed reaches was 50 

percent lower than at the ungrazed sites.

Although fine sediment (less than 2 mm) is a natural 

component of bedload (Hynes 1972), an overabundance 

of fines can affect the quality and complexity of aquatic 

environments and its inhabitants (Downes and others 

1998). Abiotic changes (that is, sedimentation) can 

decrease the diversity and abundance of macroinvertebrates that serve as food for fishes (Brown and 

Moyle 1991). Allen-Diaz and Jackson (1998) reported 

that insect family richness was significantly reduced 

on moderately grazed plot versus lightly or ungrazed 

plots. Rinne (1988a) reported similar findings in the Rio 

De Las Vacas, Santa Fe National Forest, New Mexico. 

Fine sediment can also impair reproductive success of 

gravel spawning fish. Meehan (1991) indicated that 

excessive percentage of fines negatively impacted the 

reproduction and recruitment in salmonid populations. 

In contrast, Rinne (1999b) could not identify if fines 

had negative impacts on spikedace and loach minnow and could not say it was a major limiting factor. 

Abundant populations of these two species are found 

in the Gila-Cliff Valley, New Mexico, where grazing 

is present (Propst and others 1986). Further, Rinne 

(2000) cautions that often times we are dealing with 

difficult-to-determine thresholds and discrete percentages of fines that negatively affect various stages of a 

fish life cycle. Too, streams with a greater percentage 

of suitable spawning gravels may be able to tolerate 

higher levels of sedimentation and still recruit fish 

into the population.

There is a scarcity of research on grazing effects on 

Southwestern fishes, especially nonsalmonids (Rinne 

2000), and it some cases, when dealing with cypriniform 

species, grazing may not have an impact on this group 

of fishes. Although empirical information is scarce on 

the effects of grazing in arid land watersheds (Rinne 

1999a, 2000), much research has been conducted in 

the Intermountain, Northern Rocky Mountain, and 

Pacific Northwest (Platts 1991, Rinne 1999a, 2000). 

This information can be used to lend insight into grazing 

effects in the Southwest. Platts (1991) and Rinne (1998, 

1999a, 2000) both state that more scientific rigor and 

better study design were needed in past studies and 

will be needed in future research. Nonetheless, negative effects of grazing on watershed condition, riparian 

vegetation, and aquatic habitats alluded to above have 

been documented by many practitioners throughout the 

Western United States (Ames 1977, Platts 1979, 1982, 

1991, Szarro 1989, Chaney and others 1990, Naimen 

and Descamps 1990, Elmore 1992, Belsky and others 

1999). In a review article published in the Journal 

of Soil and Water Conservation, Belsky and others 

(1999) concluded that most scientific studies document 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 157

that livestock grazing continues to be detrimental to 

stream and riparian ecosystems in the West. It may 

well be that the linkages between grazing and its direct 

effect on native fishes are still obscure (Medina and 

Rinne 1999, Rinne 1999a,b, 2000). Nevertheless, both 

watershed health and the protection of native fishes 

should be our primary objectives. Managers and land 

owners should assess whether these two ecosystem 

components will be negatively affected before deciding 

to graze livestock in Southwestern watersheds.

Management and Conservation_____

The first step in management of fish is the recognition that a species is imperiled locally or nationally. 

Managers must have up-to-date lists of species so that 

appropriate action can be taken to implement conservation measures (Miller 1972, Pister 1976, 1981). Since 

the late 1960s States across the country along with the 

Federal government has been enacting legislation and 

enforcing protection of endangered species, subspecies, 

and critical habitat. Laws such as the Endangered 

Species Preservation Act of 1966, Endangered Species 

Conservation Act of 1969, and the Endangered Species 

Act of 1973 (amended in 1978 and 1979) were put in 

place with the purpose of conserving endangered and 

threatened species and their ecosystems (Johnson and 

Rinne 1982) as well as maintaining natural habitats 

and creating refugia for some critically threatened 

fishes (Miller 1972, Miller and Pister 1991). An example 

of this is the refugia created for the Owens Valley 

pupfish. In 1971, Congress established the Desert 

Pupfish National Monument to save threatened pupfishes living east of Death Valley (Deacon and Bunnell 

1970). Similar actions could be put into play for fishes 

inhabiting Southwestern grasslands. Optimal habitat 

is infinitely more effective than imposition of restrictive 

regulations in the enhancement of fish populations 

especially with officially designated Threatened and 

Endangered species (Deacon and others 1979). Critical 

habitat has been designated for spikedace and loach 

minnow in the Southwest, and restocking programs 

for native fish are ongoing. Gila topminnow is being 

repatriated throughout its range in New Mexico and 

Arizona. Rio Grande sucker and chub are being planted 

into native waters in the upper Rio Grande basin, as 

are Rio Grande silvery minnow, razorback sucker, 

and Colorado pike minnow. These restockings have 

had mixed results due to the availability of optimum 

habitats or the continued competition with nonnative fishes. Nevertheless, these must be continued if 

we are to have any chance at successfully restoring 

native fishes.

Problems are inherent when trying to save native 

fish. First, because many native Southwestern fishes 

are nongame species considered by many, even in 

resource agencies, as “rough fish,” garnering support 

from private land owners or individuals and companies 

leasing public lands can be a challenge. Second, funding conservation efforts for native fish is a challenge 

because of budgetary constraints, political climate 

of the time, and, again, the view of these fishes as 

somewhat less desirable than game fish. Third, the 

frequent lack of information and communication 

among scientists might even be encouraged by what 

is perceived as small “cliques” of researchers whose 

intention is development of their own view on how 

the species should be managed. Fourth, water is an 

ever-decreasing resource in the Southwest, and many 

water laws permit beneficial water use without regard 

to fish and wildlife.

Finally, what is especially grim are the logistics 

and realities of trying to preserve native fishes in 

Southwestern grasslands, especially fishes that inhabit 

the grassland reaches. The grassland reaches of most 

Southwestern streams occur at low elevations in broad 

river valleys that have been occupied by humans for a 

long time and that typically have the highest human 

concentrations. Major cities such as Albuquerque, Las 

Cruces, El Paso, Tucson, and Phoenix have significant 

impacts on the water supply and quality of these low 

elevation grassland reaches. Major water withdrawals 

occur in these reaches; development and alteration 

of stream morphology is extensive along with the 

removal of riparian habitats. Pollution—both chemical and biological (encroachment of nonnative aquatic 

species)—is an ever-increasing concern. There are 

probably hundreds of thousands of nonnative fishes 

in these large-order, low-elevation grassland reaches. 

The chance of removing nonnative fishes to ensure 

the survival of native forms is small at best and, 

being realistic, probably impossible. The best land 

managers, fishery professionals, and society at large 

could expect is to identify isolated reaches on smaller 

tributary streams where logistics and isolation might 

result in some chance of success. Recovery is viable in 

small, remote locations, but at a large scale loss will 

continue and native fish community diversity will 

decline (Johnson and Rinne 1982).

This is a sad note, but one must face realities when 

prioritizing conservation projects when funding and 

worker availability are often limited. With that in 

mind, the rest of this chapter looks at a number of 

native fish in detail.

Loach Minnow (Rhinichthys (tiaroga) cobitis)

Distribution and Population Trends—Rhinichthys 

(tiaroga) cobitis is endemic to the upper Gila River 

drainage of southwestern New Mexico, southeastern 

and east-central Arizona, and northeastern Sonora 

(Miller and Winn 1951, Koster 1957, Minckley 1973, 

Propst 1999). Its historic distribution in New Mexico 
158 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

has been reported as the mainstem of the San Francisco 

River and Gila River along with many of the lower 

elevation tributaries (Koster 1957, Propst and others 

1988). Minckley (1973) indicated that loach minnow 

were historically found in the Gila River above Aqua 

Fria Creek, Salt River, Verde River, San Pedro River, 

San Francisco River, and Blue River. In Mexico, the 

species was known from the upper reaches of the San 

Pedro River in northern Sonora (Miller and Winn 1951, 

Propst 1999). Today the range of T. cobitis has been 

much reduced and the population trend is declining. 

In Arizona, loach minnow remain in Aravaipa Creek, 

Eagle Creek, White River, Blue River, and the North 

Fork of East Black River; however, populations continue 

to decline over its range. Loach minnow was listed by 

New Mexico as a threatened species (19 NMAC 33.1) 

in 1975 and is Federally listed as a threatened species 

(USFWS 1986). In Arizona it is listed as a species of 

special concern (AZGFD 1996).

Preferred Habitat—Rhinnichthys (tiaroga) 

cobitis is found in moderate to large streams where 

gradients range from 1 to 3 or 4 percent (Rinne 1989, 

Rinne and Deason 2000). Loach minnow are a cryptic 

species associated with riffle habitats (Minckley 1973, 

Propst and others 1988, Propst and Bestgen 1991, 

Propst 1999) where they live in the interstitial spaces 

of large gravel and cobble substrate (LaBounty and 

Minckley 1972, Minckley 1973, Rinne 1989, Turner 

and Tafanelli 1983, Rinne and Deason 2000). Loach 

minnow have a reduced air bladder, streamlined body 

form, and large fins, which allow them to exploit high 

velocity, riffle habitats where it breeds and feeds on 

aquatic insects, principally fly and mayfly larvae and 

nymphs; however, stone, caddis and blackfly larvae 

are also part of loach minnow diet. The species moves 

in bursts of swimming from one substrate material 

to another much as do darters of the Eastern United 

States. Loach minnow are most closely associated with 

other native riffle dwelling species. It coinhabits similar 

aquatic space with desert suckers and speckled dace 

(Rinne 1992, Propst 1999).

Breeding Biology—In New Mexico, loach minnows reproduced during a 4 to 6 week period during 

late March and early June when waters temperatures 

reach 16 to 20 0C (Britt 1982, Propst and Bestgen 

1991, Propst 1999). In contrast to New Mexico, Vives 

and Minckley (1990) found loach minnow spawning in 

the autumn in Aravipa Creek, Arizona. Bestgen and 

Propst (1991) indicate that runoff volume, timing, and 

duration of flow influences the onset of spawning for 

loach minnow. Loach minnow lay adhesive eggs on 

the undersides of rocks in flowing water habitats (42.7 

cm/sec or less) and at depths of 6.1 to 21.3 cm (Britt 

1982, Propst and Bestgen 1991, Propst and others 

1988, Propst 1999). Propst and Bestgen (1991) observed 

that when water velocities slow (less than 5 cm/s), egg 

mortality increased. Clutch size ranges from nine to 

260 eggs per nest (Britt 1982, Propst and others 1988). 

Britt (1982) reported fecundity of females ranging from 

144 to 1,200 eggs. Eggs of loach minnow hatch in 4 to 

5 days at 21 0

C (Propst and others 1988). Males may 

provide care during the incubation period (Propst and 

others 1985). Larvae average 5.4 mm TL at hatching 

(Propst 1999). Larvae use low velocity habitats after 

hatching (Propst and others 1988). Maximum life span 

of loach minnow ranges from 2 to 4 years (Minckley 

1973, Britt 1982, Propst and others 1988).

Management Issues and Recommendations—

Two major factors appear to determine the persistence 

of loach minnow: presence of nonnative species and the 

degradation of historic loach minnow habitat. Propst 

(1999) reports that at least 15 nonnative fishes are in the 

New Mexico portion of the loach minnow range with a 

similar number occurring in its historic Arizona range. 

An eminent threat is from introduced catfishes (Family: 

Ictaluridae), principally flathead catfish (Pylodictis 

olivaris), channel catfish (Ictalurus punctatus), black 

bullhead (I. melas), yellow bullhead (I. natalis), and 

brown bullhead (I. nebulosus) (Propst and others 1988). 

Flathead catfish are highly piscivorous and become so 

at an early age. Channel catfish often feed in riffles, 

habitat used by loach minnow. Propst (1999) indicates 

that where channel and flathead catfish co-occur, 

loach minnow are rare or extirpated. In addition to 

these predacious species, nonnative centrarchids are 

established throughout the range of loach minnow 

(Minckley 1973, Sublette and others 1990).

Land use activities such as cattle grazing, timber 

harvest, and road building, all of which contribute fine 

sediments to river systems, have also been implicated 

in the decline of loach minnow. Fine sediments fill 

the interstitial spaces where loach minnow live, feed, 

and breed, thereby decreasing available habitat and 

impacting spawning success. Also, dewatering of 

streams for consumptive uses eliminates spawning 

areas and core habitat.

Management activities directed toward removal of 

nonnative fishes, decreasing fine sediment in watersheds, and maintaining free-flowing unaltered stream 

systems appear to be critical for the conservation of 

loach minnow, along with translocating populations 

into suitable habitats free of nonnative fishes (Prospt 

1999).

Rio Grande Silvery Minnow (Hybognathus 

amarus)

Distribution and Population Trends—Native to 

New Mexico, the species has been historically found in 

the Rio Grande drainage downstream from Velarde, 

NM, to the Gulf of Mexico, in the Rio Chama downstream from Abiquiu, and the Pecos River downstream 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 159

from Santa Rosa (absent in the reach from Red Bluff 

Reservoir to Sheffield) to its confluence with the Rio 

Grande in Texas. The Rio Grande silvery minnow also 

occurs in coastal drainages of Texas from the Brazos 

River west to the Rio Grande drainage of New Mexico 

(Sublette and others 1990, Bestgen and Platania 1991, 

Platania 1995, Propst 1999). Bestgen and Platania 

(1991) report Hybognathusamarus as formerly common 

in the Rio Grande from Cochiti to Socorro and in the 

Pecos River from Fort Summer to Carlsbad. Likewise 

it was common near the confluence of the Pecos and 

Rio Grande in Texas (Trevino-Robinson 1959, Edwards 

and Contreras-Balderas 1991). Rio Grande silvery 

minnow appear to have been extirpated in the Pecos 

River and the lower Rio Grande below Elephant Butte 

Reservoir, the former occurring in the 1970s and the 

latter by the 1950s (Propst and others 1987, Edwards 

and Contreras-Balderas 1991, Bestgen and Platania 

1991, Propst 1999). Currently, H. amarus occupies less 

than 10 percent of its historic range (Propst 1999). The 

species is found only in perennial sections of the Rio 

Grande and incidentally in irrigation systems in the 

Middle Rio Grande reach from Cochiti to the head of 

Elephant Butte Reservoir (Sublette and others 1990, 

Platania 1991, 1993). Rio Grande Silvery minnow is 

listed by the USFWS as endangered (USFWS 2003), 

and by the New Mexico Department of Game and 

Fish as State Endangered, Group II (19 NMAC 33.1) 

(Propst 1999).

Preferred Habitat—Rio Grande silvery minnow 

are schooling fish and use a variety of habitats (Sublette 

and others 1990). Dudley and Platania (1997, 2000) 

report Rio Grande silvery minnow using mainstream 

habitats with water depths ranging from 0.2 to 0.8 m, 

water velocities of 0 to 30 cm/sec, and over silt/sand 

substrate. Propst (1999) indicates that a seasonal shift 

to slower velocity habitats with debris cover occurs 

in winter. While individuals are sometimes found in 

irrigation canals, often young and some adults have 

apparently been entrained during irrigation activities 

(Lang and Altenbach 1994, Propst 1999). Rio Grande 

silvery minnow are herbivores feeding on diatoms, 

larval insects skins, and live and decaying plant material (Sublette and others 1990).

Breeding Biology—Spawning activity of Rio 

Grande silvery minnow occurs at the end of spring 

and the beginning of summer with water temperatures 

ranging from 20 to 24 0

C (Propst 1999). Platania and 

Altenbach (1998) report that the species is a pelagic 

spawner, and the eggs are semibuoyant and nonadhesive. Females are attended by multiple males with 

several encounters (Propst 1999). Females are reported 

as producing three to 18 clutches in a 12-hour period. 

Propst (1999) reports that the majority (more than 90 

percent) of spawning individuals were age-1 adults. 

Further, it appears that mortality can be as high as 98 

percent of breeding individuals (Propst 1999). Similar 

to other larval fishes, larvae of Rio Grande silvery 

minnow drift for several days and then migrate to low 

velocity habitats at the stream margins.

Management Issues and Recommendations—

Propst (1999) lists factors related to the decline of Rio 

Grande silvery minnow. He suggests that nonnative 

fishes, impoundments, declining water quality, reduction of flow, and range fragmentation due to mainstem 

and diversion dams have together contributed to the 

decline and potential extirpation of Rio Grande silvery 

minnow in the Rio Grande drainage. Bestgen and 

Platania (1991) and Cook and others (1992) provide 

evidence that a combination of poor water quality, 

altered flow regimes, and hybridization with plains 

minnow, Hybognathus placitus, accounted for its 

elimination in the Pecos River.

Conservation strategies that maintain minimum 

flow, natural hydrographs, and improved water quality are needed to sustain Rio Grande silvery minnow. 

Maintenance of river flows will also alleviate loss 

through predation when minnows are concentrated 

in shrinking habitats.

Rio Grande Sucker (Catostomus 

(Pantosteous) plebeius)

Distribution and Population Trends—The Rio 

Grande sucker is native to Colorado, New Mexico, and 

Mexico. Historically, the species was widely distributed 

in the Rio Grande Basin of Colorado and New Mexico 

(Koster 1957, Sublette and others 1990, Langlois and 

others 1994). In New Mexico, its current distribution 

is the Rio Grande above the 36th parallel and its 

tributaries north of the 33rd parallel (Propst 1987, 

Sublette and others1990, Platania 1991, Calamusso 

unpublished data). Rio Grande suckers are found in 

the Mimbres River, and introduced populations are 

established in the Rio Hondo (Pecos drainage), Gila 

River basin, and San Francisco drainage (Sublette and 

others 1990). Rio Grande sucker also inhabit six river 

basins draining four States of Mexico (Hendrickson and 

others 1980, Abarca and others 1995). The Rio Grande 

sucker appears to be declining across its range. At one 

Rio Grande silvery minnow.
160 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

time the species was found in only one drainage in 

Colorado (Hot Creek) (Swift-Miller and others 1999a,b). 

In New Mexico, the species appears to be declining 

across its range (Michael Hatch, personal communication). Calamusso and others (2002) documented 

the decline and replacement of Rio Grande sucker by 

white sucker, Catostomus commersoni, in Rio Grande 

tributary streams in northern New Mexico. Rio Grande 

sucker was designated as State endangered by the 

Colorado legislature in 1993 when it was found that 

only one population remained (Swift and others 1999) 

in the State. While more common in New Mexico, Rio 

Grande sucker are considered rare in the Rio Grande 

mainstem (Rinne and Platania 1995) and is declining 

in the tributaries of the Middle and Upper Rio Grande 

basin (Calamusso 1992, 1996, Calamusso and Rinne 

1996, 1999).

Breeding Biology—Temperature is the controlling factor as indicated by latitudinal and elevation 

differences in spawning time. Spawning begins in 

February in the southern portion of the species range 

and occurs progressively later northward (Smith 1966). 

Rio Grande sucker were found spawning in Animas 

Creek, Sierra County, NM, in February when water 

temperature reached 9 0

C (Calamusso unpublished 

data) and in June in streams on the Santa Fe National 

Forest when water temperature also reached 9 0

C 

(Calamusso 1996). Rinne (1995b) reported spawning in 

the Rio de las Vacas peaking in June during the declining spring flows, and Rausch (1963) found the species 

spawning in the Jemez River in May. In two streams 

on the Santa Fe National Forest, suckers spawned 

where dominant substrate ranged in size from 2.0 to 

18.9 mm, water velocity ranged from 21.0 to 62.0 cm/s 

(mean 37.0 cm/s, and depth ranged from 9.0 to 28.2 

cm (mean 16.2 cm) (Calamusso 1992). Koster (1957) 

indicated that Rio Grande sucker also spawned in the 

fall. Males and females become sexually mature at age 

3 (Rausch 1963). Maximum life span of females is 7 

years, with a corresponding mean standard length of 

159 mm (Rausch 1963).

Preferred Habitat—Rio Grande sucker live in 

small to large middle elevation (2,000 to 2,600 m) 

streams over substrates that range from sand/small 

gravel to cobble (Sublette and others 1990). Calamusso 

(1996) and Calamusso and Rinne (in review) evaluated 

habitat use of Rio Grande sucker in six study streams 

on the Carson and Santa Fe National Forests of 

northern New Mexico. Rio Grande suckers were found 

in low gradient (3.2 percent or less) stream reaches 

and were associated with pool and glide habitat, low 

water column velocity (less than 20 cm/s), moderate 

depth (mean: 24.5 cm; range 7 to 62 cm), and cobble 

substrate. These habitats were used disproportionately 

to available habitats. Abundance of Rio Grande sucker 

was inversely related with stream gradient and mean 

water column velocity. Density and biomass of Rio 

Grande sucker was positively related with increasing 

amounts of pool and glide habitat at the reach level, 

as was greater biomass.

Management Issues and Recommendations—

The Rio Grande sucker was the only catostomid 

endemic to the Rio Grande (Koster 1957, Sublette and 

others 1990, Langlois and others 1994). Rio Grande 

suckers probably once occupied all low gradient (less 

than 3.2 percent), middle-elevation tributaries to 

the Rio Grande in Colorado and New Mexico north 

of the 33rd parallel (Calamusso and others 2002). 

The decline in range and numbers of the Rio Grande 

sucker is related to introduction of nonnative fishes, 

especially white sucker, into the Rio Grande drainage (Calamusso 1996, Calamusso and Rinne 1999, 

Calamusso and others 2002). In all streams where Rio 

Grande sucker have been extirpated or are declining, 

white sucker are now present (Calamusso and others 2002). Although nonnative brown trout, Salmo 

trutta, are believed to pry upon Rio Grande suckers, 

Calamusso and Rinne (1999) could not detect a negative correlation between presence of the nonnative 

trout and the native sucker. Management directed 

at protecting remaining populations of Rio Grande 

sucker using fish migration barriers is warranted. In 

addition, Rio Grande sucker should be reintroduced 

along with Rio Grande cutthroat during restorations 

for the latter. Recovery efforts for both these species 

should consider the effect of nonnative species and 

stream gradient on their persistence and abundance 

when considering reintroduction sites.

Rio Grande Chub (Gila pandora)

Distribution and Population Trends—Rio 

Grande chub are native to the Rio Grande basin of 

Colorado, New Mexico, and Texas. In New Mexico the 

species is also found in the Pecos River and Canadian 

drainages, although it may have been introduced in 

the latter. While no recent systematic review of its 

status in New Mexico has been conducted, the species appears to be stable and widespread throughout 

the montane tributaries of the Middle and Upper Rio 

Grande basin. In contrast, Rio Grande chub is considered rare or may even be extirpated in the Rio Grande 

mainstem (Rinne and Platania 1995, Propst 1999) 

and in tributaries in the southern portion of the State 

below Albuquerque. Recent ichthyofaunal surveys of 

streams draining the east slopes of the Black range, 

Sierra County, NM, indicate that Rio Grande chub 

are being replaced by nonnative longfin dace, Agosia 

chrysogaster, (Calamusso unpublished data). Propst 

(1999) suggests that legal protection for the species 

in New Mexico may be warranted. In Colorado Rio 

Grande chub is considered to be in general decline and 

listed as threatened (Zuckerman 1983, Zuckerman and 

Langlois 1990, Sublette and others 1990).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 161

Breeding Biology—Rio Grande chub spawn in 

spring and early summer (Sublette and others 1990). 

Rinne (1995a) found that spawning of Rio Grande chub 

reached its peak on the descending limb of the spring 

hydrograph in the Rio de las Vacas, New Mexico. Rio 

Grande chub are known to hybridize with longnose 

dace, Rhinichthyes cataractae, when the two occur 

in reduced habitats due to dewatering or drought 

conditions (Cross and Minckley; 1960, Suttkus and 

Cashner 1981).

Preferred Habitat—Similar to many members 

of the Genus Gila, Rio Grande chub are found in 

low velocity habitats (pools, glides, runs) of small to 

moderate sized streams where gradients rarely exceed 

2 percent and are often associated with instream 

woody debris or aquatic vegetation (Woodling 1985, 

Calamusso 1996). Rio Grande chub are midwater 

carnivores that, similar to many salmonids, feed on 

insect drift, zooplankton, and small fishes (Sublette 

and others 1990). Gila pandora also feed on detritus 

(Sublette and others 1990).

Management Issues and Recommendations—

Declines in Rio Grande chub populations appear to be 

due in large part to the introduction and range expansion of nonnative fishes more than overall degradation 

of habitat quality (Zuckerman 1983, Zuckerman and 

Langlois 1990, Propst 1999, Calamusso unpublished 

data). The latter may decrease population persistence 

through decreased habitat quality and decreased 

spawning success. Future conservation strategies 

should consider protection of tributary Rio Grande 

chub populations from intrusion by nonnative cyprinids, centrarchids, and salmonids, particularly brown 

trout. Many barriers currently protecting Rio Grande 

chub (and Rio Grande sucker) populations are unstable 

beaver dams, debris jams, and water diversions. As 

with Rio Grande sucker, Rio Grande chub should 

be introduced to suitable habitat during restoration 

activities for Rio Grande cutthroat trout. Restoration 

efforts should be conducted on a watershed scale to 

ensure stream connectivity and enhance opportunities 

for metapopulation development.

Spikedace (Meda fulgida)

Distribution and Population Trends—The 

spikedace is endemic to the Gila River Basin of New 

Mexico and Arizona (Koster 1957, Minckley 1973) and 

possibly northern Sonora (Propst 1999). Historically, 

it was common in the Gila River drainage of Arizona 

downstream to Tempe. It was also found in the Agua 

Fria, upper Verde, and Salt Rivers (Barber and 

Minckley 1966, Minckley and Deacon 1968, Minckley 

1973). In New Mexico, spikedace were common in 

the San Francisco River, Gila River, and the lower 

reaches of the three forks of the Gila River (Anderson 

1978, Propst and others 1986). Today, the range of the 

spikedace has been severely reduced (Propst 1999). 

In Arizona it now is found only in the upper Verde 

River, Aravaipa Creek, and Eagle Creek (Barber and 

Minckley 1966, Minckley 1973, Marsh and others 

1991). In New Mexico, spikedace have been extirpated 

in the San Francisco River (Anderson 1978, Propst and 

others 1986). Propst (1999) reported the distribution 

of spikedace in the Gila River mainstem as sporadic. 

The species is found in low numbers in the East Fork 

of the Gila River and regularly collected in the West 

Fork of the Gila River, although it appears to be in 

decline. Spikedace may be extirpated from the Middle 

Fork of the Gila River (Propst and others 1986, Propst 

1999). The species is common to abundant in the upper Gila River in the Gila–Cliff Valley (Propst 1986, 

Rinne 1999). Spikedace is listed by New Mexico and 

Arizona as threatened. It is Federally listed as Federal 

Threatened 31 July 1986.

Preferred Habitat—Spikedace is a small (50 to 

60 mm adult size) pelagic cyprinid species that uses 

moderately deep, low gradient riffles and runs with 

gradient ranging from 0.3 to 0.5 percent in permanently 

flowing streams (Sublette and others 1990, Rinne 

1991, 1999b). They are associated with sand and 

gravel substrates (Rinne 1991). Turner and Tafanelli 

(1983) indicated that velocities most often used by 

spikedace ranged from 40 to 82 cm/sec with preferred 

depths ranging from 15 to 18.0 cm. Larvae use low 

velocity habitats near stream margins with stream 

depth typically less than 32.0 cm. Rinne and Kroeger 

(1988) reported spikedace using velocities more than 

25 cm/sec with mean water depths of 27 cm.

Habitat use varies geographically and temporally for 

spikedace along with variation due to age of individuals (Propst and others 1986, Propst 1999). The species 

is insectivorous with its principal diet composed of 

mayflies, midges, and caddis flies along with taking 

other aquatic and terrestrial insects opportunistically 

(Barber and Minckley 1983).

Breeding Biology—The breeding season for spikedace occurs from February to March at elevations less 

than 1,000 m, and from April to June at elevations 

greater than 1,500 m (Anderson 1978, Propst and 

Bestgen 1991). At the onset of spawning, females move 

into riffle areas attended by several males. Spawning 

Spikedace. (Copyright John Rinne. Used with permission.)
162 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

takes place in riffles with moderate velocity (25 to 50 

cm/sec) over small cobble/gravel substrate (Barber and 

others 1970, Propst and Bestgen 1991). Eggs of the species are demurral and are broadcast over the substrate, 

dropping into the interstitial spaces (Anderson 1978) 

where they incubate for about 4 to 7 days (Snyder 1981). 

Fecundity is related to age and size. Eggs produced by 

female spikedace can range from 100 to 300 mature ova 

(Minckley 1973, Propst and others 1986). Propst and 

others (1986) reports one age 2 individual 55 mm SL 

female containing 319 mature ova, and age 1 females 

averaging 101 mature ova. Spikedace are 5 to 7 mm 

TL upon emergence and grow at a rate of 1 mm per 

day (Propst and others 1986).

Management Issues and Recommendations—

The decline in spikedace distribution and abundance is 

related to the alteration of aquatic and riparian habitats 

and the introduction of nonnative fishes (Propst 1999). 

Alteration of habitat for this species includes: ranching 

and farming practices that have removed instream 

flow; increased channelization; and increased siltation 

caused by grazing, road building, and recreational 

activities; construction of dams; and the introduction 

of nonnative fishes. The latter both preys upon and 

competes with spikedace for resources (Sublette and 

others 1990, Propst 1999). Management practices that 

address these vectors of decline are needed to ensure 

the persistence of spikedace. Currently, the USDI Fish 

and Wildlife Service has designated reaches in the 

Southwest as designated critical habitat.

References

Abarca, F.J., K.L. Young, I. Parra, R.H. Bettaso, and K. Cobble. 

1995. Yaqui River fishes relevant to the Madrean Province: 

U.S.—Mexico collaborations. In: Debano, L.F. and others (tech. 

coord.), Biodiversity and management of the Madrean Archipelago: 

The sky islands of Southwestern United States and Northwestern 

Mexico. 1994 Sept. 19-23, 1994; Tucson, AZ. Gen. Tech. Rep. 

RM-GTR-264. Fort Collins, CO: U.S. Department of Agriculture, 

Forest Service, Rocky Mountain Forest and Range Experiment 

Station. 669 p.

Abdel-Magid, A.H. Schuman, G.E., and R.H. Hart. 1987a. Soil bulk 

density and water infiltration as affected by grazing systems. 

Journal of Range Management 40(4):307-309.

Abdel-Magid, A.H. Schuman, G.E., and R.H. Hart. 1987b. Soil and 

vegetation responses to simulated trampling. Journal of Range 

Management 40(4): 307-309.

Allen-Diaz, B. H. and R. D. Jackson. 1998. Cattle grazing effects on 

oak woodland spring ecosystems. In: Proceedings of the American 

Water Resources Association Specialty Conference on Rangeland 

Management and Water Resources. May 27-29, 1998, Reno, NV. 

American Water Resources Association. 146 p.

Allen-Diaz, B.H., R.D. Jackson and J.S. Fehmi. 1998. Detecting 

channel morphology change in California’s hardwood rangeland 

spring ecosystems. Journal of Range Management 51:514-518.

Amaranthus, M.H., H. Jubas, and D. Arthur. 1989. Stream shading, 

summer streamflow and maximum water temperature following 

intense wildlfire in headwater streams. In: N.H. Berg (tech. coord.) 

Proceedings of the Symposium on fire and Watershed Management, 

Gen. Tech. Rep. PSW-109, Berkeley, California.

Ambos, N., G. Robertson, and J. Douglas. 2002. Dutchwoman Butte: 

A relict grassland in Central Arizona. Rangelands. 22 (2): 3-8.

Ames, C.R. 1977. Wildlife conflicts in riparian management: grazing. 

Pp. 49-51, In: B.R. Johnson, and D.A. Jones (tech. coords.). USDA 

Forest Service Gen. Tech. Rep. RM-43. Importance, preservation, 

and management of riparian habitat: a symposium. Denver, 

Colorado.

Anderson, 1978. The distribution and aspects of the life history of 

Media fulgida in New Mexico, M.S. Thesis, New Mexico State 

University, Las Cruces.

Arizona Game and Fish Department (AZGF). 1996. Wildlife of 

Special Concern in Arizona. Nongame and Endangered Wildlife 

Program, Arizona Game and Fish Department, Phoenix.

Bahre, C.J. 1991. A legacy of change: historic human impact on 

vegetation of the Arizona borderlands. University of Arizona 

Press, Tucson.

Bahre, C.J. 1995. Human impacts on the grasslands of southeastern 

Arizona. In: M.P. McClaran and T.R. Van Devender (eds). The 

desert grassland. University Arizona Press, Tucson.

Bailey, G. and R.G. Bailey. 1986. A history of the Navajos: the reservation years. School of American Research Press, Santa Fe, NM.

Baker, Malchus, B.B. Jr. and P.F. Ffolliott. 2000. Contributions of 

watershed management research to ecosystem-based management 

in the Colorado River Basin. Pp. 117-128. In: Ffolliott, P.F., M.B. 

Baker Jr., C.B. Carlton, M.C. Dillon, K.L. Mora (tech. coords.). 2000. 

Land stewardship in the 21st century: The contributions of watershed management; 2000 March 13-16; Tucson, AZ. Proceedings. 

RMRS-P-13. Fort Collins, CO: U.S. Department of Agriculture, 

Forest Service, Rocky Mountain Research Station, 438 p.

Barber, W.E. and W.L. Minckley. 1966. Fishes of Arivaipa Creek, 

Graham and Pinal Counties, Arizona. The Southwestern 

Naturalist. 11:313-314.

Barber, W.E. and W.L. Minckley. 1983. The feeding ecology of a 

southwestern cyprinid fish, the spikedace, Meda fulgida, in 

Arizona. Copeia. 1970:9-18.

Barber, W.E., D.C. Williams, and W.L. Minckley. 1970. Biology of the 

spikedace, Media fulgida, in Arizona. Copeia. 1970: 9 – 18.

Barnes, W.C. 1926. The Story of the Range. Washington, D.C.: GPO, 

1926. iv, 60 pp.

Baxster, J.O. 1987. Las Carneradas: Sheep trade in New Mexico, 

1700-1860. University of New Mexico Press, Albuquerque.

Bayer, L. and F. Montoya. 1994. Santa Ana: The people, pueblo, 

and the history of Tamaya. University of New Mexico Press, 

Albuquerque.

Behnke, R.J. 1977. Fish faunal changes associated with land use and 

water development. Great Plains – Rocky Mountain Geological 

Journal. 6(2):133-136. In: W.S. Platts 1981. Influence of Forest and 

rangeland management on anadromous fish habitat in western 

North America: Effects of livestock grazing. USDA Gen. Tech. 

Rep. PNW-124. 25 pp.

Behnke, R.J. and R.F. Raleigh. 1978. Grazing and the riparian 

zone: Impact and management perspectives. Pp. 262-267. In: 

Proceedings of the symposium, Strategies for protection and 

management of floodplain wetlands and other riparian ecosystems. 

11-13 December 1978. Callaway Gardens, GA. USDA Forest 

Service. Gen. Tech. Rep. WO-12.

Behnke, R.J. and M. Zarn. 1976. Biology and management of 

threatened and endangered western trout. Gen. Tech. Rep. RM-28. 

Fort Collins, CO: U.S. Department of Agriculture, Forest Service, 

Rocky Mountain Forest and Range Experiment Station.

Belsky, A.J., A.Matzke, and S.Uselman. 1999. Survey of Livestock 

Influences on Stream and Riparian Ecosystems in the Western 

US. Journal of Soil and Water Conservation. 54: 419-31.

Bestchta, R.L. and W.S. Platts. 1986. Morphological features of small 

streams: significance and function. Water Resources Research. 

22:369-379.

Bestgen, K.R. and S.P. Platania. 1991. Status and conservation 

of the Rio Grande silvery minnow, Hybognathus amarus. The 

Southwestern Naturalist. 36: 225-232.

Bevenger, G.S. and R.M. King. 1995. A pebble count procedure for 

assessing watershed cumulative affects. USDA Forest Service 

Research Paper RM-319:1-17.

Blinn, D.W., C. Runck, D.A. Clark, and J.N. Rinne. 1993. Effects 

of rainbow trout predation on the Little Colorado Rive spinedace. Transactions of the American Fisheries Society. 122(1): 

139-143.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 163

Bock, C.E. and J.H. Bock. 1993. Cover of perennial grasses in 

southeastern Arizona in relation to livestock grazing. Conservation 

Biology. 7:371-381.

Briggs, C.L. and J.R. Van Ness. 1987. Land, water, and culture: 

New perspectives on Hispanic land grants. University of New 

Mexico Press, Albuquerque.

Briggs, J.C. 1986. Introduction to the zoogeography of North 

American Fishes. Pages 1-6 In: C.H. Hocutt and E.O. Wiley 

(eds). The Zoogeography of North American Fishes. John Wiley 

and Sons, New York.

Britt, K.A. 1982. The reproductive biology and aspects of life history 

of Tiaroga cobitis in southwestern New Mexico. M.S. Thesis, New 

Mexico State University, Las Cruces.

Brown, L.R. and P.B. Moyle. 1991. Changes in habitat and microhabitat partitioning within an assemblage of stream fishes in 

response to predation by Sacramento squawfish (Ptychocheilus

grandis). Canadian Journal of Fisheries and Aquatic Sciences. 

48: 849-856.

Bryan, K. 1928. Date of channel trenching (arroyo cutting) in the 

arid southwest. Science. 62: 338-344.

Burr, B.M. and R.L. Mayden. 1992. Phylogenetics and North 

American freshwater fishes. Pages 287-324 In: R.L. Mayden, (ed) 

Systematics Historical Ecology, and North American Freshwater 

Fishes. Stanford University Press, Stanford, California.

Calamusso, B. 1992. Current distribution of the Catostomus plebeius

and Gila pandora on the Carson National Forest, New Mexico 

with preliminary comments on habitat preferences. Proceedings 

of the Desert Fishes Council. 24:63-64.

Calamusso, B. 1996. Distribution, abundance, and habitat of Rio 

Grande sucker (Catostomus plebeius) in the Carson and Santa Fe 

National Forests, New Mexico. M.S. Thesis, New Mexico State 

University, Las Cruces, NM.

Calamusso, B. and J.N. Rinne. 1996. Distribution of the Rio Grande 

cutthroat trout and its co-occurrence with the Rio Grande sucker 

and Rio Grande chub on the Carson and Santa Fe national forests. 

Pp. 157-167. In: D.W. Shaw and D.M. Finch (tech. coords.) Desired 

future conditions for Southwestern riparian ecosystems: Bringing 

interests and concerns together. Gen. Tech. Rep. RM-GTR-272. 

Fort Collins, CO: US Department of Agriculture, Forest Service 

Rocky Mountain Research Station.

Calamusso, B. and Rinne J.N. 1999. Native montane fishes of the 

Middle Rio Grande Ecosystem: Status, Threats, and Conservation. 

pp. 231-237 In: Finch, D.M. and others (Technical Coordinators), 

Rio Grande Ecosystems: Linking land, water, and people. Toward 

a sustainable future for the Middle Rio Grande Basin. 1998 June 

2-5; Albuquerque, NM. Proceedings RMRS-P-7. Ogden, Utah: 

U.S. Department of Agriculture, Forest Service, Rocky Mountain 

Research Station. 254 p.

Calamusso, B. and J.N. Rinne. In Review. Macro and Micro-Habitat 

use by Rio Grande sucker in the Carson and Santa Fe National 

Forests, New Mexico.

Calamusso, B., J.N. Rinne, and P.R. Turner. 2002. Distribution and 

abundance of Rio Grande sucker (Catostomus plebeius) in the 

Carson and Santa Fe National Forests, New Mexico. Southwestern 

Naturalist 47 (2): 182-186.

Carlson, A.W. 1990. The Spanish-American homeland: Four centuries in New Mexico’s Rio Arriba. John Hopkins University 

Press, Baltimore.

Carton, J.E., S.H. Stoleson, P. L Stoleson, and D.W. Shaw. 2000. 

Riparian Areas. In: Jemison, R. and C. Raish (eds) Livestock 

management in the American Southwest: Ecology, Society, and 

Economics. Elsevier, Amsterdam.

Cavender, T.M. 1986. Review of the fossil history of North American 

Fishes, in the Zoogeography of North American Freshwater Fishes 

In: C.H. Hocutt and E.O. Wiley (eds), John Wiley and Sons, New 

York, pp. 701-24.

Chaney, E., W. Elmore, and W.S. Platts. 1990. Livestock grazing on 

Western Riparian Ranges. 1990. U.S. Environmental Protection 

Agency, Northwest Resource Information Center, Eagle, Idaho.

Clarkson, R.W. and J.R. Wilson. 1995. Trout biomass and stream 

habitat relationships in the White Mountains area, east-central Arizona. Transactions of the American Fisheries Society 

124:599-612.

Clary, W.P. and J.W. Kinney. 2000. Streambank response to simulated grazing. Pp. 292 -295. In: P.F. Ffolliott, M.B. Baker Jr., C.B. 

Carlton, M.C. Dillon, K.L. Mora. (tech. coords.) Land stewardship 

in the 21st century: The contributions of watershed management; 

2000 March 13-16; Tucson, AZ. Proceedings. RMRS-P-13. Fort 

Collins, CO: U.S. Department of Agriculture, Forest Service, 

Rocky Mountain Research Station. 438 p.

Cook, J.A., K.R. Bestgen, D.L. Propst, and T.L. Yates. 1992. 

Allozymic divergence and systematics of the Rio Grande silvery 

minnow, Hybognathus amarus (Teleostei: Cyprinidae). Copeia 

1992:36-44.

Courtenay, W. Jr., D. Hensley, J. Taylor, and J. McCann. 1984. 

Distribution of exotic fishes in the Continental United States. Pp. 

41 – 77 In: W.R. Courtney Jr. and J. Stauffer (eds) Distribution, 

Biology, and Management of exotic fishes. John Hopkins University 

Press, Baltimore, Maryland.

Courtenay, W.R., Jr., and J.R. Stauffer, Jr. 1990. The introduced fish 

problem and the aquarium fish industry. Journal of the World 

Aquaculture Society 21(3):145-159.

Courtenay, W.R. and J.D. Williams. 1992. Dispersal of exotic species 

from aquaculture sources, with emphasis on freshwater fishes. 

Pp 49 – 81 In: A. Rosenfield and R. Mann (eds) Dispersal of living 

organisms into aquatic ecosystems. University of Maryland Sea 

Grant Publication, College Park.

Cox, J.R., H.L. Morton, J.T. LaBurme and K.G. Renard. 1983. 

Reviving Arizona’s rangelands. Journal of Soil and Water 

Conservation 38: 342-345.

Cross, J.N. 1967. Handbook of the fishes of Kansas. Misc. Pub. 

Museum Natural History, University of Kansas 45. 357 p.

Cross, F.B., R.L. Mayeden, and J.D. Stewart. 1986. Fishes inn the 

Western Mississippi Drainage. Pages 363-412 In: C.H. Hocutt 

and E.O. Wiley, eds. The Zoogeography of North American Fishes. 

John Wiley and Sons, New York.

Cross, F.B. and W.L. Minckley. 1960. Five natural hybrid combinations in minnows (Cyprinidae). University of Kansas Publication 

Museum of Natural History, 13: 1-18.

Davis, G.A. 1977. Management alternatives for the riparian habitat 

in the southwest. Pp. 59-67. In: Johnson, R.R., and D.A. Jones 

(tech. cords.) Importance of preservation and management of 

riparian habitat: a symposium. Gen. Tech. Rep. RM-43. USDA 

Forest Service, Fort Collins, CO.

Deacon, J.E. 1979. Endangered and threatened fishes of the West. 

Great Basin Naturalist Memoirs 3, 41-64.

Deacon, J.E. 1988. The endangered woundfin and water management in the Virgin River, Utah, Arizona, and Nevada. Fisheries 

13(1): 18-24.

Deacon, J.E. and S. Bunnell. 1970. Man and pupfish, a process of 

destruction. Cry California 5, 14-21.

Deacon, J.E., C. Hubbs, and B.J. Zahuranec. 1964. Some effects of 

introduced fishes on the native fish fauna of southern Nevada. 

Copiea 1964: 384-388.

Deacon, J.E., G.C. Kobetich, J.D. Williams, S. Contreras, and other 

members of the Endangered Species Committee of the American 

Fisheries Society. 1979. Fishes of North America Endangered, 

Threatened, or of Special Concern: 1979. Fisheries, Volume 4, 

No 2.

Deacon, J.E. and W.L. Minckley. 1974. Desert Fishes, Pp. 385-487 

In: Desert Biology, Volume 2, Academic Press, N.Y.

Dickinson, W.R. 1981. Plate tectonics and the continental margin of 

California, Pages 1–28 In: W.R. Ernst (ed) The Geotectonic development of California. Prentice-Hall, Inc., Englewood Cliffs, NJ.

Dick-Peddie, W.A. 1993. New Mexico Vegetation: past, present, and 

future. University of New Mexico Press, Albuquerque.

Dott, R.H. Jr. and R.L. Batten. 1981. Evolution of the earth, 3rd ed. 

McGraw-Hill, New York.

Douglas, M.E., P.C. Marsh, and W.L. Minckley. 1994. Indigenous 

fishes of western North America and the hypothesis of competitive displacement: Meda fulgida (Cyprinidae) as a case study. 

Copeia 1994: 9-19.

Dowling, T.E. and M.R. Childs. 1992. Impact of hybridization of a 

threatened trout of the southwestern United States. Conservation 

Biology 6(3): 355-364.

Downes, B.J., P.S. Lakes, E.S.G. Schreiber, and A. Glaister. 1998. 

Habitat structure and regulation of local species diversity in a 

stony, upland stream. Ecological Monographs 68(2), 237-257.

Dozier, E.P. 1983. The Pueblo Indians of North America. Waveland 

Press, New Haven.
164 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Dudley, R.K. and S.P. Platania. 1997. Habitat use of the Rio 

Grande Silvery Minnow. Report to the New Mexico Department 

of Game and fish, Santa Fe and the U.S. Bureau of Reclamation 

(Albuquerque Filed Projects Office) Albuquerque, NM. 188 pp.

Dudley, R.K. and S.P. Platania. 2000. Downstream transport of 

drifting semibouyant cyprinid eggs and larvae in the Pecos River, 

NM. Report to the U.S. Bureau of Reclamation. 61 pp.

Edwards, R.J., and S. Contreras-Balderas. 1991. Historical changes 

in the icthyofauna of the lower Rio Grande (Rio Bravo del Norte), 

Texas and Mexico. Southwestern Naturalist 36: 201-212.

Elmore, W. 1992. Pp. 442-457. In: Naiman, R.J. (ed). Watershed 

Management: Balancing Sustainability and Environmental 

Change. Springer- Verlag, New York.

Endler, J.A. 1977. Geographic Variation, Speciation, and clines. 

Princeton University Press, Princeton, NJ.

Etnier, D.A. and W.C. Starnes. 1993. The Fishes of Tennessee. 

University of Tennessee Press, Knoxville. 

Ferguson, D. and N. Ferguson. 1983. Sacred Cows At The Public 

Trough. Maverick Publishers Bend, Oregon.

Floyd, M.L., T.L. Fleischner, D. Hanna, and P. Whitefield. 2003. 

Effects of historic livestock grazing on vegetation at Chaco Culture 

National Park, New Mexico.

Friedman, J.M., M.L. Scott, and G.T. Auble. 1997. Water management and cottonwood forest dynamics along prairie streams. 

Pp. 49-71, In: F.L. Knopf and F.B. Samson (eds). Ecology and 

conservation of Great Plains vertebrates. Ecological Studies, 

Volume 125. Springer, New York.

Fuller, P.L., L.G. Nico, and J.D. Williams. 1999. Nonindigenous fishes 

introduced into inland waters of the United States. American 

Fisheries Society, Special Publication 27.

Gallizioli, S. 1977. Statement on improving fish and wildlife benefits 

in range management. Pp. 90-96. In: Townsend, J.F. and R.J. 

Smith (eds). Proceedings, improving fish and wildlife benefits 

in range management seminar. FWS/OBS/77-1. U.S. Fish and 

Wildlife Service.

Gilbert, C.R. 1976. Composition and derivation of the North American 

freshwater fish fauna. Florida Science, 39(2): 104-111.

Gresswell, R.E.B.A. Barton, and J.L. Kershner. (eds) 1989. Practical 

Approaches to Riparian Resource Management. U.S. Bureau of 

Land Management: Billings, MT.

Griffiths, D. 1904. Range investigations in Arizona. USDA. Bureau 

of plant industry. Bulletin 67.

Haines, G.B. and H.M. Tyus. 1990. Fish associations and environmental variables in age-0 Colorado squawfish habitats, Green 

River, Utah. Journal of Freshwater Ecology 5: 427-435.

Hanson, B. 1980. Fish survey of the streams in the Zuni River 

drainage, New Mexico. Report prepared for the Water and Power 

Resources Service, Boulder City, NV. U. S Fish and Wildlife 

Service, Albuquerque, NM.

Hastings, J.R. 1959. Vegetation change and arroyo cutting in 

Southeastern Arizona. Journal of the Arizona Academy of Science, 

(1) 60 – 67.

Hastings, J.R. and R.M. Turner. 1965. The Changing Mile. University 

of Arizona Press, Tucson, Arizona.

Heede, B. and J.N. Rinne. 1990. Hydrodynamic and fluvial morphologic processes: implications for fisheries management and 

research. North American Journal of Fisheries Management. 

10:249-268.

Heidinger, R.C. 1999. Stocking for sport fisheries enhancement. 

Pages 375-401 In: C.C. Kohler and W.A. Hubert (eds) Inland 

fisheries management in North America, American Fisheries 

Society, Bethesda, MD.

Hendrickson, D.A. and J.E. Brooks. 1991. Transplants of short-lived 

fishes of southwest North American deserts - a review, assessment 

and recommendations. Pp. 283-298 In: Battle Against Extinction 

- Desert Fish Management in the American Southwest.W.L. 

Minckley and J.E. Deacon (eds). University of Arizona Press, 

Tucson.

Hendrickson, D.A. and W.L. Minckley. 1985. Cienegas – Vanishing 

climax communities of the American Southwest. Desert Plants 

5:131-175.

Hendrickson, D.A., W.C. Minckley, R.R. Miller, D.J. Siebert, and 

P.H. Minckley. 1980. Fishes of the Rio Yaqui Basin, Mexico and 

the United States. Journal of Arizona-Nevada Academy of Science 

15(3): 65-106.

Hibbert, A.R. 1979. Managing vegetation to increase flow in the 

Colorado River Basin. USDA Forest Service, Gen. Tech. Rep. 

RM-66.

Hocutt, C.H., R.E. Jenkins, and J.R. Stauffer, Jr. 1986. Zoogeography 

of the fishes of the Central Appalachians and Central Atlantic 

Coastal Plain. Pages 161-211 In: C.H. Hocutt and E.O. Wiley 

(eds). The Zoogeography of North American Fishes. John Wiley 

and Sons, New York.

Hocutt, C.H. and E.O. Wiley. 1986. The Zoogeography of North 

American Freshwater Fishes. John Wiley and Sons, New York.

Hubbs, C.L. and R.R. Miller. 1948. The zoological evidence: correlation between fish distribution and hydrographic history in the 

desert basins of western United States. Bulletin of the University 

of Utah, 30: 17 – 66.

Hubbs, C.L., R.R. Miller, and L.C. Hubbs. 1974 Hydrographic history and relict fishes of the north-central Great Basin. California 

Academy of Science Memoirs 7: 1-259.

Johnson, J.E. and J.N. Rinne. 1982. The endangered species act 

and southwestern fishes. Fisheries. 7, 1-10.

Johnson, S.R., H.L. Gary, and Stanley L. Ponce. 1978. Range cattle 

impacts on stream water quality in the Colorado front range 

(U.S.A.). USDA, Forest Service, Rocky Mountain Forest and 

Range Experiment Station Research Note RM-359.

Johnston, A. 1962. Effects of grazing intensity and cover on 

the water-intake rate of fescue grassland. Journal of Range 

Management 15 (2).

Koster, W.J. 1957. Guide to the fishes of New Mexico. University of 

New Mexico Press, Albuquerque.

Kruse, W.H. and R. Jemison. 2000. Grazing systems of the southwest. 

In: Jemison, R. and C. Raish, (eds). Livestock management in the 

American Southwest: Ecology, Society, and Economics. Elsevier, 

Amsterdam.

LaBounty, J.F. and W.L. Minckley. 1972. Native fishes of the upper 

Gila River system, New Mexico. Symposium on rare and endangered wildlife of the Southwestern United States, September 

22-23 1972, Albuquerque. Santa Fe, NM: New Mexico Department 

of Game and Fish, pp. 134-146.

Lang, B.K. and C.S. Altenbach. 1994. Ichthyofauna of the Middle 

Rio Grande Conservancy District irrigation system: Cochiti Dam 

to Elephant Butte State Park, July-August 1993. Albuquerque 

Projects Office, U.S. Bureau of Reclamation, Albuquerque, NM.

Langlois, D., J. Alves, and J. Apker. 1994. Rio Grande sucker recovery 

plan. Colorado Division of Wildlife, Denver.

Lassuy, D.R. 1995. Introduced species as a factor in extinction and 

endangerment of native fish species. American Fisheries Society 

Symposium. 15:391-396.

Lee, D.S., C.R. Gilbert, C.H. Hocutt, R.E. Jenkins, D.E. McAllister, 

J.R. Stauffer. 1980. Atlas of North American freshwater fishes. 

North Carolina Museum of Natural History, Raleigh, NC.

Leoplod, A. 1946. Erosion as a menace to the social and economic 

future of the Southwest. Journal of Forestry. 44:627-633

Leopold, A.S. 1974. Ecosystem deterioration under multiple use. 

Proceedings of the Wild Trout Symposium, Pp. 96-98. U.S. 

Department of the Interior Fish and Wildlife Service and Trout 

Unlimited, Denver, Colorado.

Leopold, L.B., M.G. Wolman, and J.P. Miller. 1964. Fluvial processes 

in geomorphology. W.H. Freeman and Co., San Francisco.

Li, H.W. and P.B. Moyle. 1999. Management of Introduced Fishes. 

Pp. 345-374. In: C. Kohler and W. Hubert (eds). Inland Fisheries 

Management in North America. American Fisheries Society. 

Bethesda, MD.

Lisle, T.E. 1989. Sediment transport and resulting deposition in 

spawning gravels, north central California. Water Resources 

Research 25(6), 1303-1319.

Loftin, S.R., C.E. Bock, J.H. Bock, and S.L. Brentley. 2000. Desert 

Grasslands. pp. 53-96 In: R. Jemison and C. Raish (eds). Livestock 

management in the American Southwest: Ecology, Society, and 

Economics. Elsevier Science B.V., Amsterdam.

Maddux, H.R., L.A. Fitzpatrick, and W.R. Noonan. 1993. Colorado 

River endangered fishes critical habitat draft biological support 

document. U.S. Fish and Wildlife Service, Salt Lake City, UT 

225 pp.

Marsh, P.C., F.J. Arbaca, M.E. Douglas, and W.L. Minckley. 1989. 

Spikedace (Meda fulgida) and loach minnow (Tiaroga cobitis) 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 165

relative to shiner (Cyprinella lutrensis). Arizona Game and Fish 

Department, Phoenix.

Marsh, P.C., J.E. Brooks, D.A. Hendrickson, and W.L. Minckley. 

1991. Fishes of Eagle Creek, Arizona, with records for threatened 

spikedace and loach minnow (Cyprinidae). Journal of the Arizona 

Academy of Science. 23: 107 – 116.

Marsh, P.C. and D.R. Langhorst, 1988. Feeding and fate of wild 

larval razorback sucker. Environmental Biology of Fishes. 

21(1):59-67.

Matthews, W.J. 1998. Patterns in freshwater fish ecology. Chapman 

and Hall, New York.

Matthews, W.J. and F.P. Gelwick. 1990. Fishes of Crutcho Creek and 

the North Canadian River near Oklahoma City: urbanization, and 

temporal variability. Southwestern Naturalist. 35, 403-10.

May, R.M. 1988. How many species are there on Earth? Science. 

241: 1441-1449.

Mayden, R.L. 1987. Historical ecology and North American highland 

fishes: a research program in community ecology. In: W.J. Matthews 

and D.C. Heins (eds) Community and Evolutionary Ecology of 

North American Stream Fishes, University of Oklahoma Press, 

Norman, pp. 210-222.

Mayden, R.L. 1992a. An emerging revolution in comparative biology and the evolution of North American freshwater fishes, in 

Systematics, Historical Ecology and North American Fishes R.L. 

Mayden (ed) Stanford University Press.

Mayden, R.L. 1992b. Systematics, Historical Ecology and North 

American Fishes. R.L. Mayden (ed) Stanford University Press.

Medina, A.L. 1995. Native aquatic plants and ecological condition. 

USDA Forest Service, Rocky Mountain Forest and Research 

Station Gen. Tech. Rep. 272:329-335.

Medina, A.L. and J.N. Rinne. 1999. Ungulate/fishery interactions 

in southwestern riparian ecosystems: Pretensions and realities. 

Proceedings of the North America Wildlife and Natural Resources 

Conference 62:307-322.

Meehan, W.R. 1991. Influences of forest and rangeland management 

on salmonid fishes and their habitats. American Fisheries Society 

Special Publication 19., Bethesda, Maryland.

Meehan, W.R. 1996. Influence of riparian canopy on macroinvertebrate composition and food habits of juvenile salmonids in 

several Oregon streams. USDA Forest Service. Research Paper 

PNW-RP-496. 14 p.

Meehan, W.R. and W.S. Platts. 1978. Livestock grazing and the 

aquatic environment. Journal of Sol and Water Conservation 

33(6):274-278.

Meehan, W.R., F.J. Swanson, and J.R. Sedell. 1977. Influence of riparian vegetation on aquatic ecosystems with particular reference to 

salmonid fishes and their food supply. Pp. 137-145 In: Johnson, 

R.R. and D.A. Jones (tech. coord.). Importance, preservation, and 

management of riparian habitats: A symposium. 9 July 1977. 

Tucson, AZ. USDA Forest Service. Gen. Tech. Rep.

Meffe, G.K., D.A. Hendrickson, and D.L. Minckley. 1983. Factors 

resulting in the decline of the endangered Sonoran topminnow 

Poeciliopsis occidentalis (Atheriniformes: Poeciliidae) in the 

United States. Biological Conservation. 25: 135-159.

Mehrhoff, L.A. Jr. 1955. Vegetation change on a southern Arizona 

grassland – an analysis of causes. University of Arizona, Tucson. 

Masters Thesis.

Milchunas, D.G., O.E. Sala, and W.K. Lauenroth. 1988. A generalized 

model of the effects of grazing by large herbivores on grassland 

community structure. American Naturalist. 132:87-106.

Miller, R.R. 1946a. Correlation between fish distribution and 

Pleistocene hydrography in eastern California and southwestern 

Nevada, with a map of the Pleistocene waters. Journal of Geology. 

54: 43 – 53.

Miller, R.R. 1946b. The need for ichthyological studies of the major 

rivers of western North America. Science. 104, 517-519.

Miller, R.R. 1959. Origin and affinities of the freshwater fish fauna 

of western North America. Pages 187 – 222 In: C.L. Hubbs, 

ed. Zoogeography. Pub. 51 (1958). Washington, DC: American 

Association for the Advancement of Science.

Miller, R.R. 1961. Man and the changing fish fauna of the American 

Southwest. Papers of the Michigan Academy of Science, Arts, and 

Letters. 46, 365-404.

Miller, R.R. 1972. Threatened Freshwater Fishes of the United States. 

Transactions of the American Fisheries Society, No. 2. 14p.

Miller, R.R. and B. Chernoff. 1979. Status of populations of the 

endangered Chihuahua chub, Gila nigrescens. Proceedings of 

the Desert Fishes Council 11: 74-84.

Miller, R.R., C. Hubbs, and F.H. Miller. 1991. Ichthyological 

Exploration of the American West: The Hubbs-Miller Era, 1915-

1950.

Miller, R.R. and E.P. Pister. 1971. Management of the Owens pupfish, 

Cyprinodon radiosus, in Mono County, California. Transaction of 

the American Fisheries Society. 100(3): 502-507.

Miller, R.R., J.D. Williams, and J.E. Williams. 1989. Extinction 

in North American Fishes during the past century. Fisheries. 

14: 22-38.

Miller, R.R. and W.E. Winn. 1951. Additions to the known fish fauna 

of Mexico: three species and one subspecies from Sonora. Journal 

of the Washington Academy of Science. 41:83 – 84.

Minckley, W.L. 1969. Attempted re-establishment of the Gila topminnow within its former range. Copeia. 1969(1):193-194.

Minckley, W.L. 1973. Fishes of Arizona. Sims Printing Co., Phoenix, 

AZ.

Minckley, W. L. 1983. Status of the razorback sucker, Xyrauchen 

texanus, (Abott) in the lower Colorado, Southwestern Naturalist 

28: 165-187.

Minckley, W.L. and Carufel. 1967. The Little Colorado River 

spinedace, Lepidomeda vittata, in Arizona. The Southwestern 

Naturalist. 12(3):291-302.

Minckley, W.L. and J.E. Deacon. 1968. Southwestern fishes and the 

enigma of “Endangered Species.” Science. 159: 1424 – 1432.

Minckley, W.L. and J.E. Deacon. 1991. Battle against extinction: 

Native fish management in the American West. University of 

Arizona Press, Tucson.

Minckley, W.L. and M. Douglas. 1991. Discovery and extinction of 

western fishes: a blink of the eye in geologic time. Pages 717 In: 

W.L. Minckley and J.E. Deacon (eds) Battle against extinction: 

Native fish management in the American West. University of 

Arizona Press, Tucson.

Minckley, W.L., D.A. Hendrickson, and C.E. Bond. 1986. Geography 

of Western North American Freshwater Fishes: Description and 

relationships to intercontinental tectonism. Pages 519-613 In: 

C.H. Hocutt and E.O. Wiley, eds. The Zoogeography of North 

American Fishes. John Wiley and Sons, New York.

Minckley, W.L., J.N. Rinne, and J.E. Johnson. 1977. Status of the 

Gila topminnow and its co-occurrence with mosquitofish. USDA 

Forest Service Research paper RM-198, Fort Collins, CO.

Moyle, P.B. 1976. Inland fishes of California. University of California 

Press.

Moyle, P.B. and B. Herbold. 1987. Life-history patterns and community structure in stream fishes of Western North America. 

Comparison with Eastern North America and Europe, in: W.J. 

Matthews and D.C. Heins (eds) Community and Evolutionary 

Ecology of North American Stream Fishes, University of Oklahoma 

Press, Norman, pp 25-32.

Moyle, P.B., H.W. Li, and B.A. Barton. 1986. The Frankenstein effect: 

impact of introduced fishes on native fishes in North America. Pp. 

415426 In:R.H. Stroud (ed). Fish culture in fisheries management. 

American Fisheries Society, Bethesda, Md.

Moyle, P.B., and R. Nichols. 1973. Ecology of some native and introduced fishes of the Sierra Nevada foothills in central California. 

Copeia. 1973(3):478-490.

Moyle, P.B., and R. Nichols. 1974. Decline of the native fish fauna of 

the Sierra-Nevada foothills, central California. American Midland 

Naturalist. 92(1):72-83.

Mueller, G.A., and P.C. Marsh. 2002. Lost, a desert river and its native 

fishes: A historical perspective of the lower Colorado River. Fort 

Collins, CO: U.S. Geological Survey. Information and Technology 

Report USGS/BRD/ITR 2002-0010. 69 p.

Neary, D.G. and A.L. Medina. 1995. Geomorphic response of a 

montane riparian habitat to interactions of ungulate vegetation, 

and hydrology. USDA Forest Service, Rocky Mountain Forest and 

Research Station Gen. Tech. Rep. 272:143-147.

Neary, D.G. and A.L. Medina. 1996. Geomorphic response of a montane riaparian habitat to interactions of ungulates, vegetation, 

and hydrology. Pp. 143-147. In: Shaw, D.W. and D.M. Finch (tech. 

coord.) Desired future conditions for Southwestern ecosystems: 

Bringing interests and concerns together. USDA Forest Service. 

Gen. Tech. Rep. RM-GTR-272.
166 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

Neel, J.K. 1963. Impact of reservoirs. Pages 575 – 593. In: Frey, D.G. 

(ed) Limnology in North America. Wisconsin Press, Madison.

Nelson, J.S. 1994. Fishes of the World, 3rd ed. John Wiley and Sons, 

New York. Meeting of the American Society of Ichthyologists and 

Herpetologists, Austin, TX.

Nico, L.G. and P. Fuller. 1999. Spatial and temporal patterns of 

nonindigenous fishes introductions in the United States. Fisheries. 

24:16-27.

Orth, D.J. and R.J. White, 1993. Stream Habitat Management. In: 

C.C. Kohler and W.A. Hubert (eds) Inland Fisheries Management 

in North America Bethesda, Maryland, USA: American Fisheries 

Society.

Overton, K.C., G.L. Chandler, and J.A. Pisano. 1994. Northern/

Intermountain Region’s fish habitat inventory: Grazed, rested, 

and ungrazed reference stream reaches, Silver King Creek, CA. 

USDA Forest Service. Gen. Tech. Rep. INT-GTR-311. 27 p.

Patten, D.T. 1998. Riparian ecosystems of semi-arid North America: 

diversity and human impacts. Wetlands. 18:498-512.

Payne, B.A. and M.F. Lapointe. 1997. Channel morphology and lateral 

stability: effects on distribution of spawning and rearing habitat 

for Atlantic salmon in wandering cobble-bed river. Canadian 

Journal of Fisheries and Aquatic Science. 54:2627-2636.

Peiper, R. 1994. Ecological Implications of livestock grazing. Pp. 

177-211. In: Vavra, W.A. Laycok, and R.D. Peiper (eds). Ecological 

Implications of Livestock Herbivory in the West. Society for 

Range Management.

Peplow, Edward H. Jr. 1958. Livestock Industry in Arizona. In: 

History of Arizona, Volume III. Lewis Historical Publishing 

Company, New York.

Pister, E.P. 1974. Desert fishes and their habitats. Transactions of 

the American Fisheries Society. 102(3): 531-540.

Pister, E.P. 1976. The rationale for the management of non-game 

fish and wildlife. Fisheries. 1(1):11-14.

Pister, E.P. 1981. The conservation of desert fishes. Pp. 411- 444 

In: R.J. Naiman and D.L. Soltz (eds). Fishes in North American 

Deserts. John Wiley and Sons, New York.

Platania, S.P. 1991. Fishes of the Rio Chama and upper Rio Grande, 

New Mexico, with preliminary comments on their longitudinal 

distribution. Southwestern Naturalist 36: 186-193.

Platania, S.P. 1993. The fishes of the Rio Grande between Velarde and 

Elephant Butte Reservoir and their habitat associations. Report 

to the New Mexico Department of Game and Fish, Santa Fe and 

Bureau of Reclamation (Albuquerque Projects Office) Albuquerque 

Ichthyofaunal Studies Program, Museum of Southwestern Biology, 

University of New Mexico, Albuquerque.

Platania, S.P. 1995. Reproductive biology and early life-history of the 

Rio Grande silvery minnow, Hybognathus amarus. Albuquerque 

District, U.S. Army Corps of Engineers, Albuquerque, NM.

Platania, S.P. and C.S. Altenbach. 1998. Reproductive strategies 

and egg types of seven Rio Grande basin cyprinids. Copeia. 

1998: 559-569.

Platts, W.S. 1979. Livestock grazing and riparian stream ecosystems. P.39-45. In: Proc. Forum-Grazing and Riparian Stream 

Ecosystems. Trout Unlimited. Inc. In: Kauffman, J.B. & W.C. 

Krueger. 1984. Livestock impacts on riparian ecosystems and 

streamside management implications...a review. Journal of Range 

Management. 37:430-483.

Platts, W.S. 1981a. Effects of livestock grazing. Pp. 389-423 In: W.R. 

Meehan (ed). Influence of forest and rangeland management on 

anadromous fish habitat in western North America. Gen. Tech. 

Rep. PNW-124. USDA Forest Service, Pacific Northwest Forest 

and Range Experiment Station.

Platts, W.S. 1981b. Effects of sheep grazing on riparian-stream 

environment. USDA Forest Service. Research Note INT-307. 

6 p.

Platts, W.S. 1982. Sheep and cattle grazing strategies on riparian stream environments. P. 251-270. In: Wildlife-Livestock 

Relationships Symposium. Proc. 10. University of Idaho Forest, 

Wildlife & Range Exp. Sta. Moscow. In: Kauffman, J.B. & W.C. 

Krueger. 1984. Livestock impacts on riparian ecosystems and 

streamside management implications...a review. Journal of 

Range. Management. 37:430-483.

Platts, W.S. 1983. Vegetation requirements for fisheries habitats. 

Pp. 184-188. In: Proceedings of the Symposia on managing 

intermountain rangelands – improvement of range and wildlife 

habitats. 15-17 September 1981; Twin Falls, ID. 22-24 June 1982; 

Elko, NV. USDA Forest Service Gen. Tech. Rep. INT-57.

Platts, W.S. 1991. Livestock grazing. Pp. 389-423. In: Influences of 

forest and rangeland management on salmonid fishes and their 

habitats. American Fisheries Society Special Publication. 19.

Platts, W.R. and R.L. Nelson. 1985. Stream habitat and fisheries 

response to livestock grazing and instream improvement structures. Big Creek, Utah. Journal of Soil and Water Conservation, 

40(4), 374-379.

Platts, W.S., and R.L. Nelson. 1989a. Characteristics of riparian 

plant communities and streambanks with respect to grazing in 

northeastern Utah. Pp. 73-81. In: Gresswell, R.E., B.A. Barton, J.L. 

Kershner (eds). Riparian resource management: an educational 

workshop. USDI Bureau of Land Management. Billings, MT.

Platts, W.S., and R.L. Nelson. 1989b. Stream canopy and its relationship to salmonid biomass in the Intermountain West. North 

American Journal of Fisheries Management, 9: 446-457.

Platts, W.S., Nelson, R.L., Casy, O. Crispin, V. 1983. Riparian-stream 

habitat conditions on Tabor Creek, Nevada, under grazed and 

ungrazed conditions. In: Western association of fish and wildlife 

agencies annual conference proceedings, 63rd; Teton Village, WY: 

The Association; 162-174.

Pogacnik, T.M. and C.B. Marlow. 1983. The effects of time grazing 

on stream channel stability and suspended sediment loading. 

Proceedings of the annual meeting of the Society for Range 

Management. 14-16 February 1983. Albuquerque, NM.

Propst, D.L. 1999. Threatened and Endangered fishes of New Mexico. 

Technical Report No.1. New Mexico Department of Game and 

Fish, Santa Fe. 84 p.

Propst, D.L. and K.R. Bestgen. 1991. Habitat and biology of the loach 

minnow, Tiaroga cobitis, in New Mexico. Copeia. 1991: 29-39.

Propst, D.L., K.R. Bestgen, and C.W. Painter. 1986. Distribution, 

status, biology, and conservation of the spikedace (Meda fulgida) 

in New Mexico. Endangered Species Report No. 15, U.S. Fish and 

Wildlife Service, Albuquerque, New Mexico.

Propst, D.L., K.R. Bestgen, and C.W. Painter. 1988. Distribution, 

status, biology, and conservation of the loach minnow, Tiaroga 

cobitis Girard, in New Mexico. USFWS, Albuquerque, Endangered 

Species Report No. 17, pp.1-75.

Propst, D.L., G.L. Burton, and B.H. Pridgeon. 1987. Fishes of the 

Rio Grande between Elephant Butte and Caballo Reservoirs, New 

Mexico. The Southwestern Naturalist .32: 408-11.

Propst, D.L, P.C. Marsh, and W.L. Minckley. 1986. Arizona survey 

survey for spikedace and loach minnow (Tiaroga cobitis): Fort 

Apache and San Carlos Indian Reservations and Eagle Creek, 

1985. Report, U.S. Fish and Wildlife Service, Albuquerque, New 

Mexico 8 p.

Rausch, R.R. 1963. Age and growth of the Rio Grande mountainsucker, Pantosteus plebeius (Baird and Girard). M.S. Thesis, 

University of New Mexico, Albuquerque.

Reid, L.M. 1993. Research and cumulative watershed effects. Gen. 

Tech. Rep. PSW-GTR-141. Albany, CA: Pacific Southwest Research 

Station, Forest Service, U.S. Department of Agriculture. 118 p.

Reid, G.K. and R.D. Wood. 1976. Ecology of inland waters and 

estuaries. D. Van Nostrand, New York.

Rinne, J.N. 1988a. Effects of livestock grazing exclosure on aquatic 

macroinvertebrates in a montane stream, New Mexico. Great 

Basin Naturalist. 48(2), 146-153.

Rinne, J.N. 1988b. Grazing effects on stream habitat and fishes: 

research design considerations. North American Journal of 

Fisheries Management. 8: 240-247.

Rinne, J.N. 1989. Physical habitat use by loach minnow, Tiaroga 

cobitis (Pisces: Cyprinidae), in southwestern (USA) desert streams. 

Southwestern Naturalist. 34(1): 109-117.

Rinne, J.N. 1990. An approach to management and conservation 

of a declining regional fish fauna: Southwestern United States, 

pp55-60. In: Maruyama and others (eds). Wildlife Conservation: 

Present trends and perspectives for the 21st century. International 

Symposium on Wildlife, 5th International Congress Zoology, August 

21-25, 1990 Tskuba and Yokohama, Japan.

Rinne, J.N. 1991. Physical habitat use by spikedace, Meda fulgida

in southwestern desert streams with reference to probable habitat 

competition by red shiner, Notropis lutrensis (Pisces: Cyprinidae). 

Southwestern Naturalist 36(1): 7-13.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005 167

Rinne, J.N. 1992. Physical habitat utilization of fish in a Sonoran 

Desert Stream, Arizona, southwestern United States. Ecology of 

Freshwater Fishes 1: 35 – 41.

Rinne, J.N. 1995a. Reproductive biology of the Rio Grande chub, 

Gila pandora, (Cypriniformes) in a montane stream, New Mexico. 

The Southwestern Natualist 40(1):107-110.

Rinne, J.N. 1995b. Reproductive biology of the Rio Grande sucker, 

Catostomus plebeius, (Cypriniformes) in a montane stream, New 

Mexico. The Southwestern Naturalist 40(2):102-105.

Rinne, J.N. 1998. Grazing and fishes in the southwest: Confounding 

factors for research, 75-84. In: Potts, D.F. (ed). Proceedings 

American Water Resources Agency Specialty Conference: 

Rangeland Management and Water Resources American Water 

Resources Agency/Society for Range Management Specialty 

Conference on Rangeland and Water Resources. May 26-30, 

Reno, NV. Herndon, VA.

Rinne, J.N. 1999a. Fish and grazing relationships: The facts and 

some pleas. Fisheries 24(8): 12-21.

Rinne, J.N. 1999b. The status of spikedace, Meda fulgida, in the 

Verde River, 1999. Implications for research and management. 

Hydrology and Water Resources in the Southwest 29.

Rinne, J.N. 2000. Fish and grazing relationships in southwestern 

national forests. Pp. 329- 371 In: Jemison, R. and C. Raish (eds). 

Livestock management in the American Southwest: Ecology, 

Society, and Economics. Elsevier Science B.V., Amsterdam..

Rinne, J.N., P. Boucher, D. Miller, A. Telles, J. Montzingo, R. Pope, 

B. Deason, C. Gatton, and B. Merhage. 1999. Comparative fish 

community structure in two southwestern desert rivers In: S. 

Leon, P. Stine, and C. Springer (eds) Restoring native fish to the 

lower Colorado River: Interactions of native and non-native fishes: 

A symposium and Workshop.

Rinne J.N. and B.P. Deason. 2000. Habitat availability and utilization by two native, threatened fish species in two southwestern 

rivers. Hydrology and Water Resources in Arizona and the 

Southwest. 30: 43-52.

Rinne, J.N. and K. Kroger. 1988. Physical habitat used by spikedace, 

Meda fulgida, In: Aravaipa Creek, Arizona. Proceedings of the 68th

Annual Conference of the Western Association of Fish and Wildlife 

Agencies, Albuquerque.

Rinne, J.N. and W.L. Minckley. 1985. Patterns of variation and distribution in Apache trout (Salmo apache) relative to co-occurrence 

with introduced salmonids. Copeia 1985(2): 285-292.

Rinne J.N. and W.L. Minckley. 1991. Native Fishes of Arid Lands: 

A Dwindling Natural Resource of the Desert Southwest. USDA 

Forest Service, Gen. Tech. Rep. RM-206: 1-45. Rocky Mountain 

Forest and Range Experiment Station, Fort Collins, CO.

Rinne, J.N. and S.P. Platania. 1995. Fish Fauna. pp. 165 – 175 In:

Ecology, diversity, and sustainability of the Middle Rio Grande 

Basin. Gen. Tech. Rep. RM-GTR-268. Fort Collins, CO: U.S. 

Department of Agriculture, Forest Service, Rocky Mountain 

Forest and Range Experiment Station. 186 pp.

Rinne, J.N. and P.R. Turner. 1991. Reclamation and alteration as 

management techniques,and a review of methodology in stream 

renovation, Pp. 219-244 In: W.L. Minckley and J.E. Deacon 

(eds). Battle Against Extinction--Native Fish Management in 

the American West. University of Arizona Press.

Roberts, B.C. and R.G. White. 1992. Effects of angler wading on 

survival of tout eggs and pre-emergent fry. North American 

Journal of Fisheries Management 12. 450-459.

Robison, H.W. 1986. Zoogeographic implications of the Mississippi 

River Basin. Pp 267-285 In: C.H. Hocutt and E.O. Wiley, eds. 

The Zoogeography of North American Fishes. John Wiley and 

Sons, New York.

Robison, H.W. and J.K. Beedles. 1974. Fishes of the Strawberry River 

System in northcentral Arkansas. Proceedings of the Arkansas 

Academy of Science. 28: 65-70.

Robison, H.W. and T.M. Buchanan. 1988. Fishes of Arkansas. 

University of Arkansas Press, Fayetteville.

Ryder, R.A. 1980. Effects of grazing on bird habitats. Pp 51-65. 

In: Workshop proceedings, Management of western forests and 

grasslands for nongame birds. 11-14 February 1980. Salt Lake 

City, UT. USDA Forest Service. Gen. Tech. Rep. INT-86.

Schickendanz, J.G. 1980. History of grazing in the southwest. 

Pp 1- 9, In: Kirk C. McDaniel and Chris Allison (eds). Grazing 

management systems for southwest rangelands: A symposium. 

The Range Improvement Task Force, Albuquerque, NM.

Schmidt, R.E. 1986. Zoogeography of the Northern Appalachians. 

Pages 137-159 In: C.H. Hocutt and E.O. Wiley (eds). The 

Zoogeography of North American Fishes. John Wiley and Sons, 

New York.

Schoenherr, A.A. 1981. The role of competition in the replacement of 

native fishes by introduced species. Pp 173-203 In: R.J. Naiman 

and D.L. Soltz (eds). Fishes of North American Deserts. John 

Wiley and Sons, New York.

Schultz, T.T. and W.C. Leininger. 1990. Differences in riparian 

vegetation between grazed areas and exclosures. Journal of Range 

Management. 43(4), 295-299.

Scurlock, D. 1986. Settlements and Missions, 1606 – 1680. In: J.L. 

Williams (ed) New Mexico in maps, 2nd edition. University of 

New Mexico Press.

Scurlock, D. 1998. From Rio to the Sierra: an environmental history of the middle Rio Grande Basin. USDA Forest Service Gen. 

Tech. Rep. RMRS-GTR-5, Rocky Mountain Research Station, 

Fort Collins, Colorado

Simberloff, D.S. 1974. Equilibrium theory of island biogeography and 

ecology. Annual Review of Ecological Systematics. 5: 161 – 182.

Smith, G.R. 1966. Distribution and evolution of the North American 

catostomid fishes of the subgenus Pantosteus, genus Catostomus. 

Systematical Zoology. 20:282-297.

Smith, G.R. 1978. Biogeography of intermountain fishes. Great 

Basin Naturalist Memoirs. 2: 17 – 46.

Smith, G.R. 1981a. Effects of habitat size on species richness and 

adult body size of desert fishes. Pages 125 – 172 In: R.J. Naiman 

and D.L. Soltz (eds). Fishes in North American Deserts.

Smith, G.R. 1981b. Late Cenozoic freshwater fishes of North America. 

Annual Review of Ecological Systematics. 12:163-193. John Wiley 

and Sons, New York.

Smith, M.L. 1981. Late cenozoic fishes in the warm deserts of 

North America: A reinterpretation of desert adaptations. Pages 

11 – 38 In: R.J. Naiman and D.L. Soltz (eds). Fishes in North 

American Deserts.

Smith, M.L. and R.R. Miller. 1986. The evolution of the Rio Grande 

basin as inferred from its fish fauna. Pages 457- 486 In: C.H. 

Hocutt and E.O. Wiley (eds). The Zoogeography of North American 

Fishes. John Wiley and Sons, New York.

Snyder, D.E. 1981. Contributions to a guide to the cypriniform fish 

larvae of the upper Colorado River system in Colorado. Biological 

Science Series No. 3, U.S. Bureau of Land Management, Denver, 

Colorado.

Sponholtz, P.J. and J.N. Rinne. 1997. Refinement of aquatic 

macrohabitat definition in the Upper Verde River, Arizona. Pp. 

17-24 In: Hydrology and water resources in Arizona and the 

Southwest. Volume 28. Proceedings of the 1998 Meetings of the 

Arizona Section, American Water Resource Association and the 

Hydrology Section, Arizona-Nevada Academy of Sciences, April 

17, 1999, Flagstaff, Arizona.

Starnes, W.C. and D.A. Etnier. 1986. Drainage evolution and fish 

biogeography of the Tennessee and Cumberland Rivers Drainage 

Realm. Pp 325-361 In: C.H. Hocutt and E.O. Wiley, eds. The 

Zoogeography of North American Fishes. John Wiley and Sons, 

New York.

Steirer, F.S., Jr. 1992. Historical perspective on exotic species. Pp 1-4 

In:M.R. DeVoe, ed. Introductions and transfers of marine species. 

South Carolina Sea Grant Consortium. Charleston.

Stewart, G. 1936. History of range use. Pp. 119-133 In: U.S. Forest 

Service: The Western Range. 74th Congress, 2nd session, Senate 

Document 199.

Stuber, R.J. 1985. Trout habitat, abundance, and fishing opportunities in fenced vs. unfenced riparian habitat along Sheep Creek, 

Colorado. Pp. 310-314. In: Johnson, RR, C.D. Ziebell, D.R. Patton, 

P.F. Ffolliott, and R.H. Hamre (tech. Coords.). Riparian ecosystems and their management: Reconciling conflicting uses. 16-18 

April 1985. Tucson, AZ. USDA Forest Service. General Technical 

Report RM-120.

Sublette, J.E., M.D. Hatch, and M. Sublette. 1990. The Fishes of 

New Mexico. University of Arizona Press, Tucson.

Suttkus, R.D. and R.C. Cashner. 1981. The intergeneric hybrid 

combination, Gila pandora x Rhinichthys cataractae (Cyprinidae), 
168 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-135-vol. 2. 2005

and comparisons with parental species. Southwestern Naturalist. 

26(1): 78-82.

Swift-Miller, S.M., B.M. Johnson, and R.T. Muth. 1999a. Factors 

affecting the diet and abundance of northern populations of 

the Rio Grande sucker (Catostomus plebeius). Southwestern 

Naturalist. 44:148-156.

Swift-Miller, S.M., B.M. Johnson, R.T. Muth, and D. Langlois. 1999b. 

Distribution, abundance, and habitat use of Rio Grande sucker 

(Catostomus plebeius) in Hot Creek, Colorado. Southwestern 

Naturalist. 44:42-48.

Szarro, R.C. 1989. Riparian forest and scrubland community types 

of Arizona and New Mexico. Desert Plants. 9(3-4):72-138.

Taylor, J.N., W.R. Courtenay, Jr., and J.A. McCann. 1984. Known 

impact of exotic fishes in the continental United States. Pp. 322-

373 In: W.R. Courtenay, Jr. and J.R. Stauffer, eds. Distribution, 

biology, and management of exotic fish. Johns Hopkins University 

Press, Baltimore, Maryland.

Trevino-Robinson, D.T. 1959. The icthyofauna of the lower Rio 

Grande, Texas and Mexico. Copeia. 1959: 253-256

Tromble, J.M., K.G. Renard, and A.P. Thatcher. 1974. Infiltration 

for three rangeland soil-vegetation complexes. Journal of Range 

Management. 27(4):318-321.

Turner, P.R. and R.J. Tafanelli. 1983. Evaluation of the instream 

flow requirements of the native fishes of Aravipa Creek, Arizona 

by the incremental methodology. U.S. Fish and Wildlife Service, 

Albuquerque, 114 pp.

Tyus, H. M. and J.F. Saunders. 2000. Nonnative fish control and 

endangered fish recovery. Fisheries 25: 17-24.

Tyus, H.M., B.D. Burdick, R.A. Valdez, C.M. Haynes, T.A. Lytle, 

and C.R. Berry. 1982. Fishes of the Upper Colorado River Basin: 

Distribution, abundance, and status. Pp 12- 70 In: W.H. Miller, 

H.M. Tyus, and C.A. Carlson (eds) Fishes of the Upper Colorado 

River system: Present and Future. Western Division of the 

American Fisheries Society, Bethesda, MD.

United States Congress, Office of Technology Assessment. 1993. 

Harmful non-indigenous species in the United States. U.S. 

Government Printing Office, OTA-F-565, Washington, DC.

United States Fish and Wildlife Service (USFWS). 1986. Endangered 

and threatened wildlife and plants; Determination of threatened 

status for the loach minnow. Federal Register 51: 39468-39478.

United States Fish and Wildlife Service (USFWS). 2003. Threatened 

and Endangered Species System.

Vives, S.P. and W.L. Minckley. 1990. Autumn spawning and others 

reproductive notes on loach minnow, a threatened cyprinid fish 

of the American Southwest. The Southwestern Naturalist. 35: 

451-454.

Weedman, D.A., A.L. Girmendonk, and K.L. Young. 1997a. Gila 

topminnow sites in Arizona: Provisional extirpation report 1996-97 

field season. Nongame and Endangered Wildlife Program Technical 

Report 116. Arizona Game and Fish Department, Phoenix.

Weedmand, D.A. and K.L. Young. 1997b. Status of the Gila topminnow and desert pupfish in Arizona. Nongame and Endanagered 

Wildlife Program Technical Report 118. Arizona Game and Fish 

Department, Phoenix.

Westphall, V. 1965. The Public Domain in New Mexico: 1854-1891. 

Albuquerque: University of New Mexico Press, 1965.

Wilcox, B.P. and M.K. Wood. 1988. Hydrologic impacts of sheep 

grazing on steep slopes in semiarid rangelands. Journal of Range 

Management. 41:303-306.

Wildeman, G. and J. H Brock. 2000. Grazing in the southwest: history 

of land use and grazing since 1540. In: Jemison, R. and C. Raish 

(eds). Livestock management in the American Southwest: Ecology, 

Society, and Economics. Elsevier Science B.V., Amsterdam.

Williams, J.E., D.B Brooks, J.E. Echelle, A.A. Echelle, R.J. Edwards, 

D.A. Hendrickson, and J.J. Landye. 1985. Endangered aquatic 

ecosystems in North American deserts, with a list of vanishing 

fishes of the region. Journal of the Arizona-Nevada Academy of 

Sciences. 20: 1-62.

Williams, J.E., J.E. Johnson, D.A. Hendrickson, Salvador ContrerasBalderas, J.D. Williams, M. Navarro-Mendoza, D.E. McAllister, 

and J.E. Deacon. 1989. Fishes of North America, endangered, 

threatened, or of special concern: 1989. Fisheries. 14, 2-20.

Williams, J.E. and D.W. Sada. 1985. Endangered species technical bulletin, Volume X, No. 11, United States Fish and Wildlife 

Service.

Williams, J.E., D.W. Sada, C.D. Williams, and Other Members of the 

Western Division Endangered Species Committee. 1988. American 

Fisheries Society guidelines for the introductions of threatened 

and endangered fishes. Fisheries. 13: 5-11.

Woodling, J. 1985. Colorado’s little fishes. A guide to the minnows 

and other lesser known fishes in the state of Colorado. Colorado 

Division of Wildlife, Department of Natural Resources, Denver.

York, J.C. and W.A. Dick-Peddie. 1969. Vegetation change in New 

Mexico during the past 100 years. In: Arid lands in perspective. 

University of Arizona Press, Tucson.

Young, W.D. 1998. The history of cattle grazing in Arizona. Pp. 

3 – 17 In: Hydrology and water resources in Arizona and the 

Southwest. Volume 28. Proceedings of the 1998 Meetings of the 

Arizona Section, American Water Resource Association and the 

Hydrology Section, Arizona-Nevada Academy of Sciences, April 

18, 1998, Glendale, Arizona.

Zuckerman, L.D. 1983. Rio Grande fishes management: progress 

report, November 1982 to June 1983. Colorado State University, 

Fort Collins. 53 pp.

Zuckerman, L.D. and D. Langlois. 1990. Status of Rio Grande sucker 

and Rio Grande chub in Colorado. Research Report, Colorado 

Division on Wildlife, Montrose.
You may order additional copies of this publication by sending your 

mailing information in label form through one of the following media. 

Please specify the publication title and series number.

Fort Collins Service Center

Telephone (970) 498-1392

FAX (970) 498-1396

E-mail rschneider@fs.fed.us

Web site http://www.fs.fed.us/rm

Mailing address Publications Distribution

 Rocky Mountain Research Station

 240 West Prospect Road

 Fort Collins, CO 80526

Rocky Mountain Research Station

Natural Resources Research Center

2150 Centre Avenue, Building A

Fort Collins, CO 80526

The use of trade or firm names in this publication is for reader information and 

does not imply endorsement by the U.S. Department of Agriculture of any 

production or service

CAUTION:

PESTICIDES

Pesticide Precautionary Statement 

This publication reports research involving pesticides. It does not 

contain recommendations for their use, nor does it imply that the uses 

iscussed here have been registered. All uses of pesticides must be registered 

by appropriate State and/or Federal agencies before they 

can be recommended. 

CAUTION: Pesticides can be injurious to humans, domestic animals, 

desirable plants, and fish or other wildlife—if they are not handled or applied 

properly. Use all pesticides selectively and carefully. Follow recommended 

practices for the disposal of surplus pesticides and pesticide containers. 
The Rocky Mountain Research Station develops scientific 

information and technology to improve management, protection, 

and use of the forests and rangelands. Research is designed to 

meet the needs of the National Forest managers, Federal and State 

agencies, public and private organizations, academic institutions, 

industry, and individuals.

Studies accelerate solutions to problems involving ecosystems, 

range, forests, water, recreation, fire, resource inventory, land 

reclamation, community sustainability, forest engineering 

technology, multiple use economics, wildlife and fish habitat, and 

forest insects and diseases. Studies are conducted cooperatively, 

and applications may be found worldwide.

Research Locations

Flagstaff, Arizona Reno, Nevada

Fort Collins, Colorado* Albuquerque, New Mexico

Boise, Idaho Rapid City, South Dakota

Moscow, Idaho Logan, Utah

Bozeman, Montana Ogden, Utah

Missoula, Montana Provo, Utah

*Station Headquarters, Natural Resources Research Center, 2150 

Centre Avenue, Building A, Fort Collins, CO 80526.

The U.S. Department of Agriculture (USDA) prohibits discrimination in 

all its programs and activities on the basis of race, color, national origin, 

age, disability, and where applicable, sex, marital status, familial status, 

parental status, religion, sexual orientation, genetic information, political 

beliefs, reprisal, or because all or part of an individual’s income is derived 

from any public assistance program. (Not all prohibited bases apply to 

all programs.) Persons with disabilities who require alternative means for 

communication of program information (Braille, large print, audiotape, etc.) 

should contact USDA’s TARGET Center at (202) 720-2600 (voice and TDD). 

To file a complaint of discrimination, write to USDA, Director, Office of 

Civil Rights, 1400 Independence Avenue, S.W., Washington, DC 20250-

9410, or call (800) 795-3272 (voice) or (202) 720-6382 (TDD). USDA is 

an equal opportunity provider and employer.

Federal Recycling Program Printed on Recycled Paper

RMRS

Rocky Mountain Research Station